Changes in Tree Species Abundance in a Neotropical Forest: Impact of Climate Change Author(s): Richard Condit, Stephen P. Hubbell and Robin B. Foster
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Source: Journal of Tropical Ecology, Vol. 12, No. 2 (Mar., 1996), pp. 231-256 Published by: Cambridge University Press
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Jou1rnal of Tropical Ecology (1996) 12:231-256. With 2 figures Copyright (C 1996 Cambridge University Press
Changes in tree species abundance in a
Neotropical forest: impact of climate change
RICHARD CONDIT*, STEPHEN P. HUBBELL*t and ROBIN B.
FOSTER* +
* Smithsonian Tropical Research Institute, Unit 0948, APO AA 34002-0948, USA; or Apartado 2072, Balboa, Ancon, Panama'
t Department of Ecology, Evolution and Behavior, Princeton University, Princeton, New Jersey 08544, USA
+ Department of Botany, Field AMuseum of Natural History, Chicago, Illinois 60605, USA
ABSTRACT. The abundance of all tree and shrub species has been monitored for eight years in a 50 ha census plot in tropical moist forest in central Panama. Here we examine population trends of the 219 most numerous species in the plot, assessing the impact of a long-term drying trend.
Population change was calculated as the mean rate of increase (or decrease) over eight years, considering either all stems :10 mm diameter at breast height (dbh) or just stems :100 mm dbh.
For stems B 10 mm, 40% of the species had mean growth rates <1 % per year (either increasing or decreasing) and 12%l had changes B5/5% per year. For stems 100 mm, the figures were 38%
anid 8%.
Species that specialize on the slopes of the plot, a moist microhabitat relative to the plateau, suffered significanitly more declines in abundance than species that did not prefer slopes (stems 10 mm dbh). This pattern was due entirely to species of small stature: 91 % of treelets and shrubs that were slope-specialists declined in abundance, but just 19% of non-slope treelets and shrubs declined. Among larger trees, slope and non-slope species fared equally. For stems I100 mm dbh, the slope effect vanished because there were few shrubs and treelets with stems I100 mm dbh.
Another edaphic guild of species, those occurring preferentially in a small swamp in the centre of the plot, were no mor-e likely to decline in abundance than non-swamp species, regardless of growth form. Species that preferentially colonize canopy gaps in the plot were slightly more likely to decrease in abundance than non-colonizing species (only for stems B10 mm dbh, not B100 mm).
Despite this overall trend, however, several colonizing species had the most rapidly increasing populations in the plot.
The impact of a 25-year drying trend and an associated increase in the severity of the 4-month dry season is having an obvious impact on the BCI forest. At least 16 species of shrubs and treelets with affinities for moist microhabitats are headed for extinctioin in the plot. Presumably, these species invaded the forest during a wetter period prior to 1966. A severe drought of 1983 that caused ulnusually high tr-ee mortality contributed to this trend, anid may also have been responsible for sharp increases in abundance of a few gap-colonizers because it temporarily opened the forest canopy. The BCI forest is remarkably sensitive to a subtle climatic shift, yet we do not know whether this is typical for tropical folrests because no other lalge-scale censuses exist fol comparisonl.
KEY WORDS: demography. population dynamics, tropical populations
INTRODUCTION
The myth that tropical climates provide a stable environment for tropical forest organisms has long been buried. We know that annual shifts in moisture avail- ability can stress vegetation and limit the distribution of many species
231
232 RICHARD CONDIT ET AL.
(Hartshorn 1992, Wright 1992). Now, thanks to long-term weather records kept by the Panama Canal Commission, we also know that the vegetation of Barro Colorado Island (BCI) must face supra-annual shifts in moisture availability (Windsor 1990, Windsor et al. 1990). Total precipitation at BCI underwent an abrupt decline around 1965, averaging 2740 mm prior to 1965 and 2430 mm since, paralleling a worldwide reduction in rainfall in the northern tropics (Bradley et al. 1987, Diaz et al. 1989). Moreover, associated with a strong El Nino event, the annual dry season of 1983 was unusually long and severe, causing elevated tree mortality (Condit et al. 1992b, Condit et al. 1995, Leigh et al. 1990). The 1983 drought and the long-term drop in rainfall are undoubtedly part of the same phenomenon, as the frequency of dry seasons (15 December to 15 April) receiving less than 100 mm of rain increased from once every 6.2 years prior to 1965 to once every 3.5 years since (Windsor 1990).
How does the composition of a tropical forest change when precipitation patterns shift? Unusual droughts and variation in rainfall are recognized more and more as important in tropical forests (Foster 1982a, Hartshorn 1992, Leigh et al. 1990, Woods 1989), but we know little about how populations of individual species change as a result. Certainly we know that past climate changes have led to shifts in species' distributions (Bush & Colinvaux 1990, Bush et al. 1990, Hamilton & Taylor 1991, Sukumar et al. 1993), and in temperate forests, detailed descriptions of range shifts that accompany past climate changes are so well documented (Davis 1981, Delcourt & Delcourt, 1987) that precise pre- dictions on the impact of future climate scenarios can be made (Botkin & Nisbet 1992, Dale & Franklin 1989, Franklin et al. 1992, Overpeck et al. 1990, Pastor &
Post 1988, Shugart & Smith 1992, Solomon 1986, Urban et al. 1993). The species-specific information behind these predictions is not available for most tropical forests. Only for the Luquillo forest in Puerto Rico has a species-specific model been used to predict the impact of climate change; O'Brien et al. (1992) assessed the potential impact of increasing hurricane frequency.
To gather information on many individual species of tropical trees in one community, we established a large-scale and long-term population survey of forest at Barro Colorado Island in Panama. In order to include substantive information on populations of many species, a large plot was mapped: 50 ha of forest, with all stems above 10 mm in diameter included (Condit 1995, Hub- bell & Foster 1983). This dataset provides detailed information on change in forest composition and its relation to climate change. Similar large plots in natural forest are now being censused in India, Malaysia, Thailand, Sri Lanka, Puerto Rico, Ecuador, Cameroon and Zaire (Condit 1995, Manokaran et al.
1992, Sukumar et al. 1992, Zimmerman et al. 1994), so we will soon be able to make a worldwide assessment on the lability of the species composition of trop- ical forests.
Here we provide population estimates for 313 species of tropical trees found in the 50 ha plot in Panama between 1982 and 1990. We address specific hypo- theses about how the community is changing, in particular, how it might be
Climate variation and tropical forest change 233 affected by long-term reduction in rainfall (Condit et al. 1992b, Hubbell &
Foster 1990a, 1992). First, we consider species whose distributions within the plot are associated with moist microhabitats: a seasonal swamp and the moder- ately sloping terrain that drops off from the plateau in the centre of the plot (Becker et al. 1988, Hubbell & Foster 1983, 1986a). These areas remain wet throughout most dry seasons because a basalt cap below the plateau accumu- lates water during the wet season and drains slowly into the swamp and slopes throughout the dry season. Our prediction is that species associated with the moist microsites should be especially sensitive to the overall drying trend and will have suffered disproportionate losses in population.
In addition, we consider population changes of species that preferentially colonize light gaps within the forest. There are two different predictions about colonizing species. First, Hubbell & Foster (1990a, 1992) suggested that the plot is undergoing a slow loss of weedy species, because the region just north of the plot (plus 2 ha within the plot) was cleared of forest about 90 years ago, and has since regrown. Colonizing species probably gained abundance within the old forest because of their large populations just outside, and are now declining. If this is the case, we should be able to detect disporportionate popu- lation declines among colonizing species. The second prediction on colonizers is just the opposite, and is based on the observation that the drought opened the forest canopy briefly during 1983 (Becker & Smith 1990). With more light reaching the ground, colonizing species should increase in abundance. Popula- tion changes of species preferring gaps can tell us which of the potentially opposing forces is more important.
MATERIALS AND METHODS
Study site
Barro Colorado (BCI) is a 1500 ha island that was a hilltop until the Panama Canal was finished in 1914. The island is part of the Barro Colorado Nature Monument and has been operated as a research reserve since 1923. It is entirely forested, most in old-growth forest with no signs of human disturbance for over 500 years: 48 ha of the 50 ha plot are in old-growth, with 2 ha in an area cleared until about 1900 as part of a French settlement. Temperatures are uniform year-round at BCI, but rainfall is seasonal, with almost none falling between mid-December and mid-April. Details on climate, flora and fauna can be found in Croat (1978) and Leigh et al. (1982).
Census
A 50 ha plot on the top of the island was fully censused in 1981-1983, 1985 and 1990 (Condit et al. 1992a,b, 1993a,b, Hubbell & Foster 1983, 1986a,b, 1987, 1990a,b, 1992); we refer to the first census, which lasted two years, as the 1982 census. All free-standing, woody stems ? 10 mm diameter at breast height (dbh) were identified, tagged and mapped. The diameter of each stem
234 RICHARD CONDIT ET AL.
was measured at breast height (1.3 m) unless there were irregularities in the trunk there, in which case the measurement was taken at the nearest point downward where the stem was cylindrical. Dbhs of buttressed trees were taken above the buttresses. There were about 242,000 living stems in each census (Hubbell & Foster 1990a), and 305,875 stems over all three censuses; 28 have not been identified to species. A total of 313 species have been identified: 304, 306 and 303 in successive censuses. (Three new species have been added since Condit et al. 1992b, all rare plants that had been misidentified as more common species.) Included in the list of 313 is a single tree that appeared to be a hybrid between Apeiba membranacea and A. tibourbou, and two distinct varieties of Swartzia simplex (Croat 1978).
Analyses
Species included. Abundances for all 313 species are reported. Species' names match those from Croat (1978) and D'Arcy (1987), except for species which were discovered, or whose names have been changed, since. An Appendix lists all cases where names do not match those found in Croat (1978) or D'Arcy (1987), and allows any species listed here to be located in those floras or in our previous publications on the 50 ha plot. Authorities for all species can be found via these references.
Tests of hypotheses about changes in abundance included only those species that had at least 20 individuals 10 mm dbh in at least one of the censuses.
We used this cutoff because large percentage changes in abundance of very rare species could be caused by minor, chance events. Four species of Bactris palms were also eliminated from analyses because we changed methods for counting individuals of these species. This left 219 species for analyses of all stems
?10 mm dbh. Analyses were then repeated for changes in the number of indi- viduals >100 mm dbh, including the 136 species that had at least 20 stems in at least one census. We included an analysis with this larger cutoff because many other studies of tropical forest use the 100 mm limit (Phillips & Gentry 1994, Phillips et al. 1994).
Species characteristics. We analysed changes in abundance as they correlated with three species characteristics - growth form, moisture preference and tendency to recruit into light gaps. Species were divided into four growth forms - large trees (?20 m tall), mid-sized trees (10-20 m), treelets (4-10 m) and shrubs (1- 4 m) - based on the maximum height attained at BCI (Hubbell & Foster
1986a). Moisture regime was defined using the slopes in the 50 ha plot, which have higher soil moisture content during the dry season than the plateau above them (Becker et al. 1988), and the swamp, which is flooded throughout the wet season and remains moist in the dry season (Hubbell & Foster 1986a).
Many species have distributions clearly demarcated by the slopes and the swamp (Hubbell & Foster 1986a), and we calculated the density of all species in the different habitats (unpublished data). We used the ratio of density on
Climate variation and tropical forest change 235 the slopes (all 20 m X 20 m quadrats inclined >7Q) to density on the lower part of the plateau (quadrats with slope <7? and elevation <152 m, excluding the swamp) as an index of 'slope-specialization', and the ratio of density in the swamp (all 20 m X 20 m quadrats holding standing water through most of the wet season) to density on the lower plateau as an index for 'swamp- specialization'. We considered 'slope-specialists' and 'swamp-specialists' species with ratios > 1.5; this cutoff was chosen because chi-squared tests showed that nearly all higher ratios were significantly different from 1.0 (P<0.01), while most below did not (unpublished data). This index was preferable to a definition based on statistical significance, because the latter is sensitive to sample size.
Finally, as a 'colonizing index' for each species, we used the fraction of recruits in light gaps given in Welden et al. (1991): Hubbell & Foster (1986b) used a similar but not identical 'index of heliophily'. Colonizers were defined as those species with an index ?30; again, this corresponds roughly with a statistically significant preference for recruiting in gaps (Welden et al. 1991) but does not depend on sample size. Most colonizers are probably 'pioneers' as defined by Swaine & Whitmore (1988), but they emphasized seed germination character- istics, which we do not consider here. Species for which information was lacking were omitted from all analyses requiring that information. The slope and swamp indices were calculated for all but 27 of the very rare species, but the colonizing index was available for only the 156 species listed in Welden et al. (1991).
Statistical tests. In order to determine whether certain groups of species suffered disproportionate losses, the number of species that increased or decreased in abundance between 1982 and 1990 was tallied as a function of the four categor- ical variables. For statistical tests, a standard ANOVA was not possible because the design was unbalanced, with many empty cells. Instead, chi-squared tests were used on each of the variables: for example, a 2 x 2 contingency table for slope-specialization category and for population change provided a chi-squared statistic with one degree of freedom. To determine effects of each variable separ- ately, we proceeded as follows. Swamp effect was assessed by contingency tables for swamp and non-swamp species; since swamp status was not associated with colonizing nor slope status, the latter two categories were simply ignored when swamp effect was tested. But slope and colonizing variables were associated - there were fewer slope-colonizer species than expected by chance - so we segreg- ated species simultaneously by both categories. All tests were carried out on the four growth forms separately.
For each species, we calculated the annualized rate of population change (r) using a standard model of exponential population growth:
In Nt -In No r = t
where Nt and No are population sizes at time t and time 0 and In means the natural logarithm. The time interval, t, for each species was defined as the
236 RICHARD CONDIT ET AL.
arithmetic mean time elapsed between censuses for individuals of that species (based on the census data of each 20 m x 20 m quadrat in the plot).
Earlier publications
Hubbell & Foster (1990a, 1992) described population changes based on the 1982-1985 interval, and Condit et al. (1992b) updated this with 1990 data, but this is the first presentation on abundance for all 313 species. Discrepancies between the numbers reported here and those from earlier reports are slight and are due solely to corrections of old errors. Since this is an on-going process, future reports might give figures slightly different from those reported here.
Access to data
We hope that the table of abundances for 313 species provided here will be useful for many future studies, and we will provide computer versions of the table to anyone interested. Please send us a diskette and indicate preferred formats.
RESULTS
Changes in abundance for stems 10 mm dbh
Of the 219 more common species considered here, 105 had increases in stem number between 1982 and 1990, 108 had decreases and six did not change. For all 313 species in the plot, 136 increased, 154 decreased and 23 did not change.
As previously noted (Condit et al. 1992b, Hubbell & Foster 1990a, 1992), rare species - in this case the 94 having fewer than 20 stems in all censuses - suffered proportionally more declines than common species. Table 1 gives the abundance in all three censuses for all 313 species.
Many populations did not change by much (Figure 1). Of 219 species, 88 (40%) had population changes <1% per year between 1982 and 1990. But some species had dramatic changes in abundance: 27 species (12%) changed at rates more than 5% per year, 18 declining and nine increasing (Table 2).
Some common species underwent substantial declines. Poulsenia armata fell from 3430 to 2126 stems, and Acalypha diversifolia from 1568 to 827. The most rapid rate of decline was Piper aequale, which had 219 stems in 1982 and 83 in 1990 (Table 2). On the other hand, the population of Palicourea guianensis rose from 377 to 1475 stems, while the much less common Psychotria gracifora had the biggest rate of increase, from 10 to 44 stems over eight years (Table 2). The mean rate of change for the 219 species was -0.29% per year, while the mean rate of absolute change was 2.25% (the mean of the absolute values of rates of change).
Changes in abundance for stems ?100 mm dbh
The range of population change among larger stems was no different than for smaller (Figure 1B). Of the 136 species considered, 66 increased in abund- ance from 1982 to 1990, 61 decreased and seven stayed the same. Fifty-one
Climate variation and tropical forest change 237 A) 30
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Figur-e 1. Distribution of population growth r-ates. The 'colonizer' category inicludes all colonizing species that were not slope-specialists, and the 'slope' category includes slope species that were non-colonizing;
'slope-colonizers' are the four specializing in both areas. 'Other' inicludes all species that were neither slope nor colonizing plus all species that wer-e missing information on one or bothi categor-ies. (A) Rate of change of populations of stemns 310 mm dbh, including 219 species (see text). (B) Rate of change of populations of stems 3100 mm dbh, including 136 species (see text).
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