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Complex responses to climate and anthropogenic changes: An evaluation based on long-term data from Kibale National Park, Uganda



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5. Complex Responses to Climate and Anthropogenic Changes: An Evaluation Based on
Long-Term Data from Kibale National Park, Uganda
C.A. Chapman, L.J. Chapman, R. Ghai, J. Hartter, A.L. Jacob, J.S. Lwanga, P. Omeja, J.M.
Rothman, & D. Twinomugisha
Today the world is experiencing many changes and stresses that affect ecological systems.
For example, in the past 50 years the size of the human population has grown by 3.7 billion people
(Potts 2007) and is expected to grow to between 8 and 10 billion people within the next half
decade (United Nations 2005). Between now and 2050, 90% of this growth will occur in
developing countries, many of which are tropical and sustain the greatest proportion of the world’s
biodiversity (Potts 2007). This exponential growth in human populations has brought many
challenges to ecological systems and is responsible for the ecosystem modification that threatens
biodiversity on a global scale. Continued population growth is predicted to exacerbate the ever-
increasing demands for environmental products and services, particularly in the tropics (Houghton
1994). For example, the net loss in global forest area between 2000 2005 was ~7.3 million ha per
year (~200 km2 of forest per day, FAO 2005). This does not consider the vast areas being
selectively logged or the forests degraded by fire, both of which can impact huge areas. For
example, during the 1997/1998 El Nino, 7 million ha of forest burned in Brazil and Indonesia
alone (Chapman & Peres 2001). Even when the physical structure of the forest remains intact,
subsistence and commercial hunting can have a profound impact on forest animal populations. For
example, Chapman & Peres (2001) estimate that 3.8 million primates are eaten annually in the
Brazilian Amazon.
In Uganda, the country where this study focuses, threats to biodiversity are similarly
grave. Closed-canopy tropical forest once covered 20% of the country, but deforestation has
reduced this to just 3% (Howard et al. 2000). Uganda lost 18% of this remaining forest between
1990 and 2000 (Howard et al. 2000). The most recent estimate suggests that the annual rate of loss
of tropical high forest in Uganda is 7% (Pomeroy & Tushabe 2004).
Ecosystems are also indirectly affected from human induced changes to the environment,
possibly the most dramatic of which is climate change. The earth’s climate has warmed by
approximately 0.6°C over the past 100 years, and some estimates suggest that the climate could
warm by up to 5.8oC by the end of this century (IPPC 2001; Walther et al. 2002). Recently, there
has been a growing appreciation within the academic community and policy makers regarding the
potential magnitude of climate change on ecosystems (Hannah et al. 2002; McClean et al. 2005;
van Vliet & Schwartz 2002; Walther et al. 2002). There have been numerous documented shifts in
the distribution, population abundance, life history, and even survival of species in response to
climate change (Hannah et al. 2002; Parmesan & Yohe 2003; Pounds et al. 1999), although the
responses to climate change in particular regions, or with respect to specific species are largely
unknown, particularly for the tropics.
Our objective is to use long-term data from Kibale National Park, Uganda to consider
complex responses to climate and anthropogenic changes, focusing on changes in forest
composition, phenology patterns, disease, and primate abundance over time. In many of these
comparisons we consider changes in logged and unlogged forests. These analyses build on the data
and publications of many people. Of special importance is the work of Tom Struhsaker who not
only founded the field station in Kibale and protected the area for many years, but provided us
with data from 1970 to 1987. Jerry Lwanga started research in Kibale in 1983 and Colin and
Lauren Chapman came to Kibale in 1989. The remaining authors of this paper have been in Kibale
for shorter periods, but provide valuable data for comparison.
Kibale National Park, Uganda: The Environment and Recent Changes
Kibale National Park (795 km2) is located in western Uganda (0°13' - 0°41' N and 30°19' -
30°32' E) near the foothills of the Ruwenzori Mountains (Figure 5.1). Kibale is a mid-altitude,
moist-evergreen forest receiving 1697 mm of rainfall annually (1990-2009). Kibale has a
complicated history of human impacts, both in terms of commercial logging and agricultural
clearing (Chapman & Lambert 2000; Struhsaker 1997). Kibale was established as a forest reserve
between 1926 and 1932 with the goal of protecting the watershed and providing sustained
production of hardwood timber (Osmaston 1959; Struhsaker 1997). A polycyclic felling cycle of
70 years was initiated, and it was recommended that the canopy be opened up by approximately
50% through the harvest of trees over 1.52 m in girth (Kingston 1967). This history of logging has
led to varying degrees of disturbance among sites (Skorupa 1988; Struhsaker 1997). Kibale was
gazetted a National Park in 1993 and since that time has been well protected by the Uganda
Wildlife Authority (UWA).
Chapman and Lambert (2000) conducted a landcover analysis in the mid-1990s and
estimated that 57.9% of the park was forest, 14.6% was grassland, 5.8% was woodlands (in the far
south), and 2.2% was wetlands and lakes (Figure 5.1). Plantations (especially the pines Pinus
patula and P. caribaea and cypress Cupressus lusitanica) were planted on 393 ha of grassland
hilltops between 1953 and 1977 (Kaumi 1989; Kingston 1967; Osmaston 1959) and constituted
1.0% of the park's area. These plantations were primarily found in the northern sections of the park
near Sebatoli and in more central areas around Kanyawara. They were harvested starting in 1996
and 10 years after harvest there were 44 species found in a 4 ha area that was evaluated, in some
places the trees had formed a closed canopy, and the above ground biomass was substantial
(Omeja et al. 2009). Abandoned farms (10.3%) and degraded forest (8.7%; largely representing
secondary forest associated with agricultural encroachment: see below) covered 146 km2. Seventy-
six percent of these degraded lands were found in the southern corridor, which links Kibale with
Queen Elizabeth National Park (QENP). The game corridor area covered 340 km2, of which 134
km2 had dual status and lay within the Forest Reserve and a Game Reserve (Howard 1991; MISR
Makerere University Institute for Social Research 1989; Struhsaker 1997; van Orsdol 1986). It was
managed by a separate government agency from the rest of the Kibale Forest Reserve until 1993,
when it was incorporated into the National Park. Unfortunately, we do not know what proportion
of the farms was established on areas that were forest versus grassland. However, if the areas that
are now abandoned farms were once all forested, this means that 79 km2 of forest has been lost.
Animal populations would also have been affected by the degradation of a further 66 km2 (8.7%)
of forest; the degree to which the forest and wildlife populations can recover after this disturbance
remains to be seen, but current trends are very promising (Omeja and Jacob unpublished data).
As early as 1971, illegal destruction and encroachment occurred in the corridor. In 1976,
approximately 30 eviction orders were issued, but were not carried out. In 1983, the government
again ordered settlers out of these encroached areas, and by 1984 it was estimated that 60% of the
forest plots and 30% of the grassland plots had been abandoned. However, the situation soon
reverted to the prior state and encroachment increased. On April 1st, 1992, the government ordered
settlers off the encroached land and resettled the people with the aid of a United Nations program
(Chapman & Lambert 2000). Estimates of the number of people residing in the southern corridor
vary dramatically. Based on aerial surveys counting houses, van Orsdol (1986) estimated that
8,800 people were living in the southern corridor. A national census carried out in 1980 indicated
that as many as 17,000 people were residing within the bounds of the Kibale Forest Reserve.
Baranga (1991) estimated 40,000 people, Makerere Institute for Social Research (MISR) (1989)
reported some 60,000 people, and after the resettlement the National Environment Management
Authority (NEMA 1997) estimated that 30,000 households, or approximately 170,000 people, had
been residing in Kibale. The extreme variance in these estimates (from 8,800 to 170,000)
illustrates the need for careful research prior to the initiation of resettlement programs. Based on
surveys conducted by CAC and LJC just after the resettlement program was completed, it is our
impression that most of these estimates were high. It seems likely that the larger estimates were
politically and economically motivated.
Figure 5.1. A map of Kibale National Park and its placing within Uganda and Africa.
Climate Change: Patterns and Possible Cascading Disease Effects
There is growing appreciation that climate change has or could cause significant change to
the abundance of particular species and ecosystems. We have rainfall data between 1903 and 1971
from the Government of Uganda meteorological records for the town of Fort Portal (approximately
20 km east of Kanyawara), between 1972 to 1989 from Tom Struhsaker at Kanyawara, and
between1990 and present from CAC and LJC, again at Kanyawara. Temperature data were
available from NASA ( between1905
and 1948 at Fort Portal, between1976 and1989 from Tom Struhsaker at Kanyawara, and between
1990 and present from CAC and LJC at Kanyawara.
These data sources suggest that the Kanyawara area of Kibale receives approximately 1697 mm
of rainfall annually (1990-2009), but it receives ~300 mm more rainfall/year today than it did at the start
of the 19th century. Other climatic changes include less frequent droughts, an earlier onset of the rainy
season, and an increase of just over 4oC in average maximum monthly temperature over 33 years
(Chapman et al. 2005; but see Stampone et al. unpublished data). These changes are much higher than
global averages (IPCC 2001, but see Altmann et al. 2002 for a similar example from East Africa). Since
the discovery of the magnitude of this change, we established a network of temperature monitors and
rainfall gauges throughout the park and surrounding landscape (e.g., agricultural land and tea
plantations). This research is being conducted by JH and CAC. We have also been able to obtained
rainfall records from a number of government rainfall stations from 1940 to 1975 (Stampone et al.
unpublished manuscript). Trends identified in this older data set are consistent with current trends where
newer data are available. However, analysis of this data demonstrate a high degree of spatial variability
with respect to trends in total seasonal and annual rainfall with some areas getting progressively wetter
and some becoming drier. This indicates that significant variation is occurring at very localized scales,
which suggests that monitoring the impact of climate change must be conducted at the appropriate,
localized scale. Seimon and Phillips (this volume) similarly point out that simulation models of climate
in the Albertine Rift are both high in magnitude and spatially heterogeneous. At the regional scale these
models suggest a general increase and a temporal redistribution of precipitation.
Climate change will have direct impacts on systems (e.g., making them too dry for particular
plant species or shifting the ranges of animals); however, it can also have unanticipated or cascading
changes. It is on such complex changes that we believe academics should direct their efforts. One
potential cascading effect of climate change is the alteration of host-parasite dynamics. In fact, the
human medical literature illustrates many connections between climate and disease, with specific
diseases occurring during certain seasons or in association with specific unseasonable weather
conditions. For example, in sub-Saharan Africa, meningococcal meningitis epidemics occur during the
hot dry season (Patz et al. 1996). Similarly, in the United States, 68% of waterborne disease outbreaks,
such as Giardia and Cryptosporidium, are preceded by heavy rain events (Hunter 2003). We do not
have long-term records of disease dynamics to associate with climate variables for wildlife in Kibale,
although we have shown that in wetter areas black-and-white colobus (Colobus guereza) have a higher
number of parasite eggs per gram of feces for some parasite species than dryer areas or sites (Chapman
et al. 2010; see also Stoner 1996). While some argue that gastrointestinal parasites have little impact on
host populations (Munger & Karasov 1989), others present evidence of biologically significant impacts,
particularly when the host population is nutritionally stressed (Chapman et al. 2006; Coop & Holmes
1996). The population effects of such changes in parasite infections remain to be evaluated for these
monkeys. However, since a 32-year record of plant phenological patterns (see below) suggests that
climate change in this region has led to periods of food scarcity causing nutritional stress in primates
(Chapman et al. 2005), and since stress has been shown to increase the biological significance of
parasite infections in colobus monkeys (Chapman et al. 2006), the impact of climate change on host-
parasite interactions should receive more detailed study.
Many factors have been proposed to drive phenological patterns of tropical rain forests;
however, abiotic characters, such as rainfall, day length, irradiance, and temperature are most
commonly cited to influence the timing of fruiting, flowering, and leafing events (Ashton et al.
1988; Newbery et al. 1998; Opler et al. 1976; van Schaik 1986; van Schaik et al. 1993). If climatic
factors are important determinants of phenological patterns, we might expect that the magnitude of
the climate change documented in Kibale would result in changes in fruiting, flowering, and
leafing patterns. We have used two data sets (1970 1983, Tom Struhsaker and 1990 2002, CAC
and LJC) to describe the fruiting patterns of the tropical tree community in Kibale (Chapman et al.
2005; Chapman et al. 1999). To address variation in spatial patterns in phenology (trees > 10 cm
dbh) and take advantage of the north-south decline in rainfall, we describe fruiting over 2-3 y
among four sites each separated by 12-15 km. The 1990 - 2002 phenology data illustrated high
temporal variability in the proportion of the populations fruiting. Interannual variation in
community-wide fruit availability was also high; however, the proportion of trees that fruited has
increased over the past 12 years (5.2). At the species level a variety of patterns were exhibited. A
number of the most common species rarely fruit, and when they do, <4% of the individuals take
part in fruiting events. Combining the two datasets for specific species in the Kanyawara region
again reveals an increase in the proportion of trees fruiting between 1990 and 2002, although the
proportion of the populations fruiting decreased during the earlier period (Figure 5.3). Using this
combined dataset to consider patterns for specific species reveals a variety of changes. For
example, Pouteria altissima exhibited a relatively regular pattern of fruiting during the 1970s;
however, it rarely fruited in the 1990s or 2000s (Chapman et al. 2005 CAC & LJC unpubl. data).
Similarly, Parinari excelsa fruited consistently in the 1970s, but rarely fruited in the 1990s or
2000s. This region of Kibale has long been considered a Parinari forest by foresters (Kingston
1967; Osmaston 1959; Struhsaker 1975), thus it is interesting that little regeneration is occurring in
this dominant species.
Figure 5.2. The 4 month running average of the percentage of trees bearing ripe fruit in Kibale
National Park, Uganda, between 1990 and 2002.
Month and year
% of trees with ripe fruit
Figure 5.3. The proportion of the tree population fruiting between 1970 and 1984, and 1990 and
2002 in Kibale National Park, Uganda. Sampling started in August, so this is the time interval
reported for each year. This includes the following species: Pouteria altissima, Celtis africana,
Celtis durandii, Diospyros abyssinica, Funtumia latifolia, Parinari excelsa, Strombosia scheffleri,
Teclea nobilis and Uvariopsis congensis.
Contrasting phenological patterns at four sites along the north-south elevational and rainfall
gradient within Kibale (Kanyawara, Dura River, Mainaro and Sebatoli) in the mid-1990s revealed
that at the community level the fruiting patterns of only one of the six pair-wise site combinations
were correlated and relationships between rainfall and fruiting were variable among sites.
Contrasting changes in fruiting patterns over 30 y between four sites varying in rainfall suggests
that the changes observed in fruiting may be due to climate change. For example, at the more
northern and wetter site (Kanyawara), Trilepisium madagascariense fruited at a low level in the
first few years of monitoring that began in 1990; however, this species has not fruited since 1993.
It is possible that the drier climatic conditions found in the early half of the century were more
favorable for the flower or fruit production of this species. To the south, at a drier site (Dura
River), a large proportion of the T. madagascariense population has repeatedly fruited in recent
years (Chapman et al. 2002a,b, CAC & LJC unpubl. data), suggesting that drier conditions may
promote fruiting. It seems possible that the wetter conditions now found in the north are no longer
suitable for the fruiting of this species, or these wetter conditions could have negatively impacted
this species’ pollinators.
Changes in Forest Structure: Climate Change or Forest History?
It seems feasible that such changes in climate could influence the ability of some species
to persist or grow to adulthood. No credit to the foresight of members of our group to anticipate
climate change, we established 11 permanent vegetation plots in the undisturbed forest near
Kanyawara (K30) in December 1989. Each plot was 200 m x 10 m, providing a sampling area of
2.2 ha. Plots were placed randomly within the existing trail system. Each tree with a diameter at
breast height (DBH) > 10 cm within 5 m of each side of the trail was identified, individually
Proportion of the population fruiting
Month and year
marked with a numbered aluminum tag and measured (DBH). The initial mean number of trees per
plot was 97+6.3 SE. This provided an initial sample of 1321 trees. Plots were resurveyed in May
2000 and September-November 2006. All tagged trees were re-located and measured to assess
growth and new trees recruiting into the size class of DBH > 10 cm were identified, measured, and
tagged. Mortality was noted and the cause of death was evaluated when possible.
It was our impression that the species that normally recruit into large gaps (i.e., 1 ha or
more) were the class of tree that would be most negatively affected by forest maturation, thus we
classified species into “mature forest specialists” and “large gap specialists” (Omeja et al. 2009;
Zanne & Chapman 2005). There are very few small gap specialists in the community, so this class
was excluded (Chapman et al. 2008). We contrasted two indices of tree community structure: 1)
change in cumulative DBH in the plots, and 2) annualized rate of population change. DBH has
been found to vary reliably with both fruit crop size and leaf biomass, is practical and easy to
measure, and has low inter-observer error (Brown 1978; Catchpole & Wheeler 1992; Chapman et
al. 1992; Chapman et al. 1994; Harrington 1979). Annualized rate of population change was
calculated using a standard model of exponential population growth:
r = lnNt lnN0
where Nt and N0 are population sizes at time t and time 0 and ln is the natural logarithm.
It is likely that primate populations are more strongly influenced by changes in the
abundance of tree species that produce food items than by the overall abundance of all trees. Thus,
we estimated the cumulative DBH of food trees in each plot, for each primate species, at each time
period. Kibale is an unrivaled location for this type of research because a large number of year-
long or longer studies have been conducted on the foraging behaviour of its primates and we could
rely on published diet data or raw data available to CAC. We used these data to determine what
should be considered to be food for each primate. We included foods (i.e., a specific part from a
particular species) that constituted ≥ 4% of the time spent feeding reported by Rudran (1978) and
Butynski (1990) for blue monkeys, Waser (1975) and Olupot (1994) for mangabeys, Harris and
Chapman (2007) and Oates (1977) for black-and-white colobus, Rode et al. (2006, unpublished
data) and Stickler (2004, unpublished data) for redtails, and Chapman and Chapman (2002,
unpublished data) and Struhsaker (Struhsaker 1975, unpublished data) for red colobus. We chose
the 4% cutoff because it included specific food items that were consistently considered important
by previous researchers, while avoiding large numbers of rarely consumed species.
We calculated the annualized rate of population change and the percentage of change in
cumulative DBH for species assigned to each recruitment category for each of the three sampling
periods and considered each plot to be independent sample points. To test the significance of
temporal variation in these parameters, we compared repeat samples of plots across years.
Repeated measures analysis of variance (ANOVA) tests were used to test the significance of
temporal (among years) and between recruitment categories (between subject effect), and their
interactions. Following Potvin et al. (1990), Mauchly’s criterion was used to test for the compound
symmetry of the variance-covariance matrix. When the criterion was rejected, the Greenhouse-
Geisser test, which relaxes the symmetry assumption, was used to obtain corrected significance
levels (Potvin et al. 1990). We conducted a one-way ANOVA when a single species was
As predicted, the rate of population change was more strongly negative for species classed
as recruiting into large gaps (average 1989 to 2000 r = -0.143; 2000 to 2006, r = -0.209) compared
to species recruiting into the main forest (average 1989 to 2000 r = -0.007; 2000 to 2006, r =
0.009; between-subject effects F=24.874, P<0.001). There was no significant time effect
(P=0.404), but a marginal interaction effect (P=0.055), suggesting that the difference between the
two classes of trees is increasing over time. As predicted, the percentage change in the cumulative
DBH in the vegetation plots was more negative for species classed as recruiting into large gaps
(average percent change 1989 to 2000 = -13.069; 2000 to 2006 = -13.047) compared to species
recruiting into the main forest (average 1989 to 2000 = -5.417; 2000 to 2006, 5.731; between-
subject effects F=10.446, P=0.004). There was no significant time (P=0.392) or interaction effect
This analysis demonstrates a change in species composition over time, but does not
necessarily indicate that this is driven by climate change. Rather, it may suggest that Kibale has
been disturbed in its distant past, and species that preferentially recruit into large gaps are dying
now that the forest is maturing. The window of time with which researchers view these forests is
very short compared to the life span of the trees that make up the forests. Although Kibale
obtained its first legal status in 1932, when it was gazetted a crown forest, the first descriptions of
the area were not made until the late 1950's (Osmaston 1959) and no single researcher has worked
in the park for more than 20 years. The analysis of pollen deposited on lake bottoms offers a
means of quantifying the history of the region’s flora much further back in time. Pollen diagrams
from the Ruwenzori Mountain Lakes (Livingstone 1967) and Kigezi in south western Uganda
(Hamilton 1974; Hamilton et al. 1986), suggest extensive anthropogenic forest clearance that dates
beyond 4800 years ago. A recent 6 m long core from the Kabata Swamp, 10 km from Kibale, finds
similar evidence of clearing (Taylor et al. 1999) approximately 2500 yrs ago; this clearing could
have been associated with the entry of Bantu-speaking peoples. There is a second episode of forest
clearing about 400 years ago associated with shifts in settlement patterns from grassland areas to
wetter, more forested regions (Taylor et al. 1999; Taylor et al. 2000). Within Kibale, a number of
pits for storing grain and an array of potshards have been discovered in what has traditionally been
considered undisturbed forest (Lang Brown and Harrop 1962; Lwanga et al. 2000). It seems likely
that in African forests, such as Kibale, human activities have altered forest composition for a
considerable period of time. Differences in forest composition between areas may reflect the
period of time that the area has had to recover from human induced disturbance. However, based
on what is known about the life history of the canopy trees in Kibale, identifying an area that had
been disturbed 1000 years ago from one that had been disturbed 400 years ago is a very difficult
There have also been significant changes in the park due to regrowth of areas that were disturbed
recently. A number of researchers have quantified the regeneration of areas of former pine plantations
(Chapman et al. 2002b; Duncan & Chapman 2003; Omeja et al. 2009), grasslands (Lwanga 2003),
replanted forest in the south of Kibale (Omeja et al. in prep; UWA-FACE 2005) and areas of logged
forest (Chapman & Chapman 2004). In all areas except the logged forest, the rate of forest regeneration
seems promising. In the logged forest the rate of regeneration seems to be very slow or even arrested
because of a complex set of interactions between tree seedlings/saplings, aggressive herbaceous
vegetation, primarily Acanthus pubescens, and elephants (Lawes & Chapman 2005; Paul et al. 2004;
Struhsaker et al. 1996).
Change in Primate Abundance Over 30 Years
Starting from the initial work by Tom Struhsaker (Struhsaker 1975), research in Kibale has focused
on primates. As a result we have an extensive record of primate behaviour, ecology, and abundance starting
in 1970 and continuing to date. Here we use 26 to 36 years of population and habitat data to determine the
potential causes of population changes for five species of sympatric primates in Kibale in areas that were
disturbed to varying intensities in the late 1960s. Line-transect censuses were initiated in the 1970s to
monitor primate biomass and have been replicated several times (Struhsaker 1970-75, Skorupa 1980-81,
Chapman 1996-97) using identical methods and trails (Chapman et al. 2000; Chapman et al. 2010 a,b;
Skorupa 1988; Struhsaker 1975). We calculated primate density (groups/km2) and encounter rate (#
groups/km walked) from line transect data. We established vegetation monitoring plots in 1989 (see above)
and have repeatedly measured species composition and habitat structure (cumulative DBH) in these plots
since that time. Using these data, we compiled a long-term assessment of the variation in Kibale primate
abundance and tree community composition and structure through time and space. We use diet
records from primate behavioural studies (see above) to identify important food resources so that we can
examine the relationships between population dynamics and changes in food availability and, for the red
colobus, food quality. We used the protein to fiber ratio as an index of red colobus food quality because it
has been found to be a good predictor of folivore leaf choice (Milton 1979) and biomass (Chapman &
Chapman 2002; Chapman et al. 2002a; Chapman et al. 2004; Ganzhorn 2002; Oates et al. 1990; Waterman
et al. 1988). To incorporate the effect of food quality, we weighted our DBH measures by the protein to
fiber ratio of both young and mature leaves of species constituting ≥ 4% of total feeding time. We found
that in general mangabey populations increased, blue monkeys declined, and redtails were stable in all
areas (Figure 5.4). Red colobus monkeys were generally declining (particularly indicated by encounter
rate), but may have been stable in the heavily logged area. The black-and-white colobus population is either
stable or marginally increasing in some areas. For blue monkeys and mangabeys, there were no significant
changes in food availability over time, yet their populations changed. For redtails, neither population
measures nor food availability changed over time. For black-and-white colobus, a decrease in food
availability over time in the unlogged forest surprisingly coincides with a possible increase in population
density. Finally, while red colobus food availability and quality increased over time in the heavily logged
area, their population did not show a corresponding increase. In the other unlogged and lightly logged
forestry compartment, a possible decline in red colobus (suggested by encounter rate) was not related to a
change in food quality, and in the lightly logged forest the decline in red colobus numbers corresponded
with an increase in food availability. We suggest that these populations are in non-equilibrium states. If
such states are widespread, it suggests that large protected areas will be required to protect species so that
declines in some areas can be compensated for by increases in adjacent areas with different histories.
Figure 5.4. Left/ The density (groups / km2) and Abundance or / encounter rate (groups seen per
kilometer of transect walked) of the five common diurnal primates in Kibale National Park.
K30=unlogged forest, K14=lightly logged forest, K15=heavily logged forest. * indicates that there
is a significant change in the density of abundance from the first estimate to the last, and brackets
connecting forestry compartments indicate a significant difference in abundance between areas
differing in logging history (solid) or a marginal difference (dashed line).
1970 1980 1990 2000 2010
1970 1980 1990 2000 2010
1970 1980 1990 2000 2010
1970 1980 1990 2000 2010
1970 1980 1990 2000 2010
1970 1980 1990 2000 2010
1970 1980 1990 2000 2010
1970 1980 1990 2000 2010
Encounter Rate
Blue Monkey
BW Colobus
Red Colobus
1970 1980 1990 2000 2010
1970 1980 1990 2000 2010
Encounter Rate
Encounter RateEncounter RateEncounter Rate
Conservation Changes
Overall, the conservation and management approaches applied to Kibale have had several
programmatic iterations which have in turn affected the attitudes and support for the park by
neighboring communities. As a Crown Forest Reserve, Kibale was established to protect
watersheds and subsequently to ensure a sustainable flow of wood products that would supply
British colonial interests in East Africa. Following independence in 1967, the control of Kibale
transitioned to the Uganda Forest Department as a Central Forest Reserve. Soon thereafter in a
tumultuous 15-year period (1971-1986), where much of the country was immersed in war and
economic meltdown, management and enforcement by the Forest Department fell lax (Hamilton
1984). Neighboring communities often entered the reserve to collect fuelwood, building poles,
medicinal plants, timber, and to hunt wildlife. Longtime residents describe the relationship as an
amicable one and Kibale as a “neighbor” and a resource pool, where “you could get things if you
needed them” (J. Magume, pers. comm. 2006).
The outbreak of rinderpest in Uganda during the 1930’s forced many people living in
Kibale forest to move towards towns (many of whom moved to Bigodi and other towns on the
eastern side of Kibale). Decades later, displaced and landless people settled in the forest reserve
during the period of civil war (Naughton-Treves 1999). Many of these people were forcibly
evicted when the Forest Department re-established its presence in 1992 and then further when the
Ugandan Wildlife Authority took over control of the Game Corridor. Those who did not leave
were forcibly removed and resettled, most 100 km away in the Kibaale District (Deneve et al.
1997). Others moved to neighboring areas along the western edge of the former game corridor
(Edmunds 1997) and east towards Kyenjojo. From their perspective, very few of these people were
compensated fairly for their loss of crops and land. Forbidding people who lived near parks to use
its resources can often create strong negative attitudes towards the park (Mugisha & Jacobson
2004). The perceived heavy-handed eviction of farmers fueled animosity between neighboring
communities and the park. To local communities who had felt Kibale was part of the landscape,
the park now sought to be exclusive and separate itself from the landscape that contained it. Little
meaningful interaction between the park and local residents took place, and most regarded the park
as something that belonged to the government (Lepp 2004; Lepp & Holland 2006).
In recent years though, the tone appears to have changed in some communities around
Kibale. The cynicism and suspicion for the park has waned. Though perceptions of and attitudes
towards the park vary, research in communities outside the park has shown that many people are
positive towards the park, or at the very least indifferent. Hartter (2007) found that 61% of the
people surveyed (n=130) felt their households benefited directly from the park. Solomon (2007)
examined fishing communities on the southwest side of Kibale and found 76% (n=183) benefited in
some way from the park.
These more positive attitudes can be attributed at least in part to a number of different
factors, both directly related and unrelated to park initiatives. First, the increasing human
population density outside Kibale led to conversion of many of the natural forests and wetlands in
the surrounding landscape, thus destroying habitat and travel corridors for would-be crop raiders.
This has increased “edge” habitat, which makes it difficult for large animals to range outside the
park without crossing agricultural land (Naughton-Treves 1998). Despite this, Hartter (2007)
quantified that only 31% of respondents felt that the park harmed their household, mostly because
of problem animals. This problem does not seem to be getting worse; only one third of people
asked felt crop raiding has become worse in the last ten years (Hartter, unpublished data).
Intensive land use by the Batoro and Bakiga has not only decreased wildlife habitat, but it has led
to many farms being buffered by other farms and tea plantations. In fact, most farmers say that the
best defense against crop raiding by wildlife is having another farm to serve as a buffer (Naughton-
Treves 1997).
Crop raiding continues to be the cause of the largest proportion of negative attitudes
towards the park, and was found to be important for households within 1 km of the park boundary
(Hartter, unpublished data). Park managers acknowledge these problems. Many of the landholders
want compensation for their losses but, according to the Uganda Wildlife Statute 1996,
compensation to individual farmers for lost crops due to crop raiding is not permitted. Even though
respondents feel that the park (52%, n = 130) and the government (31%, n = 130) must manage the
crop raiding, most communities and individuals, especially those not directly neighboring the park,
are left to develop their own deterrence strategies and means to cope with crop losses.
Second, to mitigate the effects of crop raiding and to bolster public opinion of the park’s
existence, UWA mandated a revenue sharing program. Under the Uganda Wildlife Statute of 1996,
individuals cannot be compensated for property loss or damage due to wildlife. However, 20% of
the gate receipts from national parks are to be shared with local communities (Government of
Uganda 1996). The 27 parishes that border the park are therefore eligible for a portion of the entry
fees collected from the modest amount of tourists (5-6000 annually) who visit Kibale. In 2008,
Kabarole District received Ush 36.5 million (~$23,000 USD) under the revenue sharing program
(Kajubu 2008) which was dispersed to 5 sub-counties. Park officials continue to be concerned
about the increased destruction of crops by wild animals and as such much of the money recently
has gone to dig trenches (3 m wide x 3 m deep) along the park boundary, primarily to prevent
elephant raids. In addition, these monies have gone towards infrastructure improvement projects
(e.g., classrooms, community buildings). Overall, Archabald and Naughton-Treves (2001) report
improved attitudes and friendlier relations between local residents and park employees since the
inception of revenue-sharing programs.
Third, park management has worked to improve their relationship with the neighboring
communities. As such, there is a perception that even though encroachment, resource extraction,
and hunting are prohibited, park management has become less exclusionary and confrontational,
and instead seeks to work alongside park neighbors to solve problems of crop raiding and resource
scarcity. Kibale has entered into Collaborative Resource Management (CRM) Specific Issue
agreements with local communities. These agreements allow locals to access certain park
resources under specific conditions (Chhetri et al. 2003). Agreements have been established to
allow access to rattan cane, fish, beekeeping, coffee, water for cattle, firewood, medicinal plants,
and papyrus from within the park boundaries (Chhetri et al. 2003; Solomon 2007). In addition,
rangers also help landholders by firing rifles to scare wildlife away from fields. The respectfulness
with which park managers treat their neighbors seems to have a considerable influence on attitudes
to parks (Hartter, pers. obs.). By employing more local residents, UWA hopes that economic
opportunities would have positive effects on local attitudes. Though the percentage is still small
relative to the surrounding population (Hartter 2007), Kibale employs local residents permanently
as park rangers, tourist guides, and seasonal or contracted short-term labor as trail cutters, timber
cutters, and to run the tourist canteen. The presence of the park has also drawn attention by
international researchers who also hire local residents as field assistants and cooks.
Fourth, improved education, outreach, and communication through park programs, UWA,
National Environment Management Authority, local community-based natural resource groups,
and local and international NGOs have contributed to increased local ownership of issues and
public awareness of benefits of conservation and the park beyond material benefits through public
discussion, radio broadcasts, posters, and increased interaction and training of local government
officials. Regular dialogue with communities and local government officials raised awareness of
management policies and laws. Community conservation wardens have become more visible and
have visited many communities near Kibale. Of the people who believed that the park provided
benefits to their livelihoods, Hartter’s (2007) work shows that environmental services (e.g., timing
and adequate amount of rainfall, fresh air, and climate regulation) were mentioned as benefits
more often (73%, n=79) than material benefits (e.g., employment, infrastructure development,
26%). In addition, 72% of all respondents (n=130) believe Kibale should be maintained in its
current state with the same rules and regulations and not be destroyed (Hartter 2007). Furthermore,
60% (n=183) of respondents disagreed that the park should be given to those around the park to
convert to agriculture (Solomon 2007).
The revenue sharing program, community conservation programs, increased visibility of
park staff, and cooperation with community groups have all helped to improve relations between
the park and communities. Despite the trajectory of positive public opinion, the situation around
Kibale is far from perfect, especially in communities directly adjacent to the park boundary. Crop
raiding and small-scale encroachment (e.g., timber cutting, fuelwood collection, and hunting)
continue inside the park. Money dispersed through the revenue sharing program is relatively small
and does not adequately offset the cost to individuals for loss or damage of crops by wildlife
(Archabald & Naughton-Treves 2001). Education and outreach programs tend to be focused in
areas around park headquarters, ranger outposts, the research site, and main transportation arteries.
Communities that are further removed from these activities feel less of a connection to the park
(Hartter unpublished data, Solomon 2007). The park’s community conservation unit of 3-5
individuals is severely understaffed and lacks enough financial resources to have widespread
impacts in all areas around the park; although park warden plans to these address programmatic
deficiencies (Tumwesigye pers. comm. 2008) in more remote communities. Continued
disproportionate emphasis on law enforcement illustrates that Kibale’s priority still remains law
enforcement over collective action and education. Compliance continues to be driven partially by
fear of punitive measures by rangers (Lepp & Holland 2006). Despite the shortcomings, Kibale has
remained successful in defending its borders and continues to committed to conservation and
improving the relationship between the park and its neighbors.
In general, the most significant finding from our long-term research is that whatever aspect
of the environment we examined, it was changing. This suggests that Kibale is in a non-
equilibrium state, which represents a huge challenge for the management of the park. Not only
must data be collected over the long-term to identify changes, but continual monitoring is required
to evaluate the effectiveness of management actions. Some areas of Kibale are experiencing high
levels of climate change, which may be driving changes in phenological patterns, incidence of
disease, and primate abundance.
Kibale is in an almost unrivalled position to evaluate long-term changes in primate
populations and their drivers since we have data which date back to the early 1970s. We were able
to determine that two primate species appear to be in decline: blue monkeys and red colobus (as
evaluated by encounter rate). However, despite the wealth of data on these species, we could not
identify the cause of these declines. It is thought that there are only four viable populations of red
colobus in the world, and Kibale is the only one in Uganda (
accessed December 17th 2008). As a result, there needs to be continued monitoring of red colobus
abundance and research to identify the causes of its decline.
Our findings have a number of implications. First, it is important for management
authorities to work closely with researchers willing to collect long-term data in order to document
and identify drivers of change. Second, adaptive management plans must be constructed to use
limited existing data in the most efficient way. Where possible, long-term researchers should be
encouraged to use existing data to attempt to forecast future conditions, as this will help policy
makers prevent future problems rather than responding to events that are occurring or have
occurred. It should also be recognized that taking no active management is a management decision
in itself. For example, current evidence (Lwanga 2003; Omeja et al. in press) suggests that former
grassland areas in Kibale will be replaced by forest in the coming decades, thus management
activities will be required if maintaining grasslands and the animals it supports is a priority.
While not directly part of our research program, it has become evident from over two
decades of research in the region that two additional issues need to be addressed. The Uganda
Wildlife Authority is well aware of both issues and has active programs designed to address them.
The first involves the growing elephant population in the park and increasing effect they have on
adjacent communities when they leave the park to raid crops. Research can help with this issue in a
number of ways, such as by determining elephant population levels, evaluating how their
populations will change with the regeneration of grasslands, pine plantations, and the southern
corridor, and by constructing predictive models of how their population will change with existing
and alternative management strategies.
The second issue involves the relationship between the park and the adjacent villagers.
With growing human population sizes, there will be increasing pressure on the park for resources
that become scare in the surrounding landscape. Currently, the population density of Uganda is
120 people/km2 (United Nations 2008). However, in areas with fertile lands, such as those that
surround Kibale, the population density is often between 270 and 315 people/km2 (Hartter 2007).
Naughton-Treves (1998) estimated the population density neighboring the park nearly tripled
around the park between 1959 and 1990. The density will not be overly impacted by increasing
urbanization rates, since in 1950, 3% of Uganda’s population lived in urban areas, while in 2030
the percentage of the population living in urban areas is expected to grow to only 21% (Jacob et al.
2008). This means that improved protection, enhanced park-people relations, and providing
alternative sources of fuel will all become increasingly important as the growing population
requires more resources and land.
If change in ecological systems is occurring at a rapid rate, then the few long-term
researchers that are working in an area need to work hand in hand with managers to attempt to
construct informed management plans. The most profitable avenue to do this will likely involve
scenario planning where managers and long-term researchers present potential futures and there is
an informed dialogue about the probability that different scenarios will occur and the preemptive
actions that could be taken. Accordingly, we need to use the data from long-term research to
project potential futures. These exercises should consider the appropriate scale of management
(e.g., whether or not Kibale should conserve buffalo that prefer grassland when there is thriving
populations just to the south in Queen Elizabeth National Park). A joint partnership, recognizing
the different goals and priorities of the different stakeholders is needed to forge a strong future.
Funding for the research in the 1990s and 2000s was provided by the Canada Research Chairs
Program, Wildlife Conservation Society, Natural Science and Engineering Research Council of
Canada, and National Science Foundation. Permission to conduct this research was given by the
National Council for Science and Technology and the Uganda Wildlife Authority. We would like
to extend a special thanks to the field assistants of the Kibale Fish and Monkey Project, many of
which participated in both the 1995 and 2005 censuses, and to Tusiime Lawrence who established
the vegetation plots in 1989 and helped monitor them in 2006. Helpful comments on the material
presented were given by A. Plumptre and A. Seimon.
Altmann, J., Alberts, S.C., & Roy, S.B. (2002). Dramatic change in local climate patterns in
Amboseli basin, Kenya. African Journal of Ecology, 40, 248-251.
Archabald, K., & Naughton-Treves, L. (2001). Tourism revenue sharing around national
parks in western Uganda: early efforts to identify and reward local communities.
Environmental conservation, 23, 135-149.
Ashton, P. S., Givinish, T., and Appanah, S. (1988). Staggered flowering in the
Dipterocarpaceae: new insights into floral induction and the evolution of mast
fruiting in the aseasonal tropics. American Naturalist, 132, 44-66.
Baranga, J. (1991). Kibale Forest Game Corridor: man or wildlife? In: D. A. Saunders & R. J.
Hobbs (Eds.). Nature conservation: the role of corridors. Surrey Beatty and Sons,
Brown, J. (1978). Weight and density of crowns of rocky mountain conifers, Rocky Mountain
Press, Ogden,
Butynski, T. M. (1990). Comparative Ecology of Blue Monkeys (Cercopithecus mitis) in
High-Density and Low-Density Subpopulations. Ecological Monographs, 60, 1-26.
Catchpole, W., & Wheeler, J. (1992). Estimating plant biomass: a review of techniques.
Australian Journal of Ecology, 17, 121-131.
Chapman, C. A., Chapman, L.J., Wrangham, R. W., Hunt, K., Gebo, D., & Gardner, L.
(1992). Estimators of fruit abundance of tropical trees. Biotropica, 24, 527-531.
Chapman, C. A., Wrangham, R. W., & Chapman, L.J. (1994). Indices of habitat-wide fruit
abundance in tropical forest. Biotropica, 26, 160-171.
Chapman, C. A., Wrangham, R.W., Chapman, L.J., Kennard, D.K., & Zanne, A.E. (1999).
Fruit and flower phenology at two sites in Kibale National Park, Uganda. Journal of
Tropical Ecology, 15, 189-211.
Chapman, C. A., & Lambert, J. E. (2000). Habitat alteration and the conservation of African
primates: Case study of Kibale National Park, Uganda. American Journal of
Primatology, 50, 169-185.
Chapman, C. A., Balcomb, S.R., Gillespie, T., Skorupa, J., & Struhsaker, T.T. (2000). Long-
term effects of logging on African primate communities: A 28 year comparison from
Kibale National Park, Uganda. Conservation Biology, 14, 207-217.
Chapman, C. A., & Peres, C. A. (2001). Primate conservation in the new millennium: The
role of scientists. Evolutionary Anthropology, 10, 16-33.
Chapman, C. A., & Chapman, L.J. (2002). Foraging challenges of red colobus monkeys:
influence of nutrients and secondary compounds. Comparative Biochemistry and
Physiology. Part A, Physiology, 133, 861-875.
Chapman, C. A., Chapman, L.J., Bjorndal, K. A., & Onderdonk, D. A. (2002a). Application
of protein-to-fiber ratios to predict colobine abundance on different spatial scales.
International Journal of Primatology, 23, 283-310.
Chapman, C. A., Chapman, L.J., Zanne, A., & Burgess, M. A. (2002b). Does weeding
promote regeneration of an indigenous tree community in felled pine plantations in
Uganda? Restoration Ecology, 10, 408-415.
Chapman, C. A., & Chapman, L.J. (2004). Unfavorable successional pathways and the
conservation value of logged tropical forest. Biodiversity and Conservation, 13, 2089-
Chapman, C. A., Chapman, L.J., Naughton-Treves, L., Lawes, M. J.,& McDowell, L. R.
(2004). Predicting folivorous primate abundance: Validation of a nutritional model.
American Journal of Primatology, 62, 55-69.
Chapman, C. A., Chapman, L.J., Struhsaker, T. T., Zanne, A. E., Clark, C. J., & Poulsen, J. R.
(2005). A long-term evaluation of fruiting phenology: importance of climate change.
Journal of Tropical Ecology, 21, 31-45.
Chapman, C. A., Wasserman, M. D., Gillespie, T. R., Speirs, M. L., Lawes, M. J., Saj, T.
L.,& Ziegler, T. E.. (2006). Do nutrition, parasitism, and stress have synergistic
effects on red colobus populations living in forest fragments? American Journal of
Physical Anthropology, 131, 525-534.
Chapman, C. A., Kitajima, K., Zann, A. E., Kaufman, L. S.,& Lawes, M. J. (2008). A 10-yr
evaluation of the functional basis for regeneration habitat preference of trees in an
African evergreen forest. Forest Ecology and Management, 225, 3790-3796.
Chapman, C.A., Chapman, L.J., Jacob, A.L., Rothman, J.M., Omeja, P., Reyna-Hurtado, R.,
Hartter, J., & Lawes, M.J.. (2010a). Tropical tree community shifts: Implications for
wildlife conservation. Biological Conservation, 143, 366-374.
Chapman, C.A., Struhsaker, T.T., Skorupa, J.P., Snaith, T.V., & Rothman, J.M. (2010b).
Understanding long-term primate community dynamics: Implications of forest
change. Ecological Applications, 20, 179-191.
Chapman, C. A., Speirs, M. L., Hodder, S. A. M., & Rothman, J. M.. 2010. Colobus monkey
parasite infections in wet and dry habitats: implications for climate change. African
Journal of Ecology 48:555-558.
Chhetri, P., Mugisha, A., & White, S. (2003). Community resources use in Kibale and Mt.
Elgon National Parks, Uganda. Parks, 13, 28-49.
Coop, R. L., & Holmes, P.H. (1996). Nutrition and parasite interaction. International Journal
of Parasitology, 26, 951-962.
Deneve, R., Odwedo, M., Okwakol, R., & Thies, A. (1997). National Parks. To be or not to
be? Kibale Semliki Conservation and Development Project, Kampala.
Duncan, R.S.,& Chapman, C.A. (2003). Consequences of plantation harvest during tropical
forest restoration in Uganda. Forest Ecology And Management, 173, 235-250.
Edmunds, D. (1997). Continuity and change in the resource management institutions of
communities bordering the Kibale forest Park, Uganda. Clark University, Worcester.
FAO (2005). Global Forest Resources Assessment 2005: progress towards sustainable forest
management. FAO Forestry Paper 147, Rome.
Ganzhorn, J. U. (2002). Distribution of a folivorous lemur in relation to seasonally varying
food resources: Integrating quantitative and qualitative aspects of food
characteristics. Oecologia, 131, 427-435.
Government of Uganda (1996). The Uganda Wildlife Act. Kampala.
Hamilton, A. C. (1974). Distribution patterns of forest trees in Uganda and their historical
significance. Vegatatio, 29, 21-35.
Hamilton, A. C. (1984). Deforestation in Uganda. Oxford University Press, Oxford.
Hamilton, A. Taylor C., D., & Vogel, J. C. (1986). Early forest clearance and environmental
degradation in south-west Uganda. Nature, 320, 164-167.
Hannah, L., Midgley, G.F., Lovejoy, T., Bond, W. J., Bush, M., Lovett, J. C., Scott, D., &
Woodwards, F.L. (2002). Conservation of biodiversity in a changing climate.
Conservation Biology, 16, 264-268.
Harrington, G. (1979). Estimation of above-ground biomass of trees and shrubs. Australian
Journal of Botany, 27, 135-143.
Harris, T. R., & Chapman, C.A.. (2007). Variation in the diet and ranging behavior of black-
and-white colobus monkeys: Implications for theory and conservation. Primates, 28,
Hartter, J. (2007). Landscape changes around Kibale National Park, Uganda: Impacts on
land cover, land use and livelihoods. Ph.D. University of Florida, Gainesville.
Houghton, R.A. (1994). The worldwide extent of land-use change. Bioscience, 44, 305-313.
Howard, P. C. (1991). Nature conservation in Uganda's tropical forest reserves. IUCN,
Gland, Switzerland.
Howard, P. C., Davenport, T.R.B., Kigenyi, F.W., Viskanic, P., Balzer, M.C., Dickinson,
C.J., Lwanga, J.S., Matthews, R.A.,& Mupada, E. (2000). Protected area planning in
the tropics: Uganda's national system of forest nature reserves. Conservation Biology,
14, 858-875.
Hunter, P.R. (2003). Climate change and waterborne and vector-borne disease. Journal of
Applied Microbiology, 94, 37S-46S.
IPCC (2001). Climate Change 2001: The scientific basis. Cambridge University Press,
Jacob, A.L., Vaccaro, I., Sengupta, R., Hartter, J., & Chapman, C.A.. (2008). How can
conservation biology best prepare for declining rural population and ecological
homogenization? Tropical Conservation Science, 4, 307-320.
Kajubu, E. (2008). Kibale give Kabarole 36 m sh. New Vision.
Kaumi, S.Y.S. (1989). Rehabilitation of industrial plantations: forestry rehabilitation project.
Ministry of Environmental Protection, Kampala.
Kingston, B. (1967). Working plan for Kibale and Itwara Central Forest Reserves. Uganda
Forestry Department, Entebbe, Uganda.
Lang Brown and Harrop 1962
Lawes, M. J., & Chapman, C.A. (2005). Does the herb Acanthus pubescens and / or elephants
suppress tree regeneration in disturbed Afrotropical forests? Forest ecology and
management, 221, 274-284.
Lepp, A.P. (2004). Tourism in a rural Ugandan village: impacts, local meaning and
implications for development. Ph.D. University of Florida, Gainesville.
Lepp, A.P., & Holland, S. (2006). A comparison of attitudes toward state-led conservation
and community-based conservation in the village of Bigodi, Uganda. Society and
Natural Resources 19, 609-623.
Livingstone, D.A. (1967). Postglacial vegetation of the Ruwenzori Mountains in equatorial
Africa. Ecological Monographs, 37, 25-52.
Lwanga, J.S. (2003). Forest succession in Kibale National Park, Uganda: implications for
forest restoration and management. African Journal of Ecology, 41, 9-22.
Lwanga, J.S., Butynski, T.M., & Struhsaker, T. T. (2000). Tree population dynamics in
Kibale National Park, Uganda 1975-1998. African Journal of Ecology, 38,238-274.
McClean, C.J., Lovett, J.C., Kuper, W., Hannah, L., Sommer, J.H., Barthlott, W., Termansen,
M., Smith, G.F., Tokumine, S., & Taplin, J.R.D. (2005). African plant diversity and
climate change. Annals of the Missouri Botanical Gardens, 92, 139-152.
Milton, K. (1979). Factors influencing leaf choice by howler monkeys: a test of some
hypotheses of food selection by generalist herbivores. American Naturalist, 114, 363-
MISR (Makerere University Institute for Social Research) (1989). Settlement in Forest
Reserves, Game Reserves, and National Parks. Makerere University, Kampala,
Mugisha, A.R., & Jacobson, S.K. (2004). Threat reduction assessment of conventional and
community-based conservation approaches to managing protected areas in Uganda.
Environmental conservation, 31, 233-241.
Munger, J.C., & Karasov, W.H. (1989). Subleathal parasites and host energy budgets:
Tapeworm infection in white-footed mice. Ecology, 70, 904-921.
Naughton-Treves, L. (1997). Farming the forest edge: Vulnerable places and people around
Kibale National Park, Uganda. Geographical Review, 87, 27-49.
Naughton-Treves, L. (1998). Predicting patterns of crop damage by wildlife around Kibale
National Park, Uganda. Conservation Biology, 12,156-168.
Naughton-Treves, L. (1999). Whose Animals? A history of property rights to wildlife in Toro,
western Uganda. Land Degradation and Development, 10, 311-328.
NEMA (1997). National Environmental Authority Report: Kabarole District Environment
Profile. NEMA, Kampala, Uganda.
Newbery, D.M., Songwe, N.C., & Chuyong, G.B. (1998). Phenology and dynamics of an
African rainforest at Korup, Cameroon. In: D.M. Newbery, H. H. T. Prins, & Brown,
N.D. (Eds.). Dynamics of tropical communities. Blackwell, Oxford.
Oates, J.F. (1977). The guereza and its food. In: T. H. Clutton-Brock (Ed.). Primate Ecology.
Academic Press, New York. Pages 275-321
Oates, J.F., Whitesides, G.H., Davies, A.G., Waterman, P.G., DaSilva, G.L., & Mole, S.
(1990). Determinants of variation in tropical forest primate biomass: New evidence
from West Africa. Ecology, 71, 328-343.
Olupot, W. (1994). Ranging patterns of the grey-cheeked mangabey Cercocebus albigena
with special reference to food finding and food availability in Kibale National Park.
M.SC. Makerere University, Kampala.
Omeja, P.A., Chapman, C.A., & Obua, J. (2009). Enrichment planting does not promote
native tropical tree restoration in a former pine plantation. African Journal of
Ecology, 45, 650-657.
Omeja, P.A., Obua, J., & Chapman, C.A. (in prep). Intensive tree planting for forest
restoration: A case study of UWA -FACE Project, Kibale National Park, Uganda.
Opler, P.A., Frankie, G.W., & Baker, H.G.. (1976). Rainfall as a factor in the release, timing,
and synchronization of anthesis by tropical trees and shrubs. Journal of
Biogeography, 3, 231-236.
Osmaston, H. A. (1959). Working plan for the Kibale and Itwara Forests. Ugandan Forest
Department, Entebbe.
Parmesan, C., & Yohe, G.A. (2003). A globally coherent fingerprint of climate change
impacts across natural systems. Nature, 421, 37-42.
Patz, J., Epetein, P., Burke, T., & Balbus, J. (1996). Global climate change and emerging
infectious diseases. Journal of the American Medical Association, 275, 217-223.
Paul, J.R., Randle, A.M., Chapman, C.A., & Chapman, L.J. (2004). Arrested succession in
logging gaps: is tree seedling growth and survival limiting? African Journal of
Ecology, 42, 245-251.
Pomeroy, D., & Tushabe, H. (2004). The state of Uganda's Biodiversity 2004. Makerere
University, Kampala.
Potts, M. (2007). Population and environment in the twenty-first century. Population and
Environment, 28, 204-211.
Potvin, C., Lechowicz, M.J., & Tardif, S. (1990). The statistical-analysis of ecophysiological
response curves obtained from experiments involving repeated measures. Ecology,
71, 1389-1400.
Pounds, J.A., Fogden, M.P.L., & Campbell, J.H. (1999). Biological response to climate
change on a tropical mountain. Nature, 398, 611-615.
Rode, K.D., Chapman, C.A., McDowell, L.R., & Stickler, C. (2006). Nutritional correlates of
population density across habitats and logging intensities in redtail monkeys
(Cercopithecus ascanius). Biotropica, 38, 625-634.
Rudran, R. (1978). Socioecology of the Blue Monkeys (Cercopithecus mitis stuhlmanni) of
the Kibale Forest, Uganda. Smithsonian Contributions to Zoology, 249, 88.
Skorupa, J.P. (1988). The effect of selective timber harvesting on rain forest primates in
Kibale Forest, Uganda. Unpublished PhD Thesis, University of California, Davis.
Solomon, J. (2007). An evaluation of collaborative resource management and the
measurement of illegal resource use in a Ugandan National Park. Ph.D. University
of Florida, Gainesville.
Stampone, M., Hartter, J., Chapman, C.A., Ryan, S.J. (in prep.). Localized precipitation
trends in and around a forest park in Equatorial East Africa.
Stickler, C.M. (2004). The effects of logging on primate-habitat interactions: A case study of
redtail monkeys (Cercopithecus ascanius) in Kibale National Park, Uganda.
Unpublished M.Sc.Thesis, University of Florida, Gainesville, Florida.
Stoner, K. (1996). Prevalence and intensity of intestinal parasites in mantled howling
monkeys (Alouatta palliata) in northeastern Costa Rica: Implications for conservation
biology. Conservation Biology, 10, 539-546.
Struhsaker, T.T. (1975). The Red Colobus Monkey. University of Chicago Press, Chicago.
Struhsaker, T.T. (1997). Ecology of an African rain forest: logging in Kibale and the conflict
between conservation and exploitation. University of Florida Press, Gainesville.
Struhsaker, T.T., Lwanga, J.S., & Kasenene, J.M. (1996). Elephants, selective logging, and
forest regeneration in the Kibale Forest, Uganda. Journal of Tropical Ecology, 12, 45-
Taylor, D., Marchant, R.A., & Robertshaw, P. (1999). A sediment-based history of medium
altitude forest in central Africa: A record from Kabata Swamp, Ndale volcanic field,
Uganda. Journal of Ecology, 87, 303-315.
Taylor, D., Robertshaw, P., & Marchant, R.A. (2000). Environmental change and political
economic upheaval in precolonial western Uganda. The Holocene, 10, 527-536.
United Nations (2005). World population prospects: The 2004 revision. United Nations, New
United Nations. (2008). World Urbanization Prospects: The 2007 Revision. CD-ROM
Edition Data in digital form (POP/DB/WUP/Rev.2007) Department of Economic
and Social Affairs/Population Division. New York, NY.
UWA-FACE (2005). Project plan of operation report - January to December 2005. Uganda
Wildlife Authority, Kampala.
van Orsdol, K.G. (1986). Agricultural encroachment in Uganda's Kibale Forest. Oryx, 20,
van Schaik, C.P. (1986). Phenological changes in a Sumatran rain forest. Journal of Tropical
Ecology, 2, 327-347.
van Schaik, C.P., Terborgh, J.W., & Wright, S. J. (1993). The phenology of tropical forests:
adaptive significance and consequences for primary consumers. Annual Review of
Ecology and Systematics, 24, 353-377.
van Vliet, A., & Schwartz, M.D. (2002). Phenology and climate: the timing of life cycle
events as indicators of climate variability and change. International Journal of
Climatology, 22, 1713-1714.
Walther, G.R., Post, E., Convey, P., Menzel, A., Parmesan, C., Beebee, T.J.C., Fromentin,
J.M., Hoegh-Guldberg, O., & Bairlein, F. (2002). Ecological responses to recent
climate change. Nature, 416, 389-395.
Waser, P. (1975). Monthly variations in feeding and activity patterns of the mangabey,
Cercocebus albigena (Lydekker). East African Wildlife Journal, 13, 249-263.
Waterman, P.G., Ross, J.A.M., Bennett, E.L., & Davies, A.G. (1988). A comparison of the
floristics and leaf chemistry of the tree flora in two Malaysian rain forests and the
influence of leaf chemistry on populations of colobine monkeys in the old world.
Biological Journal of the Linnean Society, 34, 1-32.
Zanne, A. E., & Chapman, C.A. (2005). Diversity of woody species in forest, treefall gaps,
and edge in Kibale National Park, Uganda. Plant Ecology, 178, 121-139.
... Rainfall and temperature has been studied over a 100-year time period near Kibale National Park, both at MUBFS and in Fort Portal Town (Chapman et al. 2005a). Analysis of the data show that there has been an increase in both rainfall and maximum monthly temperature over time with as much as a 4 o C rise in maximum temperatures over 33 years (Chapman et al. 2005b;Chapman et al. 2012). Much of this increase in temperature is attributed to deforestation and drainage of wetlands around the Kibale National Park rather than global climate change, but it shows that the impacts of climate change will be significantly exacerbated by land cover conversion to agriculture. ...
Technical Report
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This report summarises the in-depth study of existing knowledge of changes in animal behavior in the Greater Virunga Landscape (GVL) resulting from industrial development, climate changes and other factors as well as the assessment of changes in ecology resulting from habitat changes across the GVL. The objectives of this study were specifically to 1) assess the actual and potential changes in animal behavior resulting from Climate Change; 2) assess the actual and potential long-term changes in behavior resulting from oil/gas and geothermal exploration and production; 3) assess other factors (beside development and climate change impacts) that are or will affect animal behavior in the GVL; 4) map land cover changes over time in the GVL from 2000-2015/6; and 5) provide recommendations for mitigating adverse changes in animal behavior in the GVL. From the study, the results show that climate change impacts area already being seen in the GVL, including movement upslope of birds on the mountains, increasing rainfall as predicted by climate models, and increasing temperatures. Phenology patterns in forests may is also being affected by increasing temperatures, particularly for those trees where minimum temperatures may trigger flowering. Models of predicted impacts of climate change on the endemic and threatened species (108) in the GVL indicate that significant range reductions will likely occur resulting from the need to move upslope to maintain similar niches. On average species ranges have declined by 55% because of loss of habitat to agriculture and infrastructure developments in the region around the GVL and an additional 70% of what remains will be lost to climate changes by 2070, if the projected emission scenarios follow the trajectory under RCP 8.5. The bird species that have lost more than 65% of their current range include the Blue headed sunbird, Grauers rush warbler, Montane masked apalis, Neumanns warbler, Purple-breasted Sunbird, Red-faced woodland warbler, Red throated Alethe, Rwenzori batis, Handsome francolin, Dusky crimson-wing, Kivu ground thrush,Yellow-eyed black_flycatcher, Chapin's flycatcher, Dwarf honey guide, Ruwenzori nightjar, Shelleys crimson-wing, Grauers Rush_Warbler, Chapin's Flycatcher, Shelleys Crimson-wing, Grey Crowned Crane and the plants are Allanblackia kimbiliensis, Coccinia mildbraedii, Diplazium_humbertii Embelia libeniana, Grewia mildbraedii, Harungana montana, Impatiens mildbraedii, Melchiora schliebenii,Musanga leo-errerae, Oxyanthus troupinii, Rubus kirungensis, Rytigynia bridsoniae, Rytigynia kigeziensis, Thunbergia mildbraediana,Monanthotaxis orophila, Pavetta pierlotii,Tabernaemontana odoratissima, Beilschmiedia_ugandensis ,Carex runssoroensis, Casearia engleri, Entandrophragma angolense, Entandrophragma cylindricum, Entandrophragma excelsum, Helichrysum formosissimum, Lovoa swynnertonii, Lovoa trichilioides, Ocotea kenyensis, Prunus africana, Turraeanthus africanus. Among the large mammal species, Loxodonta africana, Hippopotamus amphibius, Cercopithecus lhoesti, Pan troglodytes, Cercopithecus hamlyni, Gorilla beringei are currently being impacted by climate change. Fire frequency is also increasing, which may be a result of increasing fuel load with the wetter climate. These factors will likely lead to changes in vegetation composition as certain plants become more competitive with increasing moisture and fire frequency which will in turn affect animal distributions in the savannas of the GVL in particular. The areas in the GVL where the most species are likely to persist include Rwenzori national park, Mgahinga national park and Virunga volcanoes. These areas are important for the future conservation of endemic species in the GVL. Mapping of the 2017 land cover land use showed that most of the protected areas are still covered by natural vegetation. The north western part of the landscape, however, under conversion to degraded forest, which is a mixture of small-scale agriculture and woodlands. Assessing land cover land use change showed that the grassland cover registered the highest net loss by 33% followed by wooded grassland at 29%. Natural land cover increase was highest in woodland and tree plantation land cover categories with 41% and 22% respectively. These changes were further emphasized when overall woody cover was assessed. Woody cover of QENP registered a 25% increase between 1954 and 2006. In terms of the entire landscape, there was a 14% increase in woody cover between 2006 and 2017, which is a slightly higher rate of loss over 11 years compared to the period of 52 years. Agriculture also gained 10% coverage within the same period. Overall, there has been a higher increase in woody cover inside protected areas due to reduced mammal densities and a higher decrease outside protected areas attributed to agricultural expansion, infrastructure development and human settlement. Savanna parks are becoming more woody, which may pose a threat to the survival of angulates that occur in GVL. The key drivers of land cover land use change need to be fully addressed by the regional governments under the transboundary collaboration framework already in place. On the other hand, industrial developments have the potential to affect wildlife behaviour and examples are given of the impacts of oil exploration in Murchison Falls NP on elephant behaviour, impacts of the cobalt mines near Rwenzori Mountains NP on vegetation, impacts of the hydropower dam at Mpanga on Cycads and potential impacts of the proposed cable car to the top of Karisimbi volcanoe in the Volcanoes Park. Ideally, industrial developments would not take place in the GVL and all efforts should be made to avoid having a negative impact on the natural habitats and wildlife in the landscape by using directionally directed access from outside the landscape. Where avoidance is impossible and where there is no option to stop the development then the mitigation hierarchy should be followed and the key principles of this concept, that is, additionality, permanence, equivalences should be upheld. The company should go beyond simple compensation, but work toward achieving a net positive impact. This is because this is one of the most biodiverse places on Earth, has three World Heritage Sites and 10 Key Biodiversity Areas. We noted that poaching pressure on wildlife will also have an impact on where animals range and we showed that poaching pressure varies across the landscape using ranger patrol data from SMART in Uganda and DRC. Snaring is often found near the park boundaries, but killing of larger animals such as elephants, hippos and buffalo occur in the central areas of the parks because they have been extirpated near most park boundaries. Declines of species to small populations is of real concern and the case of the “Tree climbing lions of Uganda” is highlighted as they have fluctuated greatly between 10-33 individuals over the time they have been monitored. Given that wild animals will continue to move out of protected areas into the communities either as part of their normal movement and dispersal routine, running away from extractive industries activities taking place inside the parks or simply responding to availability of food outside protected areas, a comprehensive human-wildlife conflict plan needs to be developed and communities living adjacent to these protected areas need to be supported in order to be part of the conservation agenda. As such, a performance-based reward system needs to be developed in order to incentivize communities to participate in conservation.
... This decline is mostly found in the slower growing NPLD and shadebearer species (Fig. 3) and mostly those trees that produce fleshy fruit (Fig. S6). Chapman et al. (2012) documented decreasing fruit production between 1970 and 1984 from one dataset followed by increasing production between 1990 and 2002 from another set of trees measured in the same part of Kibale National Park in western Uganda, but it is unclear why there are these long-term trends and differences between these two periods. Climate is likely to be an important factor in triggering flowering (Wright et al. 1999, Chapman et al. 2005b, Wright & Calder on 2006. ...
The occurrence of flowering and fruiting in tropical trees will be affected by a variety of factors, linked to availability of resources and suitable climatic triggers, that may be affected by increasing global temperatures. Community-wide flowering and fruiting of 2,526 trees in 206 plots were monitored over 24 years in the Budongo Forest Reserve (BFR), Uganda. Factors that were assessed included: the size of the tree, access to light, the impacts of liana load, effects of tree growth and variation between guilds of trees. Most flowering occurs at the end of the long dry season from February to April. Trees that had access to more light flowered and fruited more frequently. Pioneer and non-pioneer light demanding species tended to reproduce more frequently than shade-bearing species. Trees that grew faster between 1993-2011 also fruited more frequently. When examining all factors, growth rate, tree size, and crown position were allimportant for fruiting, while liana load but not growth rate was important in reducing flowering. Trees in BFR show a large decline in fruiting over 24 years, particularly in non-pioneer light demanders, shade-bearers, and species that produce fleshy fruits eaten by primates. The decline in fruit production is of concern and is having impacts on primate diets and potential recruitment of mahogany trees. Whether climate change is responsible is unclear but flowering of the guilds/dispersal types which show declines is correlated with months with the coolest maximum temperatures and we show temperature has been increasing in BFR since the early 1990s.
... Climate data from Kibale shows long-term, temporal variation in rainfall and temperature. Longer drought events, more time between rainfall events, and a warming trend have been identified (Hartter et al., 2012;Saulnier-Talbot et al., 2014) and an increase in rainfall suggested (Chapman et al., 2011). The development and survival of Trichuris sp., in the environment is negatively affected by high temperatures (max survival at 37-40 • C), temperature variation (Chammartin et al., 2014), and low relative humidity (Pullan and Brooker, 2012;Weaver et al., 2010). ...
... Our study also critically highlights the importance of defining ecological characteristics across available habitat types within an anthropogenic landscape and the need to monitor these as landscape characteristics change over time with shifts in climate and land use patterns [Chapman et al., 2011]. Evidently each available habitat type is not equally important in terms of resource availability for the chimpanzees. ...
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Many primate populations occur outside protected areas in fragmented anthropogenic landscapes. Empirical data on the ecological characteristics that define an anthropogenic landscape are urgently required if conservation initiatives in such environments are to succeed. The main objective of our study was to determine the composition and availability of chimpanzee (Pan troglodytes verus) food resources across fine spatial scales in the anthropogenic landscape of Bossou, Guinea, West Africa. We examined food resources in all habitat types available in the chimpanzees' core area. We surveyed resource composition, structure and heterogeneity (20 m × 20 m quadrats, N = 54) and assessed temporal availability of food from phenology trails (total distance 5951 m; 1073 individual trees) over 1 year (2012-2013). Over half of Bossou consists of regenerating forest and is highly diverse in terms of chimpanzee food species; large fruit bearing trees are rare and confined to primary and riverine forest. Moraceae (mulberries and figs) was the dominant family, trees of which produce drupaceous fruits favored by chimpanzees. The oil palm occurs at high densities throughout and is the only species found in all habitat types except primary forest. Our data suggest that the high densities of oil palm and fig trees, along with abundant terrestrial herbaceous vegetation and cultivars, are able to provide the chimpanzees with widely available resources, compensating for the scarcity of large fruit trees. A significant difference was found between habitat types in stem density/ha and basal area m(2) /ha of chimpanzee food species. Secondary, young secondary, and primary forest emerged as the most important habitat types for availability of food tree species. Our study emphasizes the importance of examining ecological characteristics of an anthropogenic landscape as each available habitat type is unlikely to be equally important in terms of spatial and temporal availability of resources. Am. J. Primatol. © 2016 Wiley Periodicals, Inc.
... As early as 1971, illegal destruction and encroachment occurred in the corridor, which led to forest conversion to farms. The estimates of people who settled in the corridor varies considerably (8,800–170,000), but all were evicted in 1993 (Chapman et al. 2010a). Outside Kibale, a mosaic of small farms (most \5 ha in size), large tea estates ([200 ha), and Landscape Ecol (2011) 26:877–890 879 interspersed forest fragments and wetlands, effectively isolate the park from other tracts of forest. ...
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Kibale National Park, within the Albertine Rift, is known for its rich biodiversity. High human population density and agricultural conversion in the surrounding landscape have created enormous resource pressure on forest fragments outside the park. Kibale presents a complex protected forest landscape comprising intact forest inside the park, logged areas inside the park, a game corridor with degraded forest, and forest fragments in the landscape surrounding the park. To explore the effect of these different levels of forest management and protection over time, we assessed forest change over the previous three decades, using both discrete and continuous data analyses of satellite imagery. Park boundaries have remained fairly intact and forest cover has been maintained or increased inside the park, while there has been a high level of deforestation in the landscape surrounding the park. While absolute changes in land cover are important changes in vegetation productivity, within land cover classes are often more telling of longer term changes and future directions of change. The park has lower Normalized Difference Vegetation Index (NDVI) values than the forest fragments outside the park and the formerly logged area—probably due to forest regeneration and early succession stage. The corridor region has lower productivity, which is surprising given this is also a newer regrowth region and so should be similar to the logged and forest fragments. Overall, concern can be raised for the future trajectory of this park. Although forest cover has been maintained, forest health may be an issue, which for future management, climate change, biodiversity, and increased human pressure may signify troubling signs. KeywordsNormalized Difference Vegetation Index (NDVI)–Kibale National Park–Forest fragments–Land cover change–Continuous analyses–Reforestation–Regrowth
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Primatus sum, nihil primatum mihi alienum puto (Yo soy un primate; Nada acerca de los primates me es extraño [Hooton, 1955]). La primatología, como la antropología no tiene límites intelectuales, todo acerca de los primates está incluida en ella (Rodman, 1999). Los antropólogos le han dado poca importancia a los Primates del Nuevo Mundo porque no están en la tendencia principal de estudio de la evolución de los homínidos. Sin embargo, los platirrinos proveen ejemplos excelentes de paralelismos que ayudan a esclarecer principios generales para todos los primates (Kinzey, 1997)
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In tropical regions, most primary ecosystems have been replaced by mosaic landscapes in which species must cope with a large shift in the distribution of their habitat and associated food resources. Primates are particularly vulnerable to habitat modifications. Most species persist in small fragments surrounded by complex human-mediated matrices whose structure and connectivity may strongly influence their dispersal and feeding behavior. Behavioral plasticity appears to be a crucial parameter governing the ability of organisms to exploit the resources offered by new matrix habitats and thus to persist in fragmented habitats. In this study, we were interested in the dietary plasticity of the golden-crowned sifaka (Propithecus tattersalli), an endangered species of lemur, found only in the Daraina region in north-eastern Madagascar. We used a DNA-based approach combining the barcoding concept and Illumina next-generation sequencing to (i) describe the species diet across its entire range and (ii) evaluate the influence of landscape heterogeneity on diet diversity and composition. Faeces from 96 individuals were sampled across the entire species range and their contents were analyzed using the trnL metabarcoding approach. In parallel, we built a large DNA reference database based on a checklist of the plant species of the Daraina region. Our results suggest that golden-crowned sifakas exhibit remarkable dietary diversity with at least 130 plant species belonging to 80 genera and 49 different families. We highlighted an influence of both habitat type and openness on diet composition suggesting a high flexibility of foraging strategies. Moreover, we observed the presence of numerous cultivated and naturalized plants in the faeces of groups living in forest edge areas. Overall, our findings support our initial expectation that P. tattersalli is able to cope with the current level of alteration of the landscape and confirm our previous results on the distribution and the dispersal ability of this species.
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Two subpopulations of blue monkeys (Cercopithecus mitis), located 10 km apart, were studied in Kibale Forest, Uganda, for most of a 6-yr period (1978-1984). This study was undertaken (1) to assess and evaluate the major differences in the environment and in the behavioral ecology of blue monkeys living at high and low densities and (2) to explain the differences in blue monkey densities in the two study sites. Methods included the enumeration of trees and primates, and assessments of the availability of fruits and of diets, time budgets, ranging patterns, and demographics of blue monkeys. There was a 10-fold difference in blue monkey densities between the two subpopulations. Major differences were found in the ecology, behavior, demography, and habitats of blue monkeys living at these two densities. Compared to the high-density subpopulation, the low-density subpopulation exhibited shorter tenure lengths for resident males, a greater density of nonresident males, a higher rate of group intrusions by nonresident males, a higher incidence of infanticide by new resident males, hybridization between blue monkeys and redtail monkeys (Cercopithecus ascanius),fewer adult females per adult male, and lower resident male reproductive success. From these observations I conclude that male-male competition for females was more intense where blue monkey densities were low. Predictor variables of food production and food competition indicate that more food was available for the low-density than for the high-density subpopulation of blue monkeys. The habitat containing the low-density subpopulation had a higher tree density and basal cover, greater tree species richness and diversity, and more fruit left uneaten. It also had a lower overall primate density, which, together with assessments of the kinds of foods eaten, dietary overlap and richness, foraging behavior, and home range size, suggests that intra- and interspecific competition for food both were less in the low-density subpopulation. Further evidence that food was more available for the low-density subpopulation is that both the birth rate and the population growth rate of blue monkeys were higher there than in the high-density subpopulations. The low-density subpopulation was apparently below carrying capacity and increasing in a food-rich habitat. In contrast, the high-density subpopulation appeared to be at carrying capacity, stable, and food limited. Therefore, contrary to what was hypothesized, the availability of food does not explain why blue monkey densities differed on these two areas. On this basis I attribute the low density of the one subpopulation to some unknown historical event rather than to current ecological differences between areas. I suggest that, during this study, the low-density subpopulation was recovering from a decline and that the responsible mechanism (e.g., disease) was no longer operating. This paper emphasizes: (1) the considerable variability found in the ecology and behavior of primates--even within one species in the same forest; (2) the need for long-term comparative studies of free-living primates, especially those at low densities; (3) the importance of investigating the density of primate populations relative to the carrying capacities of their environments, and the influence of this relationship on behavior and ecology.
The author summarizes 20 years of research in the Kibale forest in Uganda. The main body of the book demonstrates the adverse effects of logging on community structure and other aspects of forest ecology. The author provides evidence that future logging must be done at far lower intensities than is currently the norm, if intact ecosystems are to be maintained. Detailed recommendations for harvest plans compatible with the conservation of biodiversity and ecological integrity are outlined. Struhsaker addresses the underlying causes of tropical deforestation and concludes that although there are numerous proximate factors, the ultimate causes are rapidly increasing human populations and rates of consumption per capita. Comparisons with relevant studies elsewhere in the tropics are drawn and specific recommendations to address the problems are offered.
The Kibale Forest Corridor Game Reserve lies between the Queen Elizabeth National Park (QENP) and Kibale Forest Reserve in Uganda. The Game Corridor was gazetted in 1926 to allow for the free movement of animals. Since the early 1970s forest resources have been depleted rapidly and encroachers have moved into the Game Corridor and settled. At the same time elephant Loxodonta africana numbers went from 3000 in 1973 to 500 in 1989 in the QENP. The Kibale Game Corridor should be retained as a conservation area and the trend of encroachment should be reserved. -from Author
The assessment of fruit abundance is critical for studies of frugivore ecology A variety of methods have been used to estimate habitat-wide fruit abundance. However, since the methods have not been calibrated with each other, it is difficult to compare results of different studies. Here we compare three methods used simultaneously to collect fruit abundance data in the Kibale Forest, Uganda. Estimates of fruit abundance derived from fruit traps were not correlated with estimates derived from either systematic transect sampling or estimates obtained from observing fruiting phenology of key species on a fruit trail. However, estimates based on fruit trail data and transect data were correlated. We review the advantages and disadvantages of methods that have been used to assess habitat-wide fruit abundance.
One way to study a parasite's effect on the individual and population ecology of its host is to examine effects on the host's energy budget. A relatively innocuous effect of a gut parasite, such as decreased digestive efficiency, can potentially translate into an effect costly to host fitness, such as decreased reproduction, if other compensations (such as increased rate of food intake) do not occur. We found that infection by the tapeworm Hymenolepis citelli caused a 2% drop in dry-matter digestibility in host white-footed mice (Peromyscus leucopus). A crowding experiment indicated that this result should be applicable to a wide range of intensities of infection. However, we detected no compensation for this decreased digestive efficiency either in amount of food consumed or in mass change (which would indicate use of fat stores or changes in growth). In field experiments we used doubly labeled water to measure effects of tapeworms on field metabolic rate and water influx (potentially a measure of food intake rate), and temperature-sensitive transmitters to measure body temperature. We detected no compensations via these routes either. Our failure to detect compensations indicates that in the nonreproductive mice studied the decrease in digestive efficiency is not of sufficient importance to engender substantial compensations and is therefore unlikely to lead to fitness-altering effects. It is in reproductive animals or in animals subjected to food shortage that such effects would be expected.
Although a number of studies have described seasonality of flowering by tropical plants (most often trees) either at the species or community level, for tropical plants there have been few studies treating the induction of reproductive structures, their ensuing development, or the triggering of anthesis (opening of the flower). In this paper we present data which demonstrates that rainfall, either directly or indirectly, is an important timing and spacing mechanism for the flowering of at least some tropical plants. We also attempt to relate this phenomenon to some prior relevant studies on flowering in the tropics.
To explore sources of variation in tropical forest primate biomass, and, in particular, to test the hypothesis that soil conditions are a major ultimate determinant of the biomass of colobine monkeys and other primates, we compared data on the soils, vegetation, and primate community at a site in West Africa (Tiwai Island, Sierra Leone) with information from other sites, especially two other African sites (Douala-Edea in Cameroon, and Kibale Forest in Uganda). The biomass of eight anthropoid primate species in old secondary high forest on Tiwai was estimated from data on population densities assessed by transect samples combined with data on social group densities and individual body masses. Samples of soil and tree foliage were collected at the same site, and subjected to a variety of chemical and mechanical analyses. Our estimate of anthropoid biomass at Tiwai is 1229-1529 kg/km^2, including 682-889 kg/km^2 of colobines. This is one of the highest primate biomasses recorded anywhere. The soils at Tiwai were found to be relatively high in sand content and low in pH, and to have low levels of mineral nutrients. Levels of condensed tannins in the mature foliage of the trees comprising a major part of the forest canopy were higher than at other sites, but the ratio of protein to fiber in this foliage was also higher than at any other site except Kibale. It is argued that a wide range of environmental factors affect primate population densities, and that nutrient-poor soils and high tannin levels in tree foliage do not necessarily produce a low primate (or colobine) biomass, as some earlier studies had suggested. Furthermore, seeds (an important food source for Tiwai colobines) are apparently a common part of the colobine dietary repertoire and are not consumed largely as a response to a scarcity of digestible foliage.