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Forest Ecology and Management 261 (2011) 703–709
Contents lists available at ScienceDirect
Forest Ecology and Management
journal homepage: www.elsevier.com/locate/foreco
Intensive tree planting facilitates tropical forest biodiversity and biomass
accumulation in Kibale National Park, Uganda
Patrick A. Omejaa,b, Colin A. Chapmanb,c,d,∗, Joseph Obuaa, Jeremiah S. Lwangab, Aerin L. Jacobe,
Frederick Wanyamaf, Richard Mugenyif
aFaculty of Forestry and Nature Conservation, Makerere University, P.O. Box 7062, Kampala, Uganda
bMakerere University Biological Field Station, P.O. Box 967, Fort Portal, Uganda
cMcGill School of Environment, 855 Sherbrooke St. West, McGill University, Montreal, Canada H3A 2T7
dWildlife Conservation Society, 2300 Southern Boulevard, Bronx, NY 10460, USA
eDepartment of Biology, McGill University, 1205 Docteur Penfield, Montreal, Canada H3A 1B1
fUganda Wildlife Authority, P.O. Box 3530, Kampala, Uganda
article info
Article history:
Received 19 September 2010
Received in revised form
24 November 2010
Accepted 26 November 2010
Available online 24 December 2010
Keywords:
Anthropogenic disturbance
Biomass accumulation
Carbon offset
Tree planting
Restoration
Regeneration
Lantana camara
Acanthus pubescens
abstract
The extensive area of degraded tropical land and the calls to conserve forest biodiversity and sequester
carbon to offset climate change demonstrate the need to restore forest in the tropics. Deforested land
is sometimes replanted with fast-growing trees; however, the consequences of intensive replanting on
biomass accumulation or plant and animal diversity are poorly understood. The purpose of this study was
to determine how intensive replanting affected tropical forest regeneration and biomass accumulation
over ten years. We studied reforested sites in Kibale National Park, Uganda, that were degraded in the
1970s and replanted with five native tree species in 1995. We identified and measured the size of planted
versus naturally regenerating trees, and felled and weighed matched trees outside the park to calculate
region-specific allometric equations for above-ground tree biomass. The role of shrubs and grasses in
facilitating or hindering the establishment of trees was evaluated by correlating observed estimates of
percent cover to tree biomass. We found 39 tree species naturally regenerating in the restored area in
addition to the five originally planted species. Biomass was much higher for planted (15,675 kg/ha) than
naturally regenerated trees (4560 kg/ha), but naturally regenerating tree regrowth was an important ele-
ment of the landscape. The establishment of tree seedlings initially appeared to be facilitated by shrubs,
primarily Acanthus pubescens and the invasive Lantana camara; however, both are expected to hinder
tree recruitment in the long-term. Large and small-seeded tree species were found in the replanted
area, indicating that bird and mammal dispersers contributed to natural forest restoration. These results
demonstrate that intensive replanting can accelerate the natural accumulation of biomass and biodiver-
sity and facilitate the restoration of tropical forest communities. However, the long-term financial costs
and ecological benefits of planting and maintaining reforested areas need to be weighed against other
potential restoration strategies.
© 2010 Elsevier B.V. All rights reserved.
1. Introduction
Between 2000 and 2005, the world lost ∼7.3 million ha (or
∼200 km2) of forest per day (FAO, 2005). This figure does not
include vast areas of forest that are degraded by selective logging
or fire, both of which affect huge areas (Chapman et al., 2006).
This rapid rate of degradation is caused by a number of socio-
economic and political factors, partially due to large increases in
human population in most tropical countries (Brown and Pearce,
∗Corresponding author at: 855 Sherbrooke St. West, McGill University, Montreal,
Quebec, Canada H3A 2T7. Tel.: +1 514 398 1242.
E-mail address: Colin.Chapman@McGill.ca (C.A. Chapman).
1994; Myers, 2002; Wright and Muller-Landau, 2006; Jacob et al.,
2008). In many regions, widespread commercial logging of tropical
forests provides access to previously remote areas and promotes
policies for agricultural expansion. For example, in the 1970s the
governments of Brazil (Steininger, 2000), Indonesia (Lawrence,
2005), and Uganda (Hamilton, 1984; Struhsaker, 1997) adopted
agricultural reforms to expand productive areas. They provided tax
incentives to encourage migration from densely populated regions
to forested areas that could be converted to agriculture. Unfortu-
nately, many of these formerly forested regions rapidly lost soil
fertility; the subsequent decline in crop yield forced settlers to
abandon recently cleared land (Brown and Lugo, 1994; Dobson
et al., 1997). As a result, it is estimated that there are 350 millionha
deforested and another 500 million ha degraded secondary and
0378-1127/$ – see front matter © 2010 Elsevier B.V. All rights reserved.
doi:10.1016/j.foreco.2010.11.029
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704 P.A. Omeja et al. / Forest Ecology and Management 261 (2011) 703–709
primary tropical forests; many of these areas have since been aban-
doned (ITTO, 2002; Lamb et al., 2005). The current scale of tropical
deforestation and the large areas of degraded and abandoned lands
underscore the need for intervention to restore biodiversity and the
ecological functions, processes, and goods and services previously
provided by these lands (de Groot et al., 2002; Boyd and Banzhaf,
2007).
In many cases, natural succession can cause forest to return to
deforested areas within a reasonable time frame (i.e., years to a
few decades) (Reiners et al., 1994); however, succession can also
occur at a very slow rate (Brown and Lugo, 1994; Chapman and
Chapman, 1999) and be considered arrested if it does not proceed
within a reasonable time frame (Aide et al., 1996; Shono et al.,
2007). Instances of arrested succession from both anthropogenic
and natural causes have been reported throughout the tropics,
including Brazil (Nepstad et al., 1991), Colombia (Aide and Cavelier,
1994), Panama (Brokaw, 1983), Singapore (Corlett, 1991), Sri Lanka
(Ashton et al., 1997), and Uganda (Chapman and Chapman, 1999;
Chapman et al., 1999). Arrested succession may reflect a lack of tree
seeds or resprouts, high seed or seedling mortality (Nepstad et al.,
1996; Ashton et al., 1997), inhospitable abiotic and biotic site con-
ditions (e.g., a lack of mycorrhizae or limited soil nutrients; Janos,
1980; Uhl et al., 1988; Corlett, 1991; Lwanga, 2003; Lawes and
Chapman, 2005), or competitive dominance of herbs and shrubs
(Denslow et al., 1991; Duncan and Chapman, 1999; George and
Bazzaz, 1999a,b; Duncan and Chapman, 2003b).
Seed limitation often inhibits the recovery of tropical forest tree
biodiversity when deforested areas are large or far from intact (i.e.,
non-degraded) forest. The majority of tropical tree species have
animal-dispersed fruits (Howe and Smallwood, 1982; Chapman,
1995), but many frugivores avoid deforested areas (DaSilva et al.,
1996; Zanne and Chapman, 2001), especially mammals (Chapman
and Chapman, 1999). Wind-dispersed seeds may arrive at defor-
ested sites in high numbers; however, since they are small-seeded,
the harsh micro-site conditions associated with abandoned lands
often limit the establishment of these seedlings. For example, Ingle
(2003) found that the stem density of vertebrate-dispersed species
outnumbered wind-dispersed species in montane forests of the
Philippines, although 15 times more wind-dispersed seeds arrived
in these degraded areas. Furthermore, seed dispersal limitation can
be severe for large-seeded tree species because in many systems the
predominant seed dispersal agents in abandoned areas are small
birds and bats that typically only carry small seeds (Nepstad et
al., 1996; Duncan and Chapman, 1999). These limitations can be
reduced when remnant trees (Guevara et al., 1986) and shrubs
(Vieira et al., 1994; Holl, 2002) are present because they attract
large seed dispersers and facilitate forest regeneration under their
canopies. In some deforested areas grasses invade before primary
successional tree species establish. High grass biomass can increase
the likelihood and intensity of fire, further arresting natural for-
est restoration (Nepstad et al., 1990; Lwanga, 2003). In previously
forested systems, fire can also impoverish soils, reduce seedling
growth (Aide et al., 1996), and impede forest recovery (Buschbacher
et al., 1988).
The Uganda Wildlife Authority (UWA) and Face the Future (for-
merly the Forests Absorbing Carbon Emissions (FACE) Foundation),
hereafter UWA–FACE, have a collaborative reforestation program
in Kibale. This intensive program was designed to facilitate the
rapid restoration of woody vegetation in a large area of the park
that was illegally encroached by agriculturalists who cleared forest
and grassland areas to plant crops (Chapman and Lambert, 2000;
Struhsaker, 2003). The primary objectives of our study were to
(1) quantify species richness of naturally regenerating (i.e., non-
planted) trees in Kibale National Park, Uganda, and (2) compare
the biomass accumulation of planted versus naturally regenerating
trees in sites replanted by UWA–FACE.
Fig. 1. The location of Kibale National Park within Uganda and the location of the all
restoration compartments planted to date (light grey) and the study compartments
(dark grey) within Kibale.
2. Methods
2.1. Study area and UWA–FACE field activities and costs
This study was conducted between May 2005 and May 2006
in the southern section of Kibale National Park, Uganda (Fig. 1).
The park (795 km2) is located in western Uganda (0.13–0.41N
and 30.19–30.32E) near the foothills of the Ruwenzori Moun-
tains (Chapman and Lambert, 2000; Struhsaker, 1997). Kibale is a
mid-altitude, moist-evergreen forest that receives 1697 mm of rain
annually (1990–2009). Although there are some early successional
species in the park (e.g., Albizia grandibracteata,Polysciasfulva,
Trema orientalis;Zanne and Chapman, 2005), Kibale is notable for
its lack of aggressive colonizers typical of other tropical regions
(e.g., Musanga spp., Cecropia spp.). Within the park, there is a
gradual decrease in elevation from north to south, which corre-
sponds to an increase in temperature, decrease in rainfall, and
changes in forest composition (Struhsaker, 1997). In the south,
the un-encroached forest is dominated by Cynometra alexandri
and its affiliated species; however, there are also areas of mixed
forest along riverine strips and even Acacia woodland to the far
south.
In 1994, UWA–FACE started a reforestation program to estab-
lish carbon offsets on degraded land in Kibale, with tree planting
commencing in 1995. The project is based in the southern part of
Kibale in an area which was illegally occupied by agriculturalists in
the 1970s until their eviction in 1992 (van Orsdol, 1986; Baranga,
1991). Agricultural encroachment adversely affected approxi-
mately 120 km2(15% of the total area of Kibale), mainly leading to
forest destruction and grassland clearing (Chapman and Lambert,
2000). After the eviction, grassland areas became dominated by ele-
phant grass (Pennisetum purpureum) because frequent fires set by
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poachers or that spread into the park from neighboring subsistence
farms prevented natural forest regeneration (Struhsaker, 2003).
For the first phase of the replanting program (Phase 1:
1994–1997), the majority of seedlings were raised from seeds or
collected from the wild (wildings). A limited number of species
survived transplantation into the restoration areas, which were
typically dominated by the grass P. purpureum. As a result, future
replanting phases concentrated on planting Albizia spp., Bridelia
micrantha,Sapium ellipticum,Celtis durandii, and Warbugia ugan-
densis seedlings. The species were collected from the forest floor
in areas that were not affected by agricultural encroachment (one
person can collect up to 180 wildings per hour), transferred to a
nursery, watered, and kept in plastic covered chambers with high
humidity for four weeks until a new root system formed. Special
care was taken during collection and acclimatization in the nurs-
ery to ensure the roots healed before wildings were planted in the
restoration areas. Vegetative propagation by cuttings accounted
for 5% of the trees planted. Cuttings were directly set to root in
7cm×21 cm poly-bags containing a 1:1 mixture of forest topsoil
and sand.
Preparing the site for planting involved clearing the elephant
grass along a series of 2-m wide paths spaced in a 5 m ×5 m grid,
and digging a small pit every 5 m along the paths for the seedling;
thus, 400 tree seedlings were planted per hectare. The replanted
area was divided into a series of well mapped compartments to
facilitate monitoring. The size and placement of the compartments
were done by UWA–FACE and were based on convenience (e.g.,
the edge of the road formed the compartment border). A total of
16.35 km2were replanted in eight compartments during Phase 1
(Fig. 1). After planting, the seedlings were monitored and weeded
two to three times a year to protect them from fire and to limit
competition, primarily from grasses. Fire breaks were cut between
compartments and UWA–FACE staff fought fires and maintained
access roads to protect the planted areas. The total cost for these
activities was approximately US$120,000 per km2(UWA–FACE,
2005).
2.2. Sampling design
The eight study compartments were mapped using GPS units at
the time of planting and the corners marked with trenches. We
used maps from UWA–FACE to identify the compartments and
individual trees that were planted during Phase I. We randomly
established ten 10 m ×50m plots in each compartment using ran-
domly generated locations and directions. If a plot extended beyond
the edge of the compartment, it was re-oriented to 180◦oppo-
site the original direction. We identified each seedling, sapling,
and mature tree in the plot and measured their height, diame-
ter at breast height (DBH), and diameter at ground height (DGH).
Diameter at breast height is a commonly used metric that allows
comparisons with other studies; diameter at ground height allowed
us to evaluate small stems and was used in our allometric analysis
to estimate biomass. It was easy to separate planted from naturally
regenerating individuals as the seedlings were originally planted
in straight lines at set intervals.
We established small (1 m ×1 m) subplots every 5 m next to
the path used for planting. We estimated the percentage cover of
shrubs, dominant grasses, and average vegetation height (cm) and
identified plants using well recognized plant keys (Polhill, 1952;
Kingston, 1967; Hamilton, 1991; Katende et al., 1995; Lwanga,
1996).
Species richness and density data were log transformed
prior to analysis. We estimated species diversity using the
Shannon–Wiener index (Krebs, 1989) and evaluated the effects
of grass and shrub cover on species richness, biomass, and tree
height using a regression analysis. We selected and measured
trees in forested land adjacent to the park to provide a site-
specific estimate of dry tree biomass. We chose nine species
commonly found in regenerating areas: Albizia grandibracteata,
Bridelia micrantha,Celtis africana,Celtis durandii,Clausena spp.,
Maesa lanceolata,Funtumia latifolia,Milletia dura, and Trema ori-
entalis. Individuals of similar sizes to regenerating trees (both
planted and naturally regenerating stems) in the replanted area
were selected (i.e., DBH = 1.1–10.0 cm and DGH = 1.6–11.0 cm; dry
weight = 0.25–10 kg; total n = 200 stems), and we measured their
DBH and DGH. We felled the trees at ground level, removed the
branches, and measured the total dry weight of stems and leaves.
We cut the tree stems into small sections and air dried them until a
constant mass was attained. The dry biomass of trees in the planted
areas was predicted by the allometric equation log DGH (r2= 0.653,
n= 200, y= 2.053x+ 2.056).
We measured the seed size of tree species found naturally regen-
erating in the plots to evaluate the relative role of birds and large
mammals in moving seeds into the restoration areas. We mea-
sured the longest axis of 20 seeds from six adult trees in the
adjacent undisturbed forest. Based on the observations of the for-
aging activity of birds and mammals (Chapman, unpublished data),
we estimated the largest seed typically dispersed by birds in this
community to be Olea capensis (mean length = 1.39 cm) (exclud-
ing black and white casqued hornbills, Bycanister subcylindrucus).
Seeds larger than 1.39 cm were assumed to be mammal dispersed.
Smaller seeds could be dispersed by either mammals or birds,
since mammals often disperse both small-and large-seeded species
(Wrangham et al., 1994).
3. Results
The mean predicted biomass of planted trees in the eight
compartments was 15,657 kg/ha (range: 4120 kg/ha (n= 76) to
26,916 kg/ha (n= 244); Table 1), while the mean biomass of the
naturally regenerating trees was 4560 kg/ha (range: 1126 kg/ha
(n= 30) to 11,470kg/ha (n= 162); Table 1). In general, the naturally
regenerating trees represented 22.5% of the total biomass in the
restoration area.
A total of 39 tree species established themselves naturally in
the compartments in the ten years since replanting (Table 2). The
restoration program introduced only one tree species that did not
regenerate naturally in the area (Sapium ellipticum), although this
species is commonly found in the neighboring forest. Determin-
ing what proportion of the species in the adjacent forest these
39 species represent is difficult: many species in Kibale are rare
and would require extensive sampling to identify, and the forest
composition changes along a north–south gradient in elevation
and rainfall. However, species/area curves from our previous sam-
pling suggest that the number of new species found decreases after
sampling approximately 2 ha of scattered plots. Sampling 8.6 ha
throughout the park resulted in the identification of 74 species
(Chapman et al., 1997). This suggests that approximately half the
species in the adjacent forest were also present in restoration
area.
Tree species diversity and evenness varied among plots: species
diversity was highest in compartments 109 (2.36) and 114 (2.32),
while evenness was highest in compartment 109 (0.45), followed
equally by 101, 103 and 113 (all 0.38; Table 3). Compartment 109
had 23 tree species and the individuals were distributed relatively
equally among these species. The contribution of the tree planting
to species diversity was negligible (Table 3).
Acanthus pubescens and Lantana camara were the dominant
shrubs in all compartments. We did not find a relationship
between the percentage cover of shrubs and grasses in the
subplots and the biomass of naturally established or planted
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Table 1
Above ground biomass (kg/ha) and stem densities (per ha) of woody tree species, and percentage coverage of grass and shrubs in the study compartments, Kibale National
Park, Uganda.
Compartment Planted Natural regrowth
Grass Shrub Biomass NNo. of spp. Biomass NNo of spp.
101 27 27 9620 152 5 1154 30 7
102 26 36 20,786 234 5 1126 42 8
103 31 37 15,114 318 5 7318 148 14
107 40 27 4,120 152 5 2496 72 8
108 30 34 14,126 430 5 2626 66 8
109 30 37 12,440 248 5 5060 246 19
113 24 37 22,276 448 5 11,470 162 18
114 37 28 26,916 488 5 5230 134 21
Total 125,398 36,480
Mean 15,675 4560
Standard deviation 7362 3534
Table 2
Tree species, mean seed length (where available) and frequencies for the study compartments Kibale National Park, Uganda. Tree species are arranged according to their
frequency of occurrence. Trees marked with an asterisk are considered to be mammal dispersed.
Tree species Mean seed length (cm) Not planted Planted Total
Bridelia micrantha 0.74 2 173 175
Warbugia ugandensis 0.99 1 57 58
Sapium ellipticum 0.57 0 55 55
Albizia grandibracteata 0.77 4 16 20
Harrisonia abyssinica 0.39 10 0 10
Allophyllus rubifolius 0.64 9 0 9
Combretum molle 808
Acacia spp. 0.68 7 0 7
*Mimusops bagshawei 1.65 8 0 8
Tabernaemontana holstii 1.06 7 0 7
Grewia occidentalis 0.75 7 0 7
*Erythrina abyssinica 1.83 7 0 7
Cassia spectabilis 0.83 6 0 6
Ficus capensis 0.10 6 0 6
Gardenia lanciloba 505
Rauvolfia vomitoria 1.27 4 0 4
Prunus africana 0.87 2 0 2
Markhamia lutea 1.17 4 0 4
Croton macrostachyus 0.82 0 4 4
Diospyros abyssinica 0.75 3 0 3
Antidesma spp. 303
Spathodea campanulata 1.00 3 0 3
Funtumia latifolia 0.29 3 0 3
Celtis durandii 0.50 0 2 2
*Chrysophyllum albidum 2.10 2 0 2
Olea capensis 1.38 1 0 1
Uvariopsis congensis 1.24 1 0 1
Mangifera indica 101
Carissa edulis 101
*Pseudospondias microcarpa 1.47 1 0 1
Maesa lanceolata 0.32 1 0 1
Kigelia africana 1.16 1 0 1
Persea americana 001
Apodytesdimidiate 0.47 1 0 1
Coffea eugenioides 0.93 1 0 1
Psidium guajava 0.49 1 0 1
Eudenia spp. 0.97 1 0 1
Dovyalis microcarpa 0.90 1 0 1
Dasylepis eggeling 0.73 1 0 1
Blighia unijugata 1.19 1 0 1
119 305 425
Table 3
Species richness, Shannon diversity index, and evenness of woody tree species naturally growing in the eight study compartments, Kibale National Park, Uganda.
Compartment All species Natural regrowth
NDiversity Index Evenness Diversity Index Evenness N
101 180 1.71 0.38 1.71 0.63 15
102 276 1.24 0.25 1.04 0.64 21
103 233 2.09 0.38 2.13 0.49 74
107 224 1.71 0.36 1.71 0.48 36
108 248 1.38 0.25 1.71 0.49 33
109 247 2.49 0.45 2.36 0.49 23
113 305 2.16 0.38 2.22 0.55 81
114 311 2.00 0.35 0.25 0.20 67
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P.A. Omeja et al. / Forest Ecology and Management 261 (2011) 703–709 707
trees. However, after combining the planted and the naturally
established trees, the percentage of shrub cover was positively
correlated to tree biomass (r2= 0.207, p= 0.026), while there was
a negative correlation between tree biomass and percentage
grass (r2=−0.081, p= 0.018). However, the amount of variation
in tree biomass explained by either of these relationships was
small.
The largest seed that birds in Kibale have been observed
to disperse is Olea capensis, which has an average seed length
of 1.39 cm (Chapman, unpublished data; Table 2). Mimusops
bagshawei,Chrysophyllum albidum,Pseudospondias microcarpa, and
Erythrina abyssinica were found in the compartments and their seed
sizes are longer than 1.39 cm; therefore it is likely that these species
were dispersed into the area by mammals (E. abyssinica is thought
to be a mimetic seed, one that is brightly colored but without any
pulp, so it does not provide any nutritional reward for dispersal).
Kigelia africana seedlings (average seed length =1.16cm) were also
found in the restoration areas; these seeds were likely dispersed
by elephants or large primates (probably baboons or chimpanzees)
since the seeds are embedded in a large fruit with a hard rind (aver-
age 60 cm long) (Katende et al., 1995). Seeds smaller than 1.39 cm
probably do not place a size constraint on the disperser and can be
dispersed by both mammals and birds.
4. Discussion
The results of this research have five contributions, they: (1)
demonstrate that the replanting program was successful at accu-
mulating biomass of planted trees, (2) show that native trees
naturally establish under planted trees, quickly creating a reason-
ably rich tree species assemblage, (3) quantify the financial costs
of enrichment planting, which is considered relative to poten-
tial costs of other strategies, such as fire exclusion, (4) consider
the role of grasses and shrubs, particularly the invasive Lan-
tana camera, in forest regeneration, and (5) provide evidence that
large and small-seeded tree species were found in the replanted
area, indicating that bird and mammal dispersers were likely
using regenerating forest habitat and contributing to natural forest
restoration.
First, we demonstrate that the active restoration of the degraded
lands in Kibale was relatively successful as indicated by the pre-
dicted biomass the planted trees accumulated approximately 10
years after planting. Success should be measured (i) relative to
the goals of the program, which in this case was to accumu-
late biomass, and (ii) relative to other methods, of which there
have been many that have been evaluated in Kibale or elsewhere
(Chapman and Chapman, 1999; Chapman et al., 2002; Martinez-
Garza and Howe, 2007; Vieira et al., 2009). If agricultural land
in the Kibale region is abandoned and left to recover without
intervention, it will take a very long time for a closed canopy tree-
dominated community to regenerate, even if it is close to seed
sources (Chapman and Chapman, 1999). Chapman and Chapman
(1999) studied such an area and found that it took four years for
trees >5 m tall to reach a biomass of 8.9 kg/ha; even 17 years later
these areas are still dominated by shrubs and grasses (unpublished
data). However, excluding fire from degraded, regenerating agri-
cultural land increases the rate of reforestation (Lwanga, 2003) and
fire can be excluded over large areas at relatively low cost (Omeja,
unpublished data). Some researchers have advocated using timber
plantations, typically with non-native species, as a means to eco-
nomically reforest areas (Lugo, 1997; Parrotta et al., 1997; Lamb,
1998), since native forests can regenerate underneath the planta-
tions or following timber harvest, and timber sales can help pay
for restoration efforts. In Kibale, this management strategy was
successful for reforesting areas, despite the considerable damage
caused by initial and poorly planned timber extraction (Chapman
and Chapman, 1996; Duncan and Chapman, 2003a; Omeja et al.,
2009). This use of this plantation strategy of reforestation in
Kibale included all functional groups (pioneers, later-seral species
and wind and animal dispersed); however, animal dispersed tree
species will rarely be found if the plantation is not near native forest
(Zanne and Chapman, 2001).
Second, we documented that many tree species will natu-
rally establish under planted trees, creating a reasonably rich tree
species assemblage within a period of time that is reasonable for
management. The natural establishment of a diverse tree commu-
nity with accumulating biomass (almost 1/4 of the total biomass)
suggests that these initially species-poor areas (i.e., five planted tree
species) will quickly and naturally become richer and accumulate
woody tree biomass. One of the most important contributions of
programs such as the UWA–FACE restoration project is the rapid
rate of biomass accumulation of woody trees that sequester carbon
(Holl and Howarth, 2000; UWA–FACE, 2005; Coomes et al., 2008).
The fact that the diverse community of naturally regenerating
trees within these restoration areas has also quickly accumulated
biomass suggests that such programs may contribute to the rela-
tively rapid restoration of biodiversity. Furthermore, observations
of animal occurrence in the replanted areas suggest that biodiver-
sity of both plants and animals is rapidly recovering (Jacob and
Chapman, unpublished data).
Third, this research raises the question of when the ecological
gains of enrichment planting are worth the considerable financial
costs. Replanting of a mixture of early and late successional tree
species has been recommended to restore diversity more rapidly
than removing disturbance and waiting for natural regeneration
to occur (Yirdaw, 2001). However, the potential application of
enrichment planting to facilitate biodiversity restoration is based
on the premise that heightened seedling recruitment will lead to
greatly enhanced regeneration of mature trees, which are worth the
additional investment that replanting programs require (Plumptre,
1995; Chapman and Chapman, 1996). Even though planted trees
have a high probability of establishing and growing, especially
when planting is timed to optimize survival rates (Duncan and
Chapman, 2003a), enrichment planting generally involves costly
nursery maintenance and field labor. For example, while enrich-
ment planting depends on the density of the trees planted, forest
restoration is estimated to cost $250,000 US per km2on bauxite-
mined land in the Amazon (Parrotta and Knowles, 1999). It cost
$120,000 US per km2to conduct the enrichment planting and
establish and maintain the fire breaks that we evaluated in Kibale
(UWA–FACE, 2005). Given the relative success of the restoration
management in Kibale, we suggest that fire exclusion be evalu-
ated in more depth as a restoration strategy, given its relatively low
cost and ease of protecting an area (see also Lwanga, 2003). To the
credit of the UWA–FACE program, they established and maintain
an extensive network of fire breaks to protect the planted areas.
Anecdotal evidence suggests that this action has helped to protect
as yet unplanted areas that are passively regenerating. These areas,
plus others studied by Lwanga (2003), have been protected from
fire for <1 to ∼30 years; together they create a series of natural
experiments to evaluate the role of fire prevention in tropical forest
reforestation.
Fourth, the positive relationship we found between the com-
bined biomass of planted naturally establishing tree species and
the percentage of plots covered by shrubs concurs with results
from other tropical forests (Vieira et al., 1994; Aide et al., 1995;
Holl, 2002). This suggests that shrubs, in this case primarily L.
camara which forms 95% of the shrub layer, initially facilitates
woody tree species establishment. Shrubs have been found to lower
temperatures, increase soil moisture (Bertness and Callaway, 1994;
Callaway and Walker, 1997), and buffer seedlings from harsh envi-
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708 P.A. Omeja et al. / Forest Ecology and Management 261 (2011) 703–709
ronmental condition, such as heat from the sun (DaSilva et al.,
1996; Nepstad et al., 1996). These shrubs may also be beneficial in
that they deter fire (Lwanga, 2003) and reduce browsing (Sharam
et al., 2009). Generally, tropical areas with well-established ever-
green undergrowth tend to be less susceptible to fire damage than
areas with an accumulation of dry grasses (Parrotta, 1992; Lugo
et al., 1993; Parrotta, 1993). Exposed grassy sites, on the other hand,
have negative effects on site microclimate, with more pronounced
temperature fluctuations and differences in humidity and water
availability (Bazzaz, 1991) that are extremely stressful for plants
(Loik and Holl, 1999). Furthermore, grass-dominated areas are very
prone to fire in the dry season and successive fires severely curtail
forest succession.
Presently, observations and the correlation that we found sug-
gest that the shrub layer in Kibale appears to be initially helping
the regenerating trees acclimatize and form the dominant canopy;
however, studies have shown that there may be a shift in the
direction of this interaction over the long-term (Callaway, 1997).
In the future, the shrub layer, particularly that of the dense L.
camara shrubs, may not be conducive to a diverse regenerating
plant community as it may suppress seedlings from reaching forest
canopy layer (Zalucki et al., 2007). Introduced L. camara has been
reported to have harmful effects on ecosystems in other regions.
For example, Lantana sp. was ranked as the most significant weed of
non-agricultural areas in south-east Queensland, Australia (Zalucki
et al., 2007). In Kenya, the replacement of native pastures by L.
camara is threatening the habitat of the sable antelope (Hippo-
tragus niger), and acts as a safe haven for disease transmitting
tsetse-flies and a detrimental force to forest and ecosystem regen-
eration (Zalucki et al., 2007). The impenetrable thickets formed by L.
camara are also concerning because they restrict the movement of
even highly mobile animals, such as elephants (Loxodonta africana)
and baboons (Papio anubis), which likely reduces seed dispersal.
The balance between the advantages that establishing seedlings
receive from altered microclimate around L. camara versus the dis-
advantages of reduced seed input and potential long-term inhibited
growth remains unknown. Thus, if L. camara continues to form a
dense shrub layer then the regeneration pathway to forest recov-
ery may be diverted to an arrested successional state. This suggests
that there is a pressing need to investigate the long-term role
of L. camara in forest restoration and experimentation involving
removal would help clarify its role.
Lastly, the size of seeds of tree species naturally regenerating in
the area reflects the presence of different dispersing animals in the
restoration area. Most fruit-eating animals only feed on a portion
of the fruits that are available in the forest. Fruit selection no doubt
depends on a variety of factors such as morphology, nutritional
value, color, and the abundance of secondary compounds. How-
ever, seed size is a key factor known to influence which frugivores
disperse different species of fruits (Herrera, 1985; Wheelwright,
1985; Fischer and Chapman, 1993). A seed cannot be moved away
from the adult tree crown if it exceeds the size of gape of a bird
(Wheelwright, 1985; Chapman et al., 2003). Given that a num-
ber of tree species have seeds too large to be dispersed by birds,
and the fact that mammals disperse both large and small seeds,
it seems probable that mammals, such as elephants and baboons,
have played an important role in the accumulation of tree species
richness in the regenerating areas of Kibale. Bridelia micrantha (bird
dispersed) and A. grandibracteata (wind and rodent dispersed) had
higher stem densities than other species, and most of these were
growing naturally. These are trees species with small seeds, long
distance dispersal capability, long persistence in the soil seed bank,
and have the ability to colonize gaps within forest canopies (Zanne
and Chapman, 2005; Omeja et al., 2009), attributes that allow rapid
colonization of disturbed areas.
5. Conclusion
The potential for restoring degraded forests in Kibale National
Park is high following intensive planting of only five tree species.
We found that 39 tree non-planted species were able to establish
and accumulate considerable biomass following the planting pro-
gram. This regeneration facilitates the recovery of other tree and
animal species of conservation concern. There are many limitations
to forest restoration, and we speculated that the major future lim-
itation at this site is the heavy presence of L. camara that may slow
the progress of current restoration efforts. A more detailed under-
standing on the role of L. camara in future restoration needs to be
conducted.
Acknowledgments
This research was supported by the Canada Research Chairs
Program, Wildlife Conservation Society, Natural Science and Engi-
neering Research Council of Canada, Committee on Scientific
and Technological Cooperation of the Organization of Islamic
Conference, Islamabad, Pakistan, and International Foundation
for Science, Stockholm, Sweden. We thank the Uganda Wildlife
Authority, Uganda National Council for Science and Technology,
and Makerere University Biological Field Station for granting per-
mission to conduct this research. The field assistants of the Kibale
Fish and Monkey Projects provided very valuable help. Previous
versions of this manuscript benefited from comments by Lauren
Chapman and Mike Lawes.
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