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ORIGINAL PAPER
Conserving biodiversity in a changing world: land use
change and species richness in northern Tanzania
Maurus J. Msuha •Chris Carbone •Nathalie Pettorelli •
Sarah M. Durant
Received: 24 August 2011 / Accepted: 6 July 2012
ÓSpringer Science+Business Media B.V. 2012
Abstract Even though human induced habitat changes are a major driver of biodiversity
loss worldwide, our understanding of the impact of land use change on ecological com-
munities remains poor. Yet without such information it is difficult to develop management
strategies for maintaining biodiversity in the face of anthropogenic change. To address this
gap, we explored how land use practices impacted species richness in a mammalian
community in northern Tanzania. Using camera traps, we estimated the number of
mammalian species inhabiting three land use types subjected to increasing levels of
anthropogenic pressure: (1) Tarangire National Park, (2) pastoral grazing areas; and (3)
cultivated areas outside the park. Results showed that land use practice is correlated with
different levels of species richness. Interestingly, mammal species richness was highest in
the grazing areas and lowest in cultivated areas. When we focused our analyses on car-
nivores, we found little significant difference in species richness between the park and
pastoral grazing areas, however, carnivore richness were significantly lower in the culti-
vated areas. We found no significant link between species body weight and presence in the
three areas considered. Altogether, our results show that biodiversity conservation can be
achieved outside national parks, with pastoral grazing areas holding a significant propor-
tion of mammal communities; however increasing cultivation of pastoral rangelands may
represent a major threat to mammalian communities.
M. J. Msuha (&)
Tanzania Wildlife Research Institute, P.O. Box 661, Arusha, Tanzania
e-mail: Maurus.Msuha@gmail.com
M. J. Msuha
Department of Anthropology, University College London, Gower Street, London WCE 6BT, UK
M. J. Msuha C. Carbone N. Pettorelli S. M. Durant
Institute of Zoology, Zoological Society of London, Regent’s Park, London NW1 4RY, UK
S. M. Durant
Wildlife Conservation Society, International Conservation, 2300, Southern Boulevard,
New York, NY 10460-1099, USA
123
Biodivers Conserv
DOI 10.1007/s10531-012-0331-1
Keywords Camera traps Mammals Tarangire National Park
Anthropogenic pressure Body weight
Introduction
It is widely accepted that global biodiversity is decreasing at an alarming rate, and that
much of this change is induced by human activities (Estes et al. 2011). Of all human
impacts on biodiversity, land use change has been singled out as the greatest immediate
threat to terrestrial biodiversity, because it results in fragmentation and loss of habitats
(Vitousek et al. 1997; Sala et al. 2000; Jetz et al. 2007), restricts animal movements and
may lead to declines in species richness and abundance (Pimm and Raven 2000). Carni-
vores, especially large ones, are seriously affected by land use change: most large carni-
vores range widely (Woodroffe and Ginsberg 1998) and their survival is dependent upon
the availability of large strips of habitats, which are increasingly becoming scarce in the
face of increasing anthropogenic pressure (Nowell and Jackson 1996; Sillero Zubiri et al.
2004; TAWIRI 2006,2007a,b,d, ). Land use change is particularly rapid in developing
regions such as Sub Saharan Africa, where human populations re expanding and where the
majority of people depend on natural resources for their livelihoods (Ceballos and Ehrlich
2006; Thuiller et al. 2006).Unfortunately this is also the region where information on
impacts of anthropogenic land use change on biodiversity is most lacking. This is a
particular concern, since understanding how species respond to land use change, provides
important information for mitigating its impacts on biodiversity loss (Rodriguez 2003).
Historically, protected areas have been established to protect and maintain biological
diversity and their natural and associated cultural resources (Pressey 1996; Chape et al.
2005). However, while these protected areas may provide sufficient protection for smaller
species, many of which are less wide-ranging, some of the protected areas are too small to
be sufficient for the survival of larger and more wide-ranging species (Woodroffe and
Ginsberg 1998; Patterson et al. 2004; Baeza and Estades 2010). This means that successful
conservation of wildlife must not only rely on protected areas, but also the surrounding
unprotected lands as well, if ecological integrity is to be maintained (Newmark 1996;
Woodroffe 2000). Although we know that areas outside protected areas can hold signifi-
cant populations of many wildlife species (Homewood and Rodgers 1991; Western and
Gichohi 1993).
Studies have shown that increased intensity of land use reduces habitat diversity which
leads to a decrease in species diversity (Fitzherbert et al. 2008; Maitima et al. 2009;
Wretenberg et al. 2010). However it has also have shown that species richness can be
higher in areas with intermediate levels of disturbance—the intermediate disturbance
hypothesis (Connell 1978; Huston 1979), where limited disturbance reduces competitive
exclusion of certain species (Connell 1978). We might therefore expect that species
richness will vary across land use types in relation to the degree to which they are exposed
to human disturbance (H1). Moreover, because variation in body size among species has
been proposed as an important determinant of extinction probability (Lindstedt et al. 1986;
Belovsky 1987; Cardillo et al. 2005), and because large species tend to have large home
ranges, have lower population densities (Jetz et al. 2004) and are more sensitive to human
disturbances (Gittleman and Harvey 1982; Crooks 2002; Kauffman et al. 2007), we also
expect that large-bodied mammals will be more sensitive to anthropogenic pressure than
small ones (Cardillo 2003; Cardillo et al. 2005) (H2). In addition, species at higher trophic
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123
levels (e.g. mammalian carnivores) are known to live at lower population densities, have
larger home ranges and day ranges than non-carnivore species of the same size (Jetz et al.
2004; Carbone et al. 2005). We therefore explore changes in response to habitat distur-
bance for all mammal species and we treat carnivores and non-carnivores separately.
With this study we aim to address these hypotheses by exploring how land use practices
impact species richness in northern Tanzania. Using camera traps, we estimated the
number of medium-sized mammalian species inhabiting three land use types subjected to
increasing levels of anthropogenic pressure: (1) Tarangire National Park (hereafter TNP),
(2) pastoral grazing areas outside the park; and (3) cultivated areas outside the park. We go
on to explore differences in patterns of species richness between carnivores, usually at the
highest trophic level, and other mammals, mainly herbivores to inform our analysis.
Materials and methods
Study area
The Tarangire ecosystem in northern Tanzania is within the semi-arid ecological zone with
annual rainfall ranging from 450 to 600 mm (Prat et al. 1966) (Fig. 1). The ecosystem is
globally important for biodiversity conservation, and is ranked second after the Serengeti-
Mara ecosystem for high concentrations of migratory mammals in Eastern Africa (Reid
et al. 1998; Ludwig et al. 2008). About 85 % of the Tarangire ecosystem is comprised of
village and private lands outside core protected areas which include TNP (TCP 1997).
These village and private lands outside the park, particularly to the east, are key to long
term survival of wildlife populations in the ecosystem as they provide quality pasture for
migratory and lactating mammals (Voeten 1999). However, over the last three decades,
most of these areas to the east of the park in Simanjiro Plains have been under pressure
from cultivation, which threaten long term survival of wildlife in the ecosystem (IRA
2001).
Camera trapping
Data on species richness were collected using camera traps (Rowcliffe et al. 2008; Tobler
et al. 2008; Pettorelli et al. 2010). Such type of surveys use fixed cameras (in this case
DeerCam 300 passive infrared cameras: Non Typical Inc, Park Falls, USA), triggered by
infrared sensors, to ‘‘trap’’ images of passing animals. It is a form of static monitoring and
a quantitative technique that has relatively low labor costs, is non-invasive, incurs minimal
environmental disturbance, is robust to variation in ground conditions or climate and, most
importantly, can be used to gain information on highly cryptic species in difficult terrain
where other field methods are likely to fail (Silveira et al. 2003). The delay between
pictures was set to 1 min and the sensitivity for the infrared sensor was set to high.
Cameras were placed along animal trails at approximately 40 cm above the ground and
were left to operate for 24 h a day. Each camera operated between 30 and 75 days. All
camera stations were georeferenced using a Global Positioning System (GPS). Cameras
were checked regularly (every 10 days) to replace films and batteries. All films were
developed and printed, and mammals were identified to the species level.
Field surveys were conducted in 2006 and 2007 using a systematic grid of camera traps.
In 2006, a first survey was carried out in TNP from March to May during the wet season,
while a second survey was conducted in the cultivated areas outside the park from June to
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123
August during the dry season. In both cases, 40 cameras were deployed (20 camera
locations per survey, with a 2 km interval between locations). Each camera station had
double cameras placed opposite each other. In 2007 a third survey was conducted during
the wet season in the pastoral grazing areas (from mid February to mid March), while a
fourth survey was in the cultivated areas from mid March to mid April. Two last surveys
were conducted during the following dry season: one from mid June to mid August in the
park and another from end of August to the first week of October in the pastoral grazing
Fig. 1 Locations of camera traps in TNP, pastoral grazing areas and cultivated areas outside the park
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123
areas outside the park. Surveys in 2007 generally used 80 camera stations with a single
camera at each station placed at 2 km intervals. The exception to this was the survey in the
park where double cameras were only used for the grid that was surveyed in 2006 and
additional single cameras were set up outside this grid to make a total of 80 camera stations
(Fig. 1).
Statistical analysis
Camera traps provided geo-referenced and temporally informed information on species
occurrence. For each survey the number of camera traps set, the effort (estimated as the
number of trap-days), the type of land use practice, the period when the survey occurred
and the number of carnivore and non-carnivore species recorded were determined
(Table 1). For each land use type, the number of mammalian species that had been pho-
tographed (hereafter referred to as observed species richness) was recorded. Such infor-
mation was used in combination with the program EstimateS 7.5 (Colwell 2005)to
estimate species richness in the three areas considered (hereafter referred to as estimated
species richness). This program uses the first-order Jackknife that was initially designed to
estimate population size from capture-recapture data, which allows capture probabilities to
vary by individuals (Burnham and Overton 1979). The method can equally be applied to
estimate species richness (Colwell and Coddington 1994; Boulinier et al. 1998; Chazdon
et al. 1998; Nichols et al. 1998; Hughes et al. 2002). We used Chi square test to compare
observed and expected species richness and One-way ANOVA to test whether the means
for the three groups were equal. Furthermore, we used modified t-statistic to compare of
species richness among the three land use types (H1).
To investigate effect of body weight (H2) on species presence between land use types,
we used General Linear Models in Program R (R Core Development Team 2008) and
modeled species presence as a function of log-transformed body weight. Body weight for
each species was calculated as the average value between adult male and female weights
obtained from literature (Estes 1991; Kingdon 1997).
Results
20 carnivores and 23 non-carnivore species were photographed during these surveys
(Table 1). Of the 23 species of carnivores known to occur in the park (Foley 2005), 19 species,
roughly 83 % were photographed. For non-carnivore species, 18 species were photographed
in TNP. In the pastoral grazing areas 17 carnivores and 22 non-carnivore species were
photographed. In the cultivated areas, six carnivore species and 14 non-carnivores species
were photographed. Results suggested that some species might have limited distribution
patterns e.g., leopard (Panthera pardus) was only photographed in the park; lion (Panthera
leo) and spotted hyaena (Crocuta crocuta) were photographed both in TNP and in the pastoral
grazing areas but not in the cultivated areas; springhare (Pedestes capensis) was photo-
graphed only in the cultivated areas while bush duiker was photographed only in the grazing
areas). Overall analysis using Chi square test revealed significant difference in observed and
expected mammal species richness (v
2
=126, df =2, P\0.01, Fig. 2).
As expected from (H1), mammal species richness was lowest in cultivated areas in both
seasons (dry: park/cultivated t
110.8
=55.0; grazing/cultivated t
148.9
=40.9 and wet: park/
cultivated t
29.5
=28.2; grazing/cultivated t
63.2
=40.6, all p\0.001). However, richness
was significantly lower in the park than in pastoral grazing areas in the dry season but not
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Table 1 List of species photographed during the surveys conducted in 2006 and 2007 for each land use
type (N =43)
Species Park Grazing
areas
Cultivated
areas
Body
weight
Canidae
Black-backed jackal Canis mesomeles HH H 10
Bat-eared fox Otocyon megalotis HH – 4.5
Felidae
Caracal Felis caracal H–– 13
Leopard Panthera pardus H– – 53.25
Lion Panthera leo HH – 178.5
Serval Felis serval HH H 11.63
Wild cat Felis silvestris HH H 4.75
Herpestidae
Banded mongoose Mungos mungo HH – 0.80
Bushy-tailed mongoose Bdeogale crassicauda –H– 1.7
Dwarf mongoose Helogale parvula HH – 0.28
Egyptian mongoose Herpestes ichneumon HH – 3.15
Slender mongoose Galerella sanguinea HH – 0.62
White-tailed mongoose Ichneumia albicauda HH – 3.6
Hyaenidae
Aardwolf Proteles cristatus HH –10
Spotted hyaena Crocuta crocuta HH –65
Striped hyaena Hyaena hyaena HH H 40
Mustelidae
Honey badger Mellivora capensis HH – 11.5
Zorilla Ictonyx striatus HH H 1.05
Viverridae
Common genet Genetta genetta HH H 1.78
Large spotted genet Genetta maculata H– – 2.15
Cercopithecinae
Olive baboon Papio anubis HH – 28.25
Vervet monkey Cercopithecus pygerthrus HH H 5.13
Leporidae
Cape hare Lepus capensis HH H 2.25
Pedetidae
Spring hare Pedetes capensis –– H3.5
Hystricidae
Crested porcupine Hystix cristata HH H 19.5
Orycteropodidae
Aardvark Orycterpus afer HH H 61
Elephantidae
African elephant Loxodonta africana HH – 4000
Equidae
Burchell’s zebra Equus burchelli HH H 241.75
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in the wet (dry: park/grazing t
106
=2.90, p\0.01 and wet: park/grazing t
32
=0.27, n.s,
Fig. 2). Carnivore richness was also significantly lowest in cultivated areas (dry: park/
cultivated t
153
=70.0; grazing/cultivated t
114
=42.6 and wet: park/cultivated t
19
=48.0;
grazing/cultivated t
81
=48.2, all p\0.001), but was highest in the park in the wet season
(dry: park/grazing t
114
=0.47, n.s. and wet: park/grazing t
45
=8.0, p\0.001, Fig. 3).
For non-carnivores, species richness was lowest in the cultivated areas, and highest in the
grazing areas (dry: park/cultivated t
74
=11.0; grazing/cultivated t
144
=12.5; park/grazing
t
79
=6.2 and wet: park/cultivated t
36
=6.2; grazing/cultivated t
29
=17.5; park/grazing
t
25
=7.3, all p\0.001, Fig. 4). The proportion of species ‘‘loss’’ due to cultivation was
much higher for carnivores than for non-carnivores. We found that 61 % of the non-
carnivore species photographed during our surveys were found in cultivated areas whereas
only 32 % of the carnivore species recorded were photographed in cultivated areas. We
found no significant link between species body weight and species occurrence across the
three areas considered (all P[0.05, Table 2).
Discussion
This is one of the first studies to demonstrate variation in mammalian species richness
along a gradient of changing land use in areas of high mammalian diversity. Our results
support our first hypothesis (H1) that species richness will vary across land use types in
relation to the degree to which they are exposed to human disturbance. Specifically we
Table 1 continued
Species Park Grazing
areas
Cultivated
areas
Body
weight
Suidae
Bush pig Potamochoerus larvatus –HH 97.5
Warthog Phacochoerus africanus HH H 82.5
Giraffidae
Masai giraffe Giraffa camelopardalis HH H 1340
Bovinae
African buffalo Syncerus caffer HH – 550
Bushbuck Tragelaphus scriptus HH – 48.5
Lesser kudu Tragelaphus imberdis HH – 81.5
Eland Taurotrogus oryx HH H 560.5
Antelopinae
Bush duiker Sylvicapra grimmia –H– 17.5
Steinbuck Raphicerus campestris –HH 11.5
Kirk’s dikdik Madoqua kirkii HH H 5.5
Waterbuck Kobus ellipsiprymnus HH – 215
Thomson’s gazelle Gazella rufifrons –H– 23.75
Grants gazelle Gazelle grantii HH H 61.5
Aepycerotinae
Impala Aepyceros melampus HH H 56.25
Alcelaphinae
Hartebeest Alcelaphus buselaphus HH – 161
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Fig. 2 Estimated mammal species richness in TNP, pastoral grazing areas and cultivated areas TNP,
grazing areas and cultivated areas using first order Jackknife
Fig. 3 Estimated carnivore species richness in TNP, pastoral grazing areas and cultivated areas using first
order Jackknife
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showed that: (1) land use practice is an important factor influencing species richness, (2)
species richness was significantly lower in cultivated areas, and (3) overall species richness
was higher in grazing areas consistent with the intermediate disturbance hypothesis
(Connell 1978). However, our results did not support our second hypothesis (H2) that large
bodied mammals would be more sensitive to anthropogenic pressure, probably because
some large migratory herbivores were recorded in each land use type and thus swamped
any significant effect. Overall our results suggest that while grazing areas could have a
positive impact on species richness, intensive cultivation outside the park may represent a
major threat to mammalian biodiversity.
Although body size was not a significant factor affecting species richness, carnivores
appeared to exhibit a more pronounced response to habitat disturbance with the largest
species in particular not being found in the cultivated areas. More specifically, aardwolf
(Proteles cristatus), banded mongoose (Mungos mungo), dwarf mongoose (Helogale
Fig. 4 Estimated non-carnivore species richness in TNP, pastoral grazing areas and cultivated areas using
the first order Jackkinife
Table 2 Summary statistics of general linear model for effect of body weight on species presence in the
park, pastoral grazing areas and cultivated areas outside the park (N =43)
Variable Estimate SE Z P
Intercept 1.257 0.822 1.529 0.126
Pastoral grazing areas-park 0.178 1.262 0.141 0.888
Cultivated areas-park -1.480 1.014 -1.459 0.145
Log weight 0.442 0.590 0.750 0.453
Pastoral grazing areas 9log weight 0.261 0.968 0.27 0.787
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parvula), bat-eared fox (Otocyon megalotis) and honey badger (Mellivora capensis) were
mainly confined to the national park, and yet are thought to be distributed widely in the
region according to their known distribution and ecology (Kingdon 1997; TAWIRI 2006,
2007a,c). However, the presence of bushy-tailed mongoose (Bdeogale crassicauda) in the
grazing areas was unexpected, although a recent camera trap study in Tanzania (Durant
et al. 2010; Pettorelli et al. 2010) found the species in many areas not previously recorded,
suggesting that the species may be more common than previously thought.
Overall mammal species richness was higher in pastoral grazing areas consistent with
the intermediate disturbance hypothesis (Connell 1978; Huston 1979). The management of
pastoral rangelands often involves grazing, burning and moderate use of some tree species
for fuel or for construction (Homewood and Brockington 1999). It is argued that these kind
of management practices play a significant role in maintaining savannah species richness
(Homewood and Brockington 1999) and may influence regeneration that can foster more
habitat diversity and species-rich communities (Fairhead and Leach 1996; Nyerges 1996).
It has also been shown, for example, that pastoral land use practices increase frequency of
Acacia tortilis, which is an important forage crop for both wild ungulates and livestock
(Muchiru 1994; Reid and Ellis 1995). Studies elsewhere have also found high diversity
herbivores outside protected areas in abandoned pastoral settlements because of high
quality forage produced by heavy dung deposits that enrich soil nutrients (Muchiru et al.
2008). However the increased species richness we found in grazing areas was found only in
non-carnivores, and disturbance was not associated with an increase in diversity of
carnivores.
Interestingly, our results showed that some non-carnivore species were more sensitive to
environmental change than others: for example, 9 out of the 23 species that were recorded
during our surveys were not found in the cultivated areas, while spring hare (P. capensis)
was the only species recorded in cultivated areas. We believe such results could be
explained by variation in species’ ecological needs. For instance, cultivated areas were the
only land use type where spring hare was photographed, probably because the species
prefers open sandy dry soils and croplands with sufficient cereals (Augustine et al. 1995).
Similarly, grazing areas were the only land use type where bush duiker (Sylvicapra
grimmia) was photographed, although the species is thought to be rare in TNP (Charles
Foley, pers. comm.). Furthermore, carnivores were much less likely than non-carnivores to
be found in cultivated areas, suggesting that impacts of cultivation are larger on species at
higher trophic levels.
Overall body size alone was not found to be a significant factor predicting the response
of species to levels of habitat disturbance (H2). However, perhaps this is not surprising.
Often patterns predicted from large-scale assemblages of species differ from those found at
local scales (Tilman et al. 2004; White et al. 2007). At local scales, with smaller numbers
of species represented, individual species characteristics, trophic levels and diets will have
a major influence on distributions and habitat use. In this study some of the large herbi-
vores seemed to be less susceptible to intermediate disturbance levels than some of the
carnivore species. Because of limited sample size we were unable to explore the impacts of
different season on species richness. Tarangire is a migratory system, and hence it is likely
that there will be seasonal effects, and this would be worth for further investigation.
However most of the species recorded are non-migratory, and hence we expect the overall
patterns of species richness detected in this study to hold regardless of season.
Finally, our results have important implications for the conservation of wildlife in the
Tarangire ecosystem and elsewhere. The high proportion of carnivore species ‘‘loss’’ in
cultivated areas suggest that carnivores are likely to disappear first in the face of increasing
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123
anthrogenic pressure. Increasing human pressure is a reality in the Tarangire ecosystem,
particularly in the Simanjiro and Monduli districts, where agricultural expansion cause
habitat loss and fragmentation. The information we report could be crucial for wildlife
managers in their task of maintaining a full species complement within an ecosystem, by
enabling them to adapt management strategies to the relative sensitivity of species to
anthropogenic disturbance. Our results also highlight the importance of grazing areas for
wildlife (especially for non-carnivore species), suggesting that conservation efforts should
not only focus on protected areas but also in favourable habitats outside protected areas,
such as the pastoral grazing areas outside TNP.
Acknowledgments We would like to thank Paul Baran, Isaya Samwel, Benjamin Samwel, Chediel
Kazaeli and Zawadi Mbwambo and Alex Lobora for helping with field surveys and other logistics; and
Tanzania Wildlife Research Institute and Tanzania National Parks for providing permits for the study.
Additionally we would like to thank villagers and village governments in Loiborsoit, Emboeret and Lokisale
for their support during the field work. This study would not have been possible without financial support
from the Dorothy Hodgkin Postgraduate Award, the St. Louis Zoo, Wildlife Conservation Society, the
Zoological Society of London, the Howard G. Buffett Foundation, British Ecological Society and Inter-
national Foundation for Science and the Serengeti Cheetah Project.
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