Supporting conservation with biodiversity research in sub-Saharan Africa’s human-modified landscapes

Supporting conservation with biodiversity research in sub-Saharan Africa’s human-modified landscapes
Supporting conservation with biodiversity research in sub-Saharan Africa’s
human-modified landscapes
Morgan J. Trimblea & Rudi J. van Aardeb
Conservation Ecology Research Unit, Department of Zoology & Entomology, University of Pretoria, Private Bag
X20, Hatfield Pretoria 0028, South Africa
Email: a [email protected], b [email protected]
Corresponding author: Address correspondence to R.J. van Aarde, Conservation Ecology Research Unit,
Department of Zoology & Entomology, University of Pretoria, Private Bag X20, Hatfield Pretoria 0028, South
Africa, Telephone: +27 12 420-2753, Fax: +27 12 420-4523, email: [email protected]
Running title: Biodiversity in Africa‟s human-modified land
Article type: Review
Abstract
Protected areas cover 12% of terrestrial sub-Saharan Africa. However, given the inherent
inadequacies of these protected areas to cater for all species in conjunction with the effects of
climate change and human pressures on protected areas, the future of biodiversity depends
heavily on the 88% of land that is unprotected. The study of biodiversity patterns and the
processes that maintain them in human-modified landscapes can provide a valuable evidence
base to support science-based policy-making that seeks to make land outside of protected areas
as amenable as possible for biodiversity persistence. We discuss the literature on biodiversity in
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sub-Saharan Africa‟s human-modified landscapes as it relates to four broad ecosystem
categorizations (i.e. rangelands, tropical forest, the Cape Floristic Region, and the urban and
rural built environment) within which we expect similar patterns of biodiversity persistence in
relation to specific human land uses and land management actions. Available research
demonstrates the potential contribution of biodiversity conservation in human-modified
landscapes within all four ecosystem types and goes some way towards providing general
conclusions that could support policy-making. Nonetheless, conservation success in humanmodified landscapes is hampered by constraints requiring further scientific investment, e.g.
deficiencies in the available research, uncertainties regarding implementation strategies, and
difficulties of coexisting with biodiversity. However, information currently available can and
should support efforts at individual, community, provincial, national, and international levels to
support biodiversity conservation in human-modified landscapes.
Keywords: Cape Floristic Region, countryside biogeography, off-reserve conservation,
rangelands, reconciliation ecology, tropical forest
1. Introduction
Conservation of biodiversity in Africa, like elsewhere, historically focused on the fortress model,
whereby most protected areas (PAs) excluded people (see Carruthers 2009; Adams and Hulme
2001; Siurua 2006). Though PAs are essential for conservation success, they are insufficient
(Rosenzweig 2003). For example, large mammal populations have been reduced by half in some
African PAs since 1970 (Craigie et al. 2010), probably due, in part, to increasing isolation of PAs
(Newmark 2008). Weak enforcement and ineffective management plague many of Africa‟s
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current PAs (Kiringe et al. 2007; Metzger et al. 2010; Pare et al. 2010), and many fail to cater to
species with extensive spatial requirements, e.g. migratory animals (Holdo et al. 2010; Kirby et
al. 2008; Thirgood et al. 2004; Western et al. 2009) and elephants Loxodonta africana (van
Aarde and Jackson 2007). Even small-bodied species are not necessarily safe-guarded (Pauw
2007). Additionally, the configuration of PAs within the continent neglects key areas for
biodiversity (Fjeldsa and Burgess 2008; Eardley et al. 2009; Chown et al. 2003; Fjeldsa et al.
2004; Beresford et al. 2011), a problem that may escalate if climate change makes PAs
inhospitable to species they once protected (Loarie et al. 2009). If species‟ ranges shift with
shifting climate, the areas crucial for their persistence will be transient (Hole et al. 2011).
Furthermore, the scale of beta-diversity and habitat heterogeneity often extends far beyond that
of individual PAs (Gardner et al. 2007), while human activities beyond PAs influence
biodiversity within them (Hansen and DeFries 2007).
There are calls for an increased focus on biodiversity beyond African PAs (e.g. Eardley et
al. 2009) on two fronts. First, conservation of some biodiversity elements depends on how well
matrices outside of PAs cater for persistence. At the species level, for example, the Blue Crane
Anthropoides paradiseus in South Africa (McCann et al. 2007), Ethiopia‟s critically endangered
Sidamo lark Heteromirafra sidamoensis (Spottiswoode et al. 2009; Donald et al. 2010), and the
last giraffes Giraffa camelopardalis peralta in West Africa (Ciofolo 1995) all depend on humanmodified landscapes. At the ecosystem level, three biomes fall below the threshold 10%
protection status within the Afrotropic realm, i.e. tropical and subtropical dry broadleaf forests
(6%), montane grasslands and shrublands (8%), and deserts and xeric shrublands (9%), while
several ecoregions are <5% protected, especially when limited to IUCN categories I-IV, e.g.
Southern Congolian forest-savanna mosaic (0%) (Jenkins and Joppa 2009). Second, there are
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important links between biodiversity and ecosystem function, ecosystem services, and human
livelihoods in working landscapes (Daily et al. 2001; Rosenzweig 2003). For example,
maintaining natural habitat in and around farms can enhance pollination and, thus, has an
economic value to production landscapes (Carvalheiro et al. 2010; Munyuli 2012), and natural
systems in Africa provide economic and nutritional benefits to both rural and urban dwellers
(Tabuti et al. 2009; Schreckenberg 1999; Vanderpost 2006).
Though, scientists have neglected the biogeography of human-modified landscapes in
sub-Saharan Africa, ecologists are increasingly studying the capacity of such landscapes to
support biodiversity (Trimble and van Aarde 2012). Such studies are required in order for policymakers to make defensible decisions regarding land use in relation to biodiversity conservation
goals in the face of rapid economic development that could potentially decimate biodiversity.
Agriculture in Africa has been characterized by traditional, labor-intensive, smallholder
enterprise; production has been low and has remained relatively stagnant (Abate et al. 2000;
Deininger et al. 2011). However, economic development and population growth are driving
change in African landscapes, with several nations among the world‟s fastest growing economies
(IMF 2013). In 2009, the population reached one billion and is predicted to double by 2050 (UNHABITAT 2010). Urbanization is a strong force; 40% of the current African population is citydwelling, and by 2050, 60% will be urban (UN-HABITAT 2010). Even so, the rural population
will also grow substantially, predicted to increase by nearly 50% by mid-century (UN Population
Division 2012), while growing urban centers will depend heavily on rural resources. To meet this
demand and to improve food security, there are calls for both intensifying smallholder
agriculture (Baiphethi and Jacobs 2009; Snapp et al. 2010; Muriuki et al. 2005; Baudron et al.
2011) and extensifying production (Muriuki et al. 2005).
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Therefore, the interest in biodiversity in human-modified lands is timely. Although
Africa‟s natural ecosystems are more intact than many other regions‟, a proactive approach to
biodiversity conservation that strives for the most prudent management of the unprotected
matrices between PAs is clearly preferable to trying to reconnect and restore already degraded
ecosystems (Gardner et al. 2010). Thus, as policy-makers chart the future course of development
in Africa, they should consider the effects of different choices on biodiversity in human-modified
lands, what steps can be taken to prevent biodiversity loss, and the benefits and costs of
biodiversity persistence to people. Failure to do so may necessitate expensive restoration and
reintroduction efforts. Studies of biodiversity patterns and the processes that maintain them in
human-modified landscapes provide an evidence base to support defensible management that
meets the needs of people and biodiversity simultaneously. The evidence base should,
furthermore, provide for relevant ecological contexts. For example, management standards for
timber plantations aim to minimize impact on biodiversity in surrounding natural forests. Yet,
the same standards have been applied in plantations embedded in grasslands with dubious
efficacy for minimizing impacts on grassland biodiversity (Lipsey and Hockey 2010; Pryke and
Samways 2003).
This scientific focus on biodiversity in human-modified landscapes is distinct from
Africa‟s thirty-some-year experiment in community-based conservation (CBC, but also known
as Integrated Conservation and Development Projects, Community-Based Natural Resource
Management, and others), but these two fields can and should be amalgamated. Promoters of
CBC claim that it increases the chance of conservation success and simultaneously reduces rural
poverty by allowing community involvement in management and profit from natural resources,
especially large mammals (see Hackel 1999). The philosophy of linking wildlife conservation
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and rural economic development and the practical successes and failures therein have been
discussed in a large body of literature (e.g. Songorwa et al. 2000; Torquebiau and Taylor 2009;
Hackel 1999) and are subject of renewed controversy outside of Africa (see Kareiva and Marvier
2012; Doak et al. 2013). However, the discussion has focused on socioeconomics and politics
with fleeting consideration for assessing actual biodiversity persistence under different CBC
models, a problem pointed out by Caro (1999) and subsequently largely ignored.
In this review, our objective was to elucidate the state of knowledge regarding
biodiversity in sub-Saharan Africa‟s human-modified landscapes to provide an overview for
scientists and policy-makers. We also sought to identify constraints and opportunities for
integrating biodiversity conservation with land-use management that require further attention in
order to support the implementation of sensible policies.
2. Methods
2.1. Literature search
We searched the ISI Web of Knowledge (up to 2012) with keywords “Africa” and “biodiversity
or conservation” and each of the following terms: “agricultur*”, “agroforestry”, “communal”,
“farm*”, “game farm”, “game ranch”, “human-modified”, “multiple-use management”,
“peri$urban”, “private nature reserve”, “range$land”, “rural”, “suburban”, and “urban”. We also
searched for the terms “countryside biography”, “reconciliation ecology”, and “off-reserve
conservation”. Additionally, we included relevant papers found coincidentally or in reference
lists.
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3. Biodiversity in human-modified landscapes of African ecosystems
In summarizing the literature on biodiversity in Africa‟s human-modified landscapes, we
separate our discussion into four major ecosystem types (see Fig. 1) within which we expect
similar patterns to emerge. 1) Rangelands attract the bulk of our attention as Africa‟s biggest
ecosystem type, and rangeland biodiversity is perhaps the most compatible with human landuses, so biodiversity-conscious land-use planning in rangelands could yield huge benefits. 2)
Tropical forests are discussed briefly with a focus on Central and East African forests, and we
refer readers to an excellent review of the abundant literature from West Africa (Norris et al.
2010). 3) The Cape Floristic Region, though small, is extremely rich in species yet threatened by
extensive commercial development, and we discuss a growing body of literature on land-use
management in the region. Finally, 4) the urban and rural built environment will become an
increasingly important concern for biodiversity conservation in Africa where the increase of
urban land cover is predicted to be the highest in the world at nearly 600% in the first three
decades of the 21st century (Seto et al. 2012); proper management and infrastructure
development could attenuate the consequences for biodiversity.
3.1. Rangelands
Two-thirds of sub-Saharan Africa is composed of rangelands (Fig. 1), consisting of arid and
semi-arid grasslands, woodlands, savannas, shrublands, and deserts. The rural people inhabiting
rangelands are typically agropastoralists, specializing in small-scale farming or livestock keeping
or a combination. Some agricultural practices in rangelands may be harmful to biodiversity, e.g.
overcultivation, overgrazing (Kerley et al. 1995), bush fires, cultivation of marginal and easily
eroded land, and widespread use of chemicals and pesticides (Darkoh 2003). Many people in
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rangelands depend heavily on wild resources, e.g. via hunting and gathering or by profiting from
wildlife tourism (Homewood 2004). Game ranching is an increasingly popular land-use option
across African rangelands (McGranahan 2008), and so are “eco-estates” (Grey-Ross et al.
2009b), where people choose to live amongst the natural beauty of African rangelands and their
considerable species diversity, especially charismatic large mammals.
The ecological mechanisms that maintain different rangeland types in different locations,
e.g. grassland versus woodland, are not fully understood though interactions between soils,
climate, fire, herbivory, and human disturbance are thought to be important (see Bond and Parr
2010) . The biggest threats to grasslands include afforestation or bush encroachment and clearing
for agriculture (Bond and Parr 2010), while threats to the woodlands include woodcutting,
clearing for agriculture, and over-use (Tabuti 2007; Schreckenberg 1999). Many perceive that
biodiversity is declining in rangeland systems; they blame poor agricultural practices, land
conversion, and over-utilization of wild resources by rural people and worry that these patterns
will increase with population growth (e.g. Darkoh 2003; Thiollay 2006). However, documented
evidence of biodiversity loss in rural rangelands is sparse. Many areas have likely lost some
species, but surprisingly, long-inhabited regions lacking formal PAs, e.g. Kenya‟s Laikipia
district, maintain abundant wildlife including large carnivores and elephants (Gadd 2005;
Kinnaird and O'Brien 2012) that might seem at odds with human occupation (Woodroffe et al.
2007). Rangeland systems are characterized by disturbances such as fire, unpredictable rainfall,
grazing and browsing pressure, and physical disturbance. Therefore, rangeland biodiversity may
be relatively resilient to anthropogenic disturbance due to the ability to disperse, colonize, and
persist in patchy, fluctuating environments (Homewood 2004). Thus, human-modified
landscapes have the potential to maintain a relatively large portion of rangeland biodiversity.
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Nonetheless, conservation in rangelands has traditionally excluded people from
designated PAs. In South Africa, for example, conservation planning often dichotomizes „human
land-use‟ and „conservation‟ with little consideration for different land-use options that may be
variably amenable to biodiversity (e.g. Wessels et al. 2003; Chown et al. 2003). On the other
hand, some authors have called to “mainstream” conservation into human-modified lands (e.g.
Soderstrom et al. 2003; Pote et al. 2006). O‟Connor and Kuyler (2009) used expert opinion to
rank the impact of land uses in moist grasslands on overall biodiversity integrity (from least to
most impact: conservation, game farming, livestock, tourism, crops, rural, dairy, timber, and
urban). Empirical studies are amassing to assess such assertions, which could support land-use
planning for conservation. Here we discuss emerging research on biodiversity in several of the
most common rangeland land uses.
3.1.1. Grazing
Grazing is important to the maintenance of grassland and savanna habitats, economic
development, and management for biodiversity. However, plant responses to grazing are
idiosyncratic and incompletely understood (see Watkinson and Ormerod 2001; Rutherford et al.
2012). Overgrazing can lead to degradation and bush encroachment (the proliferation of woody
plants at the expense of grasses), while too little grazing can result in succession to woodland
(Watkinson and Ormerod 2001). Of course, grazing effects on vegetation can affect higher
trophic levels as well, so it is important to understand vegetation responses to grazing, not only
for livestock production, but also because vegetation dynamics affect many other species.
However, not all grazing landscapes are alike; unique vegetation dynamics in different
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ecosystems mean that landscapes can respond disparately to grazing pressure (Todd and
Hoffman 2009).
Table 1. General Conclusions regarding practices that support biodiversity in human-modified landscapes and
scientific and implementation concerns requiring further investigation
Practices that tend to support species diversity and richness in human-modified landscapes
Prefer diversity in selection of crops grown (i.e. polyculture) and land use (i.e. land-use mosaics, over homogenous
monocultures)
Encourage traditional agricultural practices over large-scale, mechanized farming
Leave as much remnant natural vegetation as possible and monitor or assist maintenance of keystone structures or
species, e.g. large trees
Ensure strict protection for specialist and endemic species and expand PA coverage focused on these groups
Encourage appreciation and understanding of conservation goals among land users
Discourage urban sprawl and maintain and manage urban green spaces
Favor use of native species in gardens and cultivation
Avenues for further investigation into scientific uncertainties and implementation practices
Researching poorly documented combinations of species group, ecosystem type, and land use, e.g. mammals in
rangeland agroforests
Moving beyond occurrence data to likelihood of persistence, e.g. how dependent are species in human modified
landscapes on nearby PAs or remnant habitat?
Investigating the value of reintroduction or rewilding in human-modified landscapes
Going beyond the species level of biodiversity to integrate genetic and ecosystem concepts
Supporting technological and traditional knowledge for cultivating useful native species and investigating the effects
of these practices on other taxa
Creating frameworks for valuing importance of different species in different landscapes at a local level within a
global context, e.g. commonness versus rarity and specialist versus generalist
Developing policies that integrate and account for local, regional, and global conservation needs and land use
systems
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Research is emerging that investigates aspects of grazing management and biodiversity in Africa;
we summarize 30 such studies in Online Resource Table 1. Generally, these studies look at
grazing intensity, or proxies such as bush encroachment, and show that many wild species may
be maintained depending on management and location. For example, traditional pastoral
practices, i.e. burning and boma creation, may even be necessary to maintain avian diversity in
some East African savanna areas (Gregory et al. 2010). Contrastingly, bush encroachment due to
overgrazing in Ethiopia may provoke Africa‟s first avian extinction (Donald et al. 2010;
Spottiswoode et al. 2009).
Online Resource Table 1 shows that only about a third of studies compared biodiversity
of livestock grazing landscapes to controls with indigenous grazers such as PAs. Most studies
came from South Africa (67%), and most assessed grazing effects on plants (43%) or insects
(27%). Many areas of investigation remain open, such as the role of vegetation structure
including keystone, isolated trees in maintaining biodiversity in human land-use areas; such trees
are important for maintaining diversity in natural systems (Dean et al. 1999). A common
conclusion with regards to plant diversity is that spatial heterogeneity in grazing management
that includes PAs will enhance gamma diversity because different species thrive at different
grazing intensities (e.g. Fabricius et al. 2003).
3.1.2. Agricultural mosaic
While extensive grazing is common in arid-savannas and xeric shrublands, a land-use mosaic of
grazing and cropping interspersed with settlements is common in more mesic savannas and
grasslands. This mosaic effect may have important consequences for the maintenance of
biodiversity, and studies of biodiversity in agricultural mosaics (24 studies summarized in Online
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Resource Table 2) identify some common themes. Compared to strict PAs, agricultural mosaics
may actually be beneficial to some species groups. For example, Caro (2001) illustrated greater
diversity and abundance of the small mammal assemblage in the agricultural matrix outside
Katavi National Park, Tanzania than inside, a pattern also found for Niokolo Koba National
Park, Senegal (Konecny et al. 2010). Richness of birds, amphibians, small mammals, butterflies,
and trees is similar at 41 sites across a land-use gradient from Katavi National Park to nonintensive agricultural land; however, composition changes along the gradient, and although the
PA holds some unique species, some species found outside the PA are absent within (Gardner et
al. 2007). Thus, agricultural mosaics may contribute to greater gamma diversity at the landscape
scale; nonetheless understanding the conservation implications of higher gamma diversity may
require a regional or global perspective on species rarity and commonness.
It is a common finding that agricultural intensification (e.g. mechanization, shortening
fallows, destruction of remnant habitat patches, and introduction of cash crops) can have
detrimental effects on the biodiversity value of agricultural mosaics. The mosaic effect of
traditionally managed farms in KwaZulu-Natal, South Africa may support, and even enhance,
bird diversity (Ratcliffe and Crowe 2001), but intensification results in species declines due to
loss of „edge‟ habitats. In Burkina Faso, common butterfly species occur in cultivated areas,
while specialists are more common in old fallows and grazed areas, probably because grazing
maintains host plants and, thus, diversity (Gardiner et al. 2005). In this case, an agricultural
mosaic of shifting fallows could support butterfly meta-populations that allow species
persistence, while intensification could be detrimental (Gardiner et al. 2005). In Ethiopian
grasslands, low-intensity agriculture supports moderate plant diversity, while larger-scale,
mechanized farms reduce tree cover and diversity (Reid et al. 1997). Similarly, In the Serengeti12
Mara ecosystem, commercial mechanized agriculture is associated with declining wildlife
populations (Homewood et al. 2001; Homewood 2004).
3.1.3. Cropping
Cropping is perhaps more at odds with biodiversity than grazing is because cropping involves the
direct removal of indigenous vegetation and planting of, generally, non-indigenous species.
Nonetheless, crops can support wild species, and conservation value may depend on the crops
planted, the farming methods employed, and the arrangement of fields with respect to natural
habitat. Relatively few studies assessed biodiversity solely in cultivated areas (10 studies
summarized in Online Resource Table 3), as opposed to agricultural mosaics (Online Resource
Table 2). This perhaps reflects the current state of African agriculture, where most farms are
smallholder or subsistence based rather than expansive, commercial cultivation. Although there
are exceptions, average farm size is just 2 to 3 ha (Deininger et al. 2011). Where commercial
cultivation does occur, loss of biodiversity may be seen as a foregone conclusion not worth
investigating (see Thiollay 2006). Many studies of biodiversity in cultivation were concerned
primarily with the benefits of that diversity for production via pest control, fertility enhancement,
or pollination services, rather than for its value to conservation (e.g. Carvalheiro et al. 2010;
Midega et al. 2008; Tchabi et al. 2008).
3.1.4. Agroforestry
Agroforestry, the integration of trees into agriculturally productive landscapes, has garnered
much attention in the global conservation community because it has been shown to provide
habitat for relatively high levels of forest species diversity (see Bhagwat et al. 2008). In African
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rangelands, agroforestry can be divided into two types: technological and traditional.
Technological agroforestry deals with the expertise to plant and maintain tree species that
increase productivity in agricultural production systems. Kenyan farmers, for example, plant
crops of fodder trees, which raise milk yields of cows and goats (Pye-Smith 2010a). Government
programs in Niger, Zambia, Malawi, and Burkina Faso support large-scale „evergreen
agriculture‟ projects to plant indigenous trees such as Faidherbia albida among crops, which
maintain green cover year-round, increase yields by improving soil fertility, and provide fodder
and firewood (Garrity et al. 2010). Evergreen agriculture and other technological agroforestry
projects are touted as having greater biodiversity value than do monoculture crops (see PyeSmith 2010a, b; Kalaba et al. 2010; Garrity et al. 2010). Yet, evidence to support these claims
remains mostly anecdotal, warranting further research because plans are underway to expand
technological agroforestry projects throughout Africa (Garrity et al. 2010).
Traditional agroforestry, on the other hand, is a millennia-old practice, particularly
evident in the parkland savannas of West Africa, of people maintaining savanna tree species in
pastures, fields, and villages. These trees provide shade, food, wood, and even cash when
commercially traded (e.g. shea, baobab), and traditional agroforestry may contribute to the
maintenance of tree species in addition to species for which trees provide habitat. Many studies
have enumerated tree diversity in farmlands (Online Resource Table 4). Even so, the
conservation value of agroforestry varies. Augusseau et al. (2006) report that in Burkina Faso,
few indigenous species are important to farmers and none are planted. Even where tree richness
is maintained at a relatively high level, the persistence of trees in traditional agroforestry can be
compromised if the economic value of totally clearing the land, e.g. for mechanized agriculture
or firewood, outpaces the value of non-timber products (Tabuti et al. 2009). Additionally, based
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on demographic profiles of tree species, tree regeneration appears problematic in many humanmodified landscapes (e.g. Fandohan et al. 2010; Schumann et al. 2010; Venter and Witkowski
2010). For example, a study in Benin shows that the largest shea trees are often in villages or
fields, but seedling survival is low compared to nearby PAs (Djossa et al. 2008). Regeneration
potential can also be diminished when harvesting tree products affects recruitment, as is the case
for Khaya senegalensis in Benin (Gaoue and Ticktin 2008). Where natural regeneration potential
is compromised, intervention may be required to ensure rejuvenation (Kindt et al. 2008;
Ouinsavi and Sokpon 2008), especially if farmers abandon traditional rotational land-use systems
such as long fallow, where trees are often most capable of regenerating (Raebild et al. 2007;
Schreckenberg 1999).
Fortunately, agroforestry management in rangeland ecosystems is an active area of
research with regards to developing strategies to encourage tree persistence (Kindt et al. 2008;
Tabuti et al. 2009; Augusseau et al. 2006). Yet, there is a surprising lack of research to assess the
value of savanna agroforests for faunal diversity or even non-tree plant diversity (Online
Resource Table 4), aspects that have been more thoroughly studied in the tropical forest context
(Bhagwat et al. 2008), and this dearth should be remedied.
3.1.5. Game ranching and private nature reserves
The wildlife industry, including game ranching, game farming, and private nature reserves, has
become big business, especially in southern and East African rangelands. These land-use options
involve profiting from consumptive (e.g. trophy hunting, live animal sales, meat) or nonconsumptive (e.g. tourism, aesthetic value) use of wildlife on communal or private land. South
Africa alone has an estimated 9,000 private game ranches, covering 20.5 million ha, many of
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which were converted from traditional livestock ranches (NAMC 2006). Ranching game rather
than domestic livestock may ameliorate effects of overgrazing because indigenous species have
coevolved with indigenous vegetation (Kerley et al. 1995), and indigenous browsers may control
bush encroachment (McGranahan 2008; Taylor and Walker 1978). Thus, the wildlife industry
may be a boon to biodiversity conservation; however, very few studies have actually assessed
impacts on biodiversity, which may be positive or negative and likely depend on management
actions (Cousins et al. 2008).
Occurrence and abundance of mammal species on private land has increased due to game
ranching (Lindsey et al. 2009). Nonetheless, some aspects of the wildlife industry are worrying.
Privatization of wildlife (and sometimes legislative requirements) begets ubiquitous game
fencing (McGranahan 2008; Lindsey et al. 2009) with substantial ecological consequences
including the interruption of natural movements, inbreeding, and overstocking (Lindsey et al.
2009; Hayward and Kerley 2009). Ranches are often quite small (South African provincial
averages range from 8.2 to 49.2 km2), and smaller ranches necessitate more intensive
management interventions (Lindsey et al. 2009; Bothma 2002). Additionally, the industry‟s
focus on trophies may skew natural communities in favor of valuable species and induce semidomestication (Mysterud 2010), and it has resulted in extra-limital introductions, questionable
breeding practices, and persecution of predators (Lindsey et al. 2009). Even within the mammal
community, generally the focus of game ranching, the full complement of species of a given
ecosystem may not be maintained on ranches despite deliberate re-introductions (Grey-Ross et
al. 2009a).
Thus, much more research is needed on the biodiversity value of the wildlife industry and
what measures, e.g. promoting conservancies over single game ranches (Lindsey et al. 2009), can
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improve this value. Best-practice management in terms of grazing pressure, fire regimes, bush
encroachment, wildlife ownership policies, and fencing needs more attention (McGranahan
2008). Furthermore, surprisingly little is known about the impacts of game ranching on species
other than large mammals. Even so, game ranches are likely more amenable to most indigenous
biodiversity than are many other commercial land-use options. For example, large eagles in
South Africa‟s Karoo shrublands are much more common in areas stocking indigenous mammals
than in areas with domestic livestock and cultivation (Machange et al. 2005).
3.2. Tropical forests
Though rangelands cover the majority of Africa, tropical forests also make up a considerable
portion (~20% (Brink and Eva 2009)) (Fig. 1), particularly rich in biodiversity. Research on
biodiversity in human-modified landscapes is biased towards tropical forests (Trimble and van
Aarde 2012). Nonetheless, biodiversity in human-modified tropical forest landscapes in Africa
has received much less scientific attention than in other regions, especially South and Central
America (Gardner et al. 2010). African tropical forests tend to be in less conflict with high
human population densities than elsewhere (e.g. Southeast Asia and Brazilian Atlantic forests)
(Gardner et al. 2010), although in West Africa 80% of the original forest extent is now an
agricultural-forest mosaic home to 200 million people (Norris et al. 2010).
We do not attempt a comprehensive review of African tropical forest biodiversity in
human-modified landscapes and refer readers to Norris et al. (2010) for an excellent treatment of
the West African scenario. They lament the lack of data regarding biodiversity in African
agricultural-forest mosaics but are able to reach some general conclusions. Land uses that
maintain tree cover are more amendable to forest biodiversity than those that do not. Species
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Figure 1. Map of sub-Saharan Africa showing ecosystem types adapted from Olson et al. (2001): rangelands (desert
and xeric shrubland, montane grassland/shrubland, flooded grassland/savanna, and tropical/subtropical
grassland/savanna/shrubland), tropical forests (moist and dry tropical forest), the Cape Floristic Region
(Mediterranean forest/woodland/scrub), and the urban and rural built environment represented by the human
influence index (Wildlife Conservation Society and Center for International Earth Science Information Network
2005), a dataset comprising nine data layers incorporating population pressure (population density), human land use
and infrastructure (built-up areas, nighttime lights, land use, land cover), and human access (coastlines, roads,
railroads, navigable rivers).
richness increases in some modified habitats, such as logged and secondary forest, for some
species groups, but endemic forest species are often lost. Additionally, relatively high species
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richness in modified habitats comprises, in part, species not present in the baseline forest
comparison, so species richness alone likely overestimates the value of modified habitats for
forest species. Furthermore, habitat modification seems to affect richness of forest plant species
more negatively than of some animal groups.
Although logically, it seems more difficult to encourage the persistence of biodiversity in
human-modified landscapes embedded in tropical forests than in rangelands, research can
indicate best practices for land-use planning. In contrast to West Africa, Central Africa still
maintains large tracts of relatively undisturbed forest that are becoming increasingly threatened
by development, and lessons from studying African forest biodiversity in human-modified
landscapes should be incorporated into development policy for the region (Norris et al. 2010).
The tropical forest biome extends to East and southern Africa where forests are less
extensive, confined largely to high altitudes inland and a linear belt along the coast. These
geographic constraints present unique challenges for conservation and heighten the importance
of maintaining endemic species and retaining connectivity in fragmented forests. Fewer studies
consider East and southern African tropical forests than West African forests, but work is
emerging to support land-use planning in the region, and results largely conform to those found
for West Africa. Agroforestry in Ethiopian and Tanzania supports less diversity than forests but
more than other land uses (Gove et al. 2008; Hemp 2006; Hall et al. 2011; Negash et al. 2012).
While Schmitt et al. (2010) found higher overall plant richness in Ethiopian coffee agroforests
than natural forests, richness of typical forest species was lower. In Kenya, connectivity of
coastal forest fragments for primates may be influenced by matrix structure (Anderson et al.
2007). Farmland outside tropical forest remnants, especially structurally complex subsistence
farms, support higher bird richness than forests; however, many forest species are lost,
19
highlighting the importance of maintaining the forest remnants but also supporting traditional
farming techniques over commercial monocultures (Laube et al. 2008; Mulwa et al. 2012).
Furthermore, structurally diverse farmland surrounding forest remnants may enhance forest
pollinator communities (Hagen and Kraemer 2010). Similarly, South African forest remnants
embedded in various matrix types have similar bird species richness, but abundance is highest in
fragments in agricultural matrices due to the presence of forest generalists and open-habitat
species, while forest specialists are rare (Neuschulz et al. 2011). Additionally, herpetofaunal
richness does not decline monotonically along a land-use gradient from forest to cultivation,
while richness of functional groups erodes along the gradient due to sensitivity of some specialist
groups (Trimble and van Aarde 2014). Forest fragments and grasslands in the agricultural
mosaic outside a PA in southern Mozambique have more beetle species and higher abundance,
while endemic beetle species are better represented inside the PA (Jacobs et al. 2010).
3.3. Cape Floristic Region
While small in area (approximately 90,000 km2, see Fig. 1), the Cape Floristic Region (CFR) of
South Africa is a biodiversity hotspot of global significance (Myers et al. 2000) consisting of a
Mediterranean-type ecosystem with high species turnover across the landscape and high
endemicity. Systematic conservation planning has been conducted for the region but focused on
pristine habitat that could be formally protected (see Cowling and Pressey 2003). Because spatial
turnover of species is so high, however, successful conservation will depend heavily on efforts in
human-modified landscapes beyond PAs (Cox and Underwood 2011). Based on species-area
curves for plants and vertebrates in the CFR, practicing biodiversity friendly management on just
20
25% of the land that is beyond PAs, but still in a natural or semi-natural state, might add an
additional 541 species to the 7,340 estimated to occur in PAs (Cox and Underwood 2011).
However, in contrast to many areas of Africa dominated by subsistence agriculture, the
CFR is characterized by large areas of intensively managed agricultural monocultures with low
biodiversity value (Giliomee 2006). Overall, only 26% of the CFR has been transformed, but the
CFR is made up of different habitat types, and some, especially in the fertile lowlands, have lost
much more of their area to cultivation, urbanization, and heavy invasion of exotic plants; for
example, coast renosterveld is more than 80% transformed (Rouget et al. 2003a).
Transformation threatens not only the CFR‟s plants but also endemic and vulnerable animals
such as the Black Harrier Circus Maurus, which has been displaced from the inland plains by
cereal agriculture and now breeds, less successfully, in the coastal strip and inland mountain
habitats (Curtis et al. 2004). Though the Black Harrier can forage in cultivated areas, it relies on
intact vegetation to breed (Curtis et al. 2004).
PAs within the CFR are concentrated in areas of low agricultural value (e.g. mountains
and coastlines), so biodiversity in fertile areas depends on conservation on privately owned land
(Giliomee 2006; Rouget et al. 2003b). To increase the biodiversity value of agricultural areas,
the primary focus should be on conserving remnants of natural vegetation on farms (Giliomee
2006). This is being attempted through incentive-driven stewardship agreements that protected
almost 70,000 ha of vegetation on private land between 2003 and 2007 (Von Hase et al. 2010).
Additionally, farm management practices may be variably amenable to biodiversity. For
example, though vineyards have very different arthropod communities than those in natural
vegetation, organic vineyards support greater diversity than do more intensively managed
vineyards (Gaigher and Samways 2010). However, these effects may be taxon dependent; for
21
instance, organic vineyard management benefits richness of monkey beetles (crucial pollinators),
but not bees (Kehinde and Samways 2012). Similarly, apple orchards support less arthropod
diversity than natural vegetation does, but orchards that are not sprayed with pesticides have a
higher diversity than sprayed sites (Witt and Samways 2004). On the other hand, farms with a
mixture of different crops and remnants of natural vegetation maintain most fynbos bird species
and attract several additional species, while single crop sites without remnant vegetation have
much less bird diversity and lose many fynbos species (Mangnall and Crowe 2003). Clearly,
maintaining remnant vegetation and connectivity in agricultural areas of the CFR is crucial, but
more research is needed to tailor agricultural practices to better conserve CFR species in
production landscapes.
3.4. Urban and rural built environment
Plant and vertebrate species richness and endemism are correlated with human population
density and human infrastructure in sub-Saharan Africa (Burgess et al. 2007; Fjeldsa and
Burgess 2008; Balmford et al. 2001), which is substantial in many regions (see Fig. 1). That the
pattern endures in relatively developed South Africa means either that species persist to some
degree with humans in disturbed habitats at current levels, that human-disturbed habitats actually
attract more species, or that a major extinction debt is yet to be paid (Fairbanks 2004; Chown et
al. 2003). Regardless, areas with high human density (predicted to increase dramatically in
Africa, outpacing growth in all other regions in the coming decades (Seto et al. 2012)) require
appropriate regulations to ensure they remain as amenable as possible to biodiversity
conservation. This will be especially important in Africa‟s most biologically rich yet rapidly
urbanizing regions; by 2030 for example, the urban area within the Eastern Afromontane and
22
Guinean Forests of West Africa hotspots is forecasted to be 1,900% and 920% of 2000 levels
respectively (Seto et al. 2012).
Some obvious steps include discouraging urban sprawl; providing appropriate housing
for low income populations while controlling illegal settlements in biodiversity sensitive areas;
designing relevant green spaces that include aquatic habitats and indigenous plants; and
managing invasive species, waste, and pollutants (Puppim de Oliveira et al. 2011; Muriuki et al.
2011). Research on managing Africa‟s urban and rural built environments for biodiversity is in
its infancy and is mostly constrained to South Africa. Clearly, more research is needed, yet
several studies provide pertinent information for planners.
While urban environments might not seem particularly hospitable to biodiversity, even
small home gardens in African cities can harbor a remarkable number of species, especially in
the tropics, both intentionally cultivated and otherwise (Cumming and Wesolowska 2004; Lubbe
et al. 2010; Bigirimana et al. 2012). In South Africa, socioeconomics, urbanicity, and ecological
factors influence plant diversity and the proportion of invasive species in home gardens (Lubbe
et al. 2010; Molebatsi et al. 2010). Gardens with a high number of non-indigenous species
contribute to biotic homogenization and pose the risk of new introductions that could prove
detrimental to indigenous ecosystems. Therefore, invasive species in the urban landscape need to
be controlled through regulation and removal, especially in threatened and fragile ecosystems
(Alston and Richardson 2006; Dures and Cumming 2010; Cilliers et al. 2008; Bigirimana et al.
2012).
Green spaces such as city parks, tree-lined streets, and even golf courses in urban
environments can support certain species. Dures and Cumming (2010) show that habitat quality,
rather than patch metrics such as area, have the greatest influence on bird diversity in fynbos in
23
an urban gradient in Cape Town. Thus, controlling invasive species even in high-density housing
areas may be more beneficial for birds than expanding the low quality network of urban reserves.
Alien pine tree removal helps restore invertebrate species diversity in Cape Town, and
fragments of natural vegetation and gardens with indigenous plants help maintain it (Pryke and
Samways 2009). In the Durban Metropolitan Open Space System, complex habitats (i.e. with
trees and shrubs) support higher invertebrate diversity than simplified habitats (i.e. mown lawns);
however, simple habitats might cater for certain rare species (Whitmore et al. 2002). Green
spaces in urban Pretoria contribute to butterfly and moth diversity (McGeoch and Chown 1997)
and also support indigenous birds (van Rensburg et al. 2009), while maintaining urban riparian
vegetation is necessary for dragonfly conservation in Pietermaritzburg (Samways and Steytler
1996). Better ecological planning for developments such as golf courses could increase the
likelihood for biodiversity persistence and minimize negative consequences, even in the CFR
(Fox and Hockey 2007). Additionally, habitat engineering, e.g. creating biotopes for dragonflies
(Steytler and Samways 1995), might be a useful tool in the urban context to promote
biodiversity, although continual management of these habitats may be necessary to ensure
species persistence (Suh and Samways 2005).
When species are range-restricted such that a single metropolitan area may affect most of
their range, special attention is required. For example, two small forest parks in Durban suburbs
are home to the last remnant populations of the rare tree Oxyanthus pyriformis whose specialist
pollinators, the long-tongued hawkmoths, appear unable to tolerate suburban living. Hand
pollination and planting of seedlings will be necessary to maintain the species (Johnson et al.
2004). Similarly, conservation of plants in Cape Town is hampered by apparent sensitivity of
specialist pollinator birds to urbanization, which is concerning given the increasing urbanization
24
in the CFR (Seto et al. 2012). Durban covers a large portion of the range of the black-headed
dwarf chameleon Bradypodion melanocephalum, and translocations from sites demarcated for
development to sites reserved for conservation have proven somewhat successful, dependent on
adequate alien-plant-control and habitat restoration (Armstrong 2008). Unique landscape features
within urban areas may also require special attention. For example, Table Mountain in Cape
Town harbors endemic species whose conservation depends not only on the PA of Table
Mountain but also on management of lower elevation suburban woodlands (Pryke and Samways
2010).
On the rural end of the settlement spectrum, less attention has been given to biodiversity
persistence. Some agricultural mosaic studies consider rural settlements, but a few studies treat it
explicitly. For example, similar to shifting cultivation, some cultures practice shifting settlement,
and abandoned settlements have been shown to provide valuable seasonal resources, e.g. fruit
trees, to chimpanzees Pan troglodytes in Mali (Duvall 2008). Even road verges may provide for
some types of biodiversity. For example, verges in the Karoo support some plant species not
found in adjacent grazing lands, though many species from pastures are not found in verges
(O'Farrell and Milton 2006). Verges also support invertebrates and could prove valuable to
conservation because verges are public spaces that can be managed for biodiversity (Tshiguvho
et al. 1999).
Understanding more about urban settlement and biodiversity may even benefit
conservation in once remote PAs where rural sprawl and infrastructure for wildlife tourism can
be dramatic (Wittemyer et al. 2008). For example, recent decades have seen substantial increases
in rural sprawl along with the construction of 60 tourist lodges, 1,200 boreholes, and 540 km of
roads in the Okavango Delta, one of Botswana‟s premiere conservation areas (Vanderpost 2006).
25
4. Constraints and opportunities
The scope of this review was broad, covering four major ecosystem types within sub-Saharan
Africa. Specific conclusions are difficult because species‟ responses to land use are clearly
idiosyncratic and dependent on local factors. Nonetheless, we have identified some general
conclusions and constraints summarized in Table 1 and discussed below.
4.1. The science of biodiversity in human-modified landscapes
As others have pointed out, understanding the value of human-modified landscapes for
biodiversity, especially in Africa, is hampered by data constraints (Norris et al. 2010; Pettorelli et
al. 2010; Waltert et al. 2011; Trimble and van Aarde 2012). Many studies are limited in temporal
and spatial scale, and poor study design may result in insufficient sampling of habitats. The focus
on species richness of certain habitat types while failing to account for the importance of species
from other habitats in assigning conservation value to different land-use options may neglect the
bigger picture; Bond and Parr (2010), for example, call for more collaboration between forest
conservationists and others. More consideration for the value of different species in terms of
commonness and rarity also needs to be developed because human-modified landscapes often
fail to cater for endemic and specialist species (Waltert et al. 2011), and a better understanding of
beta and gamma diversity at a landscape scale is necessary.
Additionally, further investigation into the relationship between occurrence and
persistence is required, as are more studies that delve beyond species richness into the processes
that support the observed patterns of biodiversity. For example, studies of demographic
processes (e.g. Venter and Witkowski 2010; Djossa et al. 2008; Schumann et al. 2010) and
population trends (e.g. Trimble and van Aarde 2011; Stoner et al. 2007) for species inhabiting
26
human-modified landscapes can provide insight beyond mere patterns of occurrence.
Furthermore, as elsewhere (Gardner et al. 2010), studies of biodiversity in African humanmodified landscapes is biased towards certain taxa—and the patterns exhibited by these species
might not apply to others (Caro 2001). Also, genetic diversity, has not generally been considered
though may be important in terms of traits valuable to humans and for conservation (Ashley et
al. 2006). Conservation in human-modified landscapes may be particularly important in
conserving genetic diversity because the traditional fortress PA model may encompass relatively
little, especially for plants (Atta-Krah et al. 2004).
Many authors lament erosion of ecological knowledge to maintain species, especially
trees, medicinal plants, and wild food plants, and urge more effort towards domestication,
cultivation, and marketing to provide farmers with the means to conserve species while easing
pressure on wild stock and improving food security and economic stability (Kindt et al. 2008;
Ntupanyama et al. 2008; Tabuti et al. 2009; Leakey and Tchoundjeu 2001; Dold and Cocks
2002; Dovie et al. 2007; Khumalo et al. 2012). However, care must be taken to ensure that
genetic diversity is maintained in the process (Lengkeek et al. 2006; Muchugi et al. 2008).
Development of domestication and cultivation methods could promote the use of native species
in human-dominated lands, and these native plants may contribute to conservation of other taxa
(Dovie et al. 2007), but more research is clearly required.
4.2. Implementing policies
Given the limitations of the available science, it is difficult to develop strategies to encourage
land uses that are of the highest conservation value. The effect of policy on biodiversity
conservation in human-modified landscapes under different land tenure systems and different
27
settlement patterns needs more research because decisions are largely opinion driven and not
evidence based (Homewood 2004; Duvall 2008). Perhaps the community-based conservation
literature, which has focused heavily on implementation and policy, could lend some insight. A
review of this literature stresses that better implementation results are achieved when there is
quality governance, resilient local institutions with local power and accountability, consideration
for local context, integration across social and ecological systems, and mutual learning involving
communities and other involved parties, e.g. outside experts (Balint and Mashinya 2008). NGO‟s
and foreign aid are more likely to encourage successful conservation when projects are flexible,
small-scale, and targeted at local interests, and when they prioritize innovation, learning, and
experimentation (Nelson 2009). Conservationists must also take cognizance of perspectives and
needs of local communities in both rural and urban settings in order to better engage them in
conservation management (Ferketic et al. 2010). CBC projects that are independent of PAs are
excellent opportunities to maintain biodiversity on human-modified land of marginal use for
agriculture; and expert opinion, monitoring, and ecological modeling tools can help communities
manage their natural resources (Du Toit 2002).
We have indicated several gaps in the literature on biodiversity in African humanmodified landscapes, and while much more work is required to create sensible policies that meet
conservation needs and those of governments and people (Ashley et al. 2006), as it stands,
current research can go some way towards supporting policy-making. Studies of biodiversity
persistence in different land-use options for a given region can be incorporated into scenario
modeling for future development. For example, Turpie et al. (2007) amalgamated studies of
plants, invertebrates, birds, and mammals in human-modified landscapes to predict how varying
28
levels of afforestation or dairy production in the Drakensberg grasslands of South Africa would
influence alpha diversity.
Some generalities emerge from the literature that may be helpful in working towards
sensible policies. Generally, diversifying human-modified landscapes at all levels, e.g.
polyculture cropping, diverse agroforestry, and maintaining farmlands with high heterogeneity in
terms of both crops and vegetation structure, is likely to support more species than do more
homogenous land uses, while potentially also providing economic stability, given fluctuating
markets for specific crops (Franzen and Mulder 2007). Endemic and specialist species tend not
to persist in human-modified landscapes; thus, protected areas continue to be crucial to
conservation efforts, and expanding PA coverage within areas rich in such species is important
(Jenkins et al. 2013). Furthermore, a regional approach to management of these sensitive species
is necessary. Past and present implementation strategies are beyond the scope of this review, yet
there is literature dealing with such strategies in Africa that may be of use, e.g. certification of
sustainable and biodiversity friendly products (Lilieholm and Weatherly 2010).
4.3. Living with nature
Maintaining biodiversity in landscapes where humans live, work, and extract resources implies
that humans will have to coexist with other species. While the consequences of living without
nature may be worse than the difficulties of living with it, certain issues present considerable,
though not unmanageable, obstacles for promoting conservation beyond PAs. Many, and perhaps
most, species are easily compatible with human livelihoods and may have a positive impact via
aesthetic or functional value, but some, especially large mammals, can be problematic in humandominated landscapes, e.g. carnivores threaten livelihoods by predating livestock and,
29
occasionally, people. Nonetheless, specific and practical actions can greatly reduce the
probability of carnivore attacks. For example, in Kenyan communal lands, having a domestic
dog accompany herds reduces risk of a carnivore attack by 63%; conversely each additional
boma gate increases risk of attack by 40% (Woodroffe et al. 2007). However, carnivores are not
the only concern. Other animals, such as baboons and bush pigs, can damage structures and
destroy crops while larger herbivores, such as elephants, also threaten human lives. Knowledge
of attitudes of people employing different land uses can help land-use planners develop strategies
to reduce conflict and negative attitudes towards conservation. For example, crop agriculture
should not be encouraged in predominantly pastoral areas where elephants and people coexist
relatively peacefully (Gadd 2005). Furthermore, land-use planning that incorporates knowledge
of which crops are most likely to generate conflict could allow creation of buffer zones in areas
with high conflict potential (Hockings and McLennan 2012).
The risk of disease transmission poses an additional difficulty. Diseases of domestic
animals threaten wildlife. For example, domestic dogs are carriers of canid diseases transmissible
to wild carnivores (Butler et al. 2004) and were partly responsible for extinction of the African
wild dog Lycaon pictus and decimation of lions Panthera leo in areas of the Serengeti (see
Woodroffe 1999). Additionally, livestock can transmit animal diseases (e.g. bovine tuberculosis)
to wildlife with negative conservation outcomes, while wildlife can also transmit diseases (e.g.
foot and mouth) to livestock with immense economic consequences (Michel et al. 2006;
Thomson 2009).
Fencing has been heavily used in Africa to assist people in their ability to coexist with
nature—to reduce direct conflict and disease transmission. Laws regarding fencing differ by
country; for example, Zambia requires game fences while Namibia encourages large-scale
30
cooperation between game-farmers to discourage fencing (McGranahan 2008). Obviously,
fencing has serious ecological consequences (Trimble and van Aarde 2010; Hayward and Kerley
2009) and is anathema in many ways to the goals of conservation, especially conservation
beyond PAs (Trimble and van Aarde 2010; Woodroffe et al. 2014). However, non-traditional
fencing technologies (see Hayward and Kerley 2009), such as fences targeted at particular
problem species (e.g. elephant fences that allow other species to pass), virtual barriers, or fencing
wildlife out of villages and fields instead of into PAs, may be acceptable compromises. The
effect of fences on the persistence of species in human-modified landscapes certainly deserves
more investigation.
Economically, wild animals provide an important resource for many people in Africa
(Bharucha and Pretty 2010), which may threaten species persistence. „Sustainable use‟ is
frequently discussed with relation to bushmeat hunting, but food scarcity and population growth
dictate that it will likely be impossible to enforce rules for sustainable use unless food security
issues are addressed (Fa et al. 2003). Sustainable harvesting is also an issue for plants (Sambou
et al. 2002). Community forests must be carefully managed, e.g. by restricting harvesting of
pole-sized stems to certain species, to ensure that species are not used to extinction (Obiri et al.
2002). Additionally, rules must be assessed to ensure that they achieve the desired goals; for
example, in the Republic of Guinea, tax to the forestry administration for harvesting palm wine
counterproductively encourages harvesters to employ lethal yet profitable methods of harvesting
to compensate for the initial investment (Sambou et al. 2002).
31
5. Conclusion
There is clearly both necessity and great potential for human-modified land in sub-Saharan
Africa to contribute to the conservation of the continent‟s biodiversity. While PAs will remain
essential, and are especially important for protecting species sensitive to human disturbance
(Devineau et al. 2009), a greater focus on biodiversity conservation beyond their boundaries
could be complementary to overall conservation goals. The information gleaned from studies of
biodiversity in human-modified landscapes in Africa discussed in this review goes some way
toward providing policy-makers with evidence to support defensible decisions for land-use
planning and conservation management beyond PAs (see Table 1). Improving the amenity of
human-modified landscapes for biodiversity can be encouraged at all levels from individuals‟
choices to plant indigenous home gardens, to grass roots endeavors to manage communal
resources, to communities deciding to share their land with wildlife, to commercial farms going
organic and maintaining patches of natural habitat. Governmental intervention at the level of the
city (e.g. green space planning), region (e.g. extension agencies demonstrating biodiversity
friendly agricultural practices), nation (e.g. policy-setting for control of invasive species,
pesticide or poison usage, and land-use zoning), or even internationally (e.g. cooperative removal
of boundary fences) are also warranted.
Although several factors including lack of knowledge, implementation challenges, and
problems of coexistence with wildlife may constrain successful implementation of biodiversity
conservation in human-modified landscapes, given each constraint, opportunity exists for
progress. On the bright side, scientific interest in the topic is increasing (Trimble and van Aarde
2012), and as research accumulates, it will allow for systematic reviews useful for policy
decisions. Additionally, many issues associated with human-wildlife coexistence are primarily
32
related to large mammals and efforts to solve these problems should continue. Meanwhile, the
barriers to implementing strategies to conserve other species groups in human-modified
landscapes are far from insurmountable and such strategies should be prioritized.
Acknowledgements
M.J.T. was supported by a National Science Foundation Graduate Research Fellowship and
R.J.v.A. through various grants to the Chair in Conservation Ecology at CERU.
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