in bird communities following conversion of lowland forest to oil
palm and rubber plantations in southern Thailand
By Sirirak Aratrakorn, Somying Thunhikorn and Paul F. Donald
This article was originally published in Bird Conservation International
16 (2006) the journal of Birdlife
International and was kindly submitted by Paul
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This paper describes changes in bird communities following the conversion
of lowland forest to commercial oil palm and rubber plantations. Conversion
of forest to plantations resulted in a reduction in species richness
of at least 60%, with insectivores and frugivores suffering greater
losses than more omnivorous species. Of the 128 species recorded across
all habitats, 84% were recorded in forest, and 60% were recorded only
in that habitat. Of the 16 Globally Threatened or Near-Threatened
species recorded in the study, 15 were recorded only in forest. Species
occurring in plantations were significantly more widespread in Thailand
than species recorded only in forests and had a tendency towards smaller
body size. Species richness in plantations was unaffected by plantation
age or distance from nearest forest edge, but was significantly greater
where undergrowth was allowed to regenerate beneath the crop trees.
Bird communities in oil palm and rubber plantations were extremely
similar, and there was a strong positive correlation across species
in their relative abundance in each plantation type. The results indicate
that a high proportion of species formerly present in the region are
unable to adapt to conversion of forest to oil palm and rubber plantations,
resulting in large losses of bird species and family richness and
the replacement of species with restricted ranges and high conservation
status by those with extensive ranges and low conservation status.
Initiatives that reduce pressure to clear new land for plantations,
for example by increasing productivity in existing plantations and
improving protected area networks, are likely to be more effective
in conserving threatened forest birds than initiatives to improve
conditions within plantations, though both should be encouraged.
The loss of tropical forests, and particularly of lowland forests,
represents one of the greatest threats to bird diversity globally
(e.g. BirdLife International 2004, Niesten et al. 2004).
The greatest single cause of deforestation is the clearance of land
for agriculture, which is proceeding most rapidly in countries with
the highest biodiversity (Balmford and Long 1994) and at a higher
rate in areas holding restricted range bird species than outside
them (Scharlemann et al. 2004). Losses of tropical forest to agriculture
are likely to continue. In the developing world, 109 hectares of
pristine habitats, an area approximately equal to that of all the
planet’s remaining tropical rainforests (Mayaux et al.
1998), may be cleared for agriculture by 2050 (Tilman et al.
2001). Two of the most rapidly expanding crops in tropical regions
are oil palm Elaeis guineensis and rubber Hevea brasiliensis
(Clay 2004). The global area of oil palm production nearly tripled
between 1961 and 2000, most of this increaseat the expense of natural
habitats,1 and per capita consumption of vegetable oils increased
more rapidly than any other food during the same period (Clay 2004).
The highest rate of oil palm expansion has been in South-East Asia,
particularly Malaysia, Indonesia and Thailand, where the area more
than doubled in the 10 years from 1995 to 2004. Oil palm now covers
at least 10 million hectares globally and is the world’s second
largest source of edible oils, after soybean Glycine max.
In Malaysia, oil palm is the second largest export earner and makes
up 56% of the country’s tree cover. Rubber has also increased
greatly at the expense of natural habitats, again mainly lowland
tropical forest, approximately doubling in area between 1960 and
2000. Once again, South-East Asia dominates production, with Indonesia,
Malaysia and Thailand being the world’s largest producers.
In Thailand, the area of rubber quadrupled between 1960 and 2000.
Current high prices for both oil palm and rubber, combined with
some aggressive and often subsidised national strategies to increase
production (Wu et al. 2001, Clay 2004), mean that expansion
of both crops is likely to increase in the near future. This poses
a severe threat to the few remaining lowland forests in one of the
world’s most biodiverse regions. The areas most suitable for
oil palm and rubber production lie in the tropical lowlands 10°
either side of the equator. Lowland South-East Asia, where the spread
of these crops has been greatest, is a region of particularly high
biodiversity and one supporting some of the world’s most threatened
forests (Lambert and Collar 2002). The Sundaic lowland forests,
confined to the Thai- Malay Peninsula and the Greater Sunda islands,
have suffered catastrophic losses of area in all range states (except
Myanmar), largely because of the spread of oil palm and rubber.
Across a range
of agricultural systems, there is a general pattern of biodiversity
loss when natural habitats are converted to agricultural ones, and
a further loss of biodiversity as such systems are intensified (Donald
2004). However, despite the widely publicised threat posed to biodiversity
by oil palm and rubber plantations, remarkably few empirical studies
have documented biodiversity change following forest conversion
to commercial plantations of these crops. The few published comparisons
of biodiversity in tropical plantations and pristine habitats, summarized
in Donald (2004), suggest that rubber and oil palm plantations are
particularly poor habitats for wildlife and offer little environmental
compensation for forest loss.
The most detailed
study to date was that of Danielsen and Heegaard (1995), who documented
the almost complete loss of most vertebrate taxa as forest was converted
to oil palm. Similarly, Chung et al. (2000) found that
beetle populations were far lower in oil palm than in a range of
other natural and agricultural habitats. While most taxa appear
to be adversely affected, some are capable of using rubber and oil
palm plantations. Wild pigs Sus scrofa reach very high
densities in lowland forest surrounding oil palm plantations in
Malaysia, at least partly because they feed on the abundant fallen
oil palm fruit (Ickes 2001), and White-faced monkeys Cebus capucinus
inhabiting agricultural landscapes in Central America feed largely
on oil palm fruit (Williams and Vaughan 2001). However, the few
empirical studies available suggest that any such gains are greatly
outweighed by losses, particularly where crops are grown in intensive
monocultures. Empirical evaluation of the effects of conversion
of forest to plantations is clearly important if the true environmental
effects of plantations are to be quantified and accounted for in
global biodiversity inventories and strategies.
This paper describes
the results of research undertaken in southern Thailand to estimate
and quantify avifaunal change following the replacement of forest
by oil palm and rubber plantations, and to assess the effects of
vegetation structure and distance from forest patches on bird communities
within plantations. This was undertaken as part of a research project
designed to support efforts to save the small Thai population of
the Critically Endangered Gurney’s Pitta Pitta gurneyi,
a Sundaic lowland forest endemic species threatened by loss of forests
to oil palm and rubber plantations (BirdLife International 2001).
Study site and data collection Fieldwork was undertaken in and around
Khao Pra-Bang Khram Wildlife Sanctuary and Bang Khram National Reserve
Forest in Krabi Province, peninsular Thailand. The Wildlife Sanctuary
and the Forest Reserve contain remnants of the lowland forests that
once covered much of peninsular Thailand but have now largely disappeared.
Most of the forest loss in the areas around Khao Pra-Bang Khram has
been caused by legal and illegal conversion of forest to commercial
plantations of rubber and oil palm, largely within the last 20 years.
Most recent clearances have been for oil palm planting, although some
new rubber plantations are also being established. All counts were
undertaken within a 10 km radius of the Khao Pra-Bang Khram Wildlife
Sanctuary headquarters (Figure 1).
1 : Distribution of census points. Exact location of forest
census sites could not be plotted because GPS signals in forest were
too weak to allow locations to be recorded.
Species Counts (TSCs; e.g. Pomeroy and Dranzoa 1997, Bibby et al.
2000, Freeman et al. 2003) were used to collect data on bird
species richness and relative abundance. This method was chosen in
preference to distance sampling methods because of the difficulty
in judging distances in forest, the potentially confounding effects
on density estimates of mobile mixed-species flocks, and because estimates
of species richness and composition and relative abundance were considered
sufficient to assess the major gradients in bird communities in each
habitat type. The TSCs were converted to measures of relative abundance,
l, using the method of Freeman et al. (2003). A total of
30 oil palm plantations and 30 rubber plantations were selected using
aerial photographs of the area. This sample size was determined by
undertaking a small number (n = 5) of preliminary counts in each habitat
and using the formula:
(where M+ is the number of additional sample units needed, Q is the
required Percentage Relative Precision, s is the standard deviation
from a preliminary sample, N1 is the mean number of species recorded
in preliminary sample and M1 is the number of sample units in a preliminary
sample) to estimate the sample size required to achieve a Percentage
Relative Precision of species richness of 10% (i.e. the 95% CLs fall
within 10% of the mean). These initial counts suggested that a sample
size of around 30 would be sufficient to meet this predetermined level
of precision, though the non-normality of the response variable might
influence the level of precision actually achieved. Two counts, separated
by at least 4 weeks, were made between 3 February and 27 March 2004
from points located in the centre of each plantation, and at 30 widely
scattered points within the adjacent forest. Counts were undertaken
between 06h00 and 10h00. Points were sufficiently far apart that individual
birds could not be recorded from more than one location, although
exact distances could not be measured as portable GPS units worked
infrequently in forest. Counts lasted 20 minutes, and were broken
into five 4-minute blocks. The first time each species was detected,
it was entered on field recording forms in the appropriate 4-minute
block. As well as recording the plantation type, a score of undergrowth
density was collected, recorded as a binary variable 0 (little or
absent) or 1 (dense and extensive). Due to the lack of adequate maps
and poor functioning of portable GPS units, distance of each plantation
from the nearest forest edge was also recorded as a binary variable,
indicating whether the plantation was within or beyond 1 km from the
Generalized linear models were used to derive estimates of species
richness and their standard errors for each habitat type and to
test the null hypothesis that species richness was equal in all
habitat types. In the case of the plantations, similar models were
used to assess the effects on species richness of undergrowth structure,
plantation age and type and distance from forest edge. Models assumed
a Poisson distribution of errors and a log-link function was specified.
Minimum adequate models were derived by sequential backwards deletion
from the maximal model. Results are graphically represented as non-parametric
boxplots, as these visually convey more information about the distribution
of Poisson data than do plots of means and their associated error
bars and avoid problems such as error bars spanning proportions
of 0 or 1 (Crawley 2002).
We used the
general and well-documented relationship between range size and
population size (e.g. Gaston et al. 2000, He et al.
2002, Selmi and Boulinier 2004), and between range size and extinction
likelihood (e.g. BirdLife International 2004), to use an estimate
of the range of each species in Thailand as a surrogate estimate
of its relative population size and conservation importance. The
assumption was made that species with very extensive ranges in Thailand
are more likely to be of lower conservation concern than species
with more restricted ranges. Further assumptions about habitat specificity
might also be inferred from estimates of range size (e.g. Gregory
and Gaston 2000). No systematic data on range size, such as ornithological
atlas data, or on population size exist for birds in Thailand, so
we estimated relative range size from the maps in Robson (2002),
using a transparent graticule. Range was expressed as the proportion
of 20 equal-sized cells that overlapped at least partly with the
depicted range. Only breeding ranges were included in these analyses,
thereby excluding a number of species that only winter in Thailand.
Body size (length) was extracted from the same source. We then tested
the null hypothesis that species inhabiting plantations and forests
have equal average range sizes in Thailand and body sizes using
non-parametric tests. Each species was classified into one of a
number of foraging guilds based on information in Fogden (1971),
Wong (1986) and Wells (1999), and followed the 10 classes identified
by Wong (1986). For species included in none of these sources, categorization
was based upon the authors’ field experience.
Species accumulation curves (Figure 2) suggest that 20 minutes were
sufficient to record most species in plantations but that species
richness in forest might be underestimated, although having two count
periods is likely to have reduced this bias in the final cumulative
estimates of species richness.
Accumulation rates of species detection across 4-minute count periods
in single 20-minute counts. Unbroken line, forest; dashed line, rubber,
dotted line, oil palm.
1 : Species richness and, in parentheses, overall relative
abundance, l, of birds in different feeding guilds (following Wong
Figure 3 : Boxplot of cumulative species richness
recorded over two 20-minute observation periods in forest, oil palm
and rubber. The horizontal line represents the median, the box the
interquartile range and the vertical lines span the range of the values
lying between the interquartile and 1.5 times the interquartile range.
Outliers beyond this are represented by asterisks. The mean is represented
by a dot. GLMs indicated that the means differed significantly (F2,87
= 114.5, P < 0.0001).
all habitat types, 128 species were recorded (Table 1). The cumulative
number of species recorded across the two counts at each point ranged
from 4 to 35 (mean = 13.5, SE = 0.82). A total of 41 species was
recorded in oil palm plantations, 41 in rubber plantations and 108
in forest. Four species (3% of all species recorded) were recorded
only in oil palm plantations, 6 species (5%) only in rubber and
77 species (60%) only in forest. Mean species richness per point
was significantly higher in forest (22.6, SE = 1.04) than in rubber
(9.1, SE = 1.07) or oil palm plantations (8.8, SE = 1.07; Figure
3). Insectivores and frugivores tended to suffer relatively greater
losses, both in terms of species richness and overall abundance,
than species with mixed diets (Table 1).
species richness did not differ significantly between oil palm and
rubber, between plantations less than 1 km from forest and those
further away, or between young and mature plantations (P > 0.1
in all cases). Species richness in plantations in which a layer
of undergrowth was allowed to develop (10.04, SE = 1.05) was significantly
higher than in plantations with no undergrowth (8.09, SE = 1.05;
Figure 4). Species recorded only in forest had significantly smaller
ranges than species occurring in both forest and plantations or
plantations alone (Figure 5). There was no significant difference
in body size between species recorded only in forest and those recorded
in plantations. However, species that were recorded in both forest
and plantations had smaller body size than species recorded only
in forest (Mann– Whitney U-test, P = 0.001).
Figure 4 : Boxplot of species richness in oil palm
and rubber plantations in which undergrowth was present and absent.
Interpretation of the boxplots is as in Figure 3. The median for
the group “Absent” falls at the lower limit of the interquartile
range. GLMs indicated that the means differed significantly (F1,58
= 9.94, P < 0.002).
Figure 5 : Relative range sizes in Thailand of
species recorded only in forest (n = 68), only in plantations (n
= 17) or both (n = 26). Non-breeding species were excluded. There
was a significant difference between the three groups of species
(Kruskal–Wallis test, adjusted for ties; H = 27.9, df = 2,
P < 0.0001). In pairwise tests, species occurring only in forest
had significantly smaller national ranges than species occurring
only in plantations or species occurring in both, which did not
differ significantly from each other. Interpretation of the boxplots
is as in Figure 3. The median for the “Plantations only”
group falls along the upper limit of the interquartile range.
the 16 Globally Threatened or Near-Threatened species recorded (BirdLife
International 2000), only one was recorded outside forest. The eight
species listed by Round and Treesucon (1996) as being confined to
extreme lowland areas were also recorded only in forest. The distribution
of bird species by family is shown in the Appendix; groups well
represented in forest but absent or poorly represented in plantations
included woodpeckers, barbets, broadbills, leafbirds and babblers.
The mean relative
abundance of species occurring in both forest and plantations did
not differ significantly between the two habitat types (Mann–Whitney
U-test, P > 0.1). Mean species relative abundance did not differ
significantly between oil palm and rubber. Species composition and
abundance were extremely similar in oil palm and rubber plantations,
and there was a highly significant positive correlation between
mean relative abundances in each plantation type across the 51 species
recorded in at least one plantation type (Figure 6). Communities
in both plantation types were numerically dominated by just two
or three species.
6 : Mean relative abundance (l) in oil palm and in rubber
plantations for each of the 51 species recorded in one or both plantation
types. The two were significantly correlated (r = 0.88, n = 51, P
The results indicate that conversion of forest to commercial plantations
of oil palm or rubber results in two major changes to bird communities:
the replacement of species rich communities with species-poor communities,
and the replacement of threatened and range-restricted species by
species of lower conservation concern and with extensive ranges.
The combination of these processes means that forest conversion
to plantations represents a severe threat of biodiversity, with
the resulting habitats supporting communities dominated by a small
number of common and widespread species. Almost all species of conservation
concern disappeared following forest conversion to plantations.
These patterns match results of work undertaken in a number of agricultural
systems in different parts of the world (e.g. Pomeroy and Dranzoa
1997, Waltert et al. 2004) and appear to have some generality
across regions and agricultural systems (Donald 2004).
The demand for
agricultural products, including palm oil and rubber, is likely
to increase in line with an increasing world population, representing
probably the greatest single threat to the world’s birds (BirdLife
International 2004). Palm oil production is expected to double between
2004 and 2020. A gradient of options exists to meet this demand,
ranging between the extremes of meeting all needs by expanding the
area of land covered by low-intensity agriculture (the “wildlife-friendly
farming” option of Green et al., 2005) to maximizing
yield and keeping the area of land under production to a minimum
(“land sparing”). Determining the point along this gradient
that has the minimum impact on biodiversity requires consideration
of the responses of different species to different levels of agricultural
intensity (Green et al. 2005). In South-East Asia, almost
all oil palm and rubber plantations are in the form of intensive
monocultures (Clay 2004), and there is little history in the region
of sympathetic agroforestry systems, such as are found in some rubber
production systems in Indonesia (Joshi et al. 2003) and Brazil (Schroth
et al. 2003). Oil palm and rubber production strategies
in South-East Asia are not commercially conducive to low-intensity
mixed agroforestry systems (Clay 2004). There therefore appear to
be few opportunities to reduce the biodiversity impacts of these
crops by increasing biodiversity in plantations. The bird communities
in the structurally very different rubber and oil palm plantations
were found to be remarkably similar at all stages of crop development,
suggesting that planting a mixture of crops or age groups would
have few positive effects on species richness in agricultural landscapes.
Allowing undergrowth to develop under the crop did result in a significant,
albeit small, increase in species richness. This increase was, however,
wholly insufficient to compensate for the loss of species following
Given the lack
of management options within plantations, it would appear that protecting
pristine habitats and reducing the need to clear new land by maximizing
yield within existing systems (“land sparing”) would
appear to be more efficient strategies than trying to increase the
biodiversity value of plantations through methods such as those
reviewed by Hartley (2002). So far, increases in palm oil production
have been brought about largely by increasing the area over which
it is grown, and yields per unit area have remained relatively stable
over the last decade; in Malaysia, the area planted increased by
over 70% between 1994 and 2004, whereas yield increased over the
same period by less than 1%. However, yields can be greatly increased
through selective breeding and by developing hybrid crosses between
African and American palm species; the yield of oil palm in Costa
Rica, for example, is 3 times the global average (Clay 2004). It
seems inevitable that the only way to meet current palm oil production
targets while still conserving some lowland forests in South-East
Asia will be to increase yields. Conservationists may find themselves
in the unexpected position of having to encourage the palm oil and
rubber industries to intensify production and increase profitability
in existing plantations, as they have done previously in low-intensity
rubber systems (Joshi et al. 2003, Schroth et al.
2004). Such a strategy would need to be backed by increased protection
for non-agricultural habitats (Niesten et al. 2004).
For help with fieldwork we thank Yotin Meekaeo, Malcolm and Sheena
Davies and the staff of the Khao-Pra Bang-Kram Wildlife Sanctuary.
For helpful comments on previous drafts we thank Nigel Collar, Dieter
Hoffmann, Fiona Sanderson, Jo Phillips, Debbie Pain, Stuart Marsden
and two anonymous referees.
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Taxonomic breakdown of species recorded in forest and in plantations
of oil palm and rubber, listing number of species recorded in forest
and plantations by bird family. Taxonomy follows the Sibley and
of families represented
Aratrakorn, Bird Conservation Society of Thailand, 43 Soi Chokchai
Ruammitr, Vibhavadee Road, Dindaeng, Bangkok 10320, Thailand.
Thunhikorn, National Parks, Wildlife and Plant Conservation Department,
61 Phahonyothin Road, Ladyaow, Jatuchak, Bangkok 10900, Thailand.
Donald, Royal Society for the Protection of Birds, The Lodge, Sandy,
Bedfordshire SG19 2DL, U.K. E-mail: email@example.com
*Author to whom
correspondence should be sent.