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Journal of 









, 1003–1010


© 2006 The Authors

Journal compilation

© 2006 British 

Ecological Society


Blackwell Publishing Ltd


Less diverse forest is more resistant to hurricane 

disturbance: evidence from montane rain forests in Jamaica




Department of Plant Sciences, University of Cambridge, Downing Street, Cambridge CB2 3EA, UK, and 





Research, PO Box 69, Lincoln 8152, New Zealand 





Are more diverse ecosystems more or less resistant to disturbance? Does diversity

increase in a forest after being hit by a hurricane? We answer these questions using a 30-

year study of four Jamaican forests, which differ in soil fertility and diversity, and which

were hit by Hurricane Gilbert in 1988; the decades were: pre-Gilbert (1974 –84), Gilbert

(1984 –94), and post-Gilbert (1994 –2004).




Diversity (Shannon index) was always higher in the three forests (Col H



 3.00, Mull




 2.91 and Slope H



 2.99) on more fertile soils (C : N ratios 10–13, N : P 16 –24), and

significantly lower in the Mor forest (H



 2.26) with the least fertile soil (C : N ratio 24,

N : P ratio 44). Diversity increased during the Gilbert decade in two of  the more diverse

forests (Mull and Slope), it did not increase in the least diverse, Mor forest. The overall

increase in diversity during the Gilbert decade was due to the recruitment of eight,

mostly light-demanding, species and the increased abundance of uncommon species.




We used turnover rates (the average of  mortality and recruitment of  stems) as a

measure of  resistance. We equate low turnover with high resistance to hurricane

damage. Turnover increased during the Gilbert decade in all forests, but increased more

in the three more diverse forests (Mull 1.5% year





 1974 –84 to 3.1% year





 1984 –94;

Slope 1.3–2.6; Col 1.5–3.2); than in the least diverse Mor forest (1.2–1.9).




Stem diameter growth rates pre-Gilbert were very low in all forests and were lowest

in the Mor forest (Mor 0.3 mm year





, Mull 0.4, Slope 0.5, Col 0.6). They increased

during the Gilbert decade and remained, in the post-Gilbert decade, double those of the

pre-Gilbert decade (Mor 0.6 mm year





, Mull 0.6, Slope 0.8, Col 1.1). Smaller stems

increased growth more than larger stems. The stems recruited during the Gilbert and

post-Gilbert decades grew faster than those present in 1974.




Thus, in montane forest in Jamaica the least diverse forest was most resistant to

hurricane damage, and although there was a strong similarity in species rank abundances

over 30 years including a hurricane, the hurricane increased diversity.




: diversity, hurricanes, mortality, recruitment, resistance 


Journal of Ecology






, 1003–1010

doi: 10.1111/j.1365-2745.2006.01149.x




Natural disturbances affect all ecosystems, killing all

or some resident organisms and disrupting ecosystem

processes. With global loss of biodiversity a pressing

issue, determination of relationships between ecosystem

diversity and response to disturbance is of  particular

interest: this has been the subject of studies over many

years (MacArthur 1955; Hooper 


et al


. 2005). However

predictions of  relationships between diversity and

response to disturbance remain elusive. While one recent

review concluded that ‘communities with greater diversity

tend to be stable’, with stable meaning ‘less oscillatory

and less susceptible to invasion by exotic species’



et al


. 2001), others have concluded the opposite

(Goodman 1975).

Responses to disturbance may be anywhere in the

space defined by the following two axes: (i) resistance

(‘staying essentially unchanged’, Grimm & Wissel 1997);

and (ii) resilience (‘returning to a reference state’, Grimm


Correspondence: E. V. J. Tanner (e-mail




E. V. J. Tanner & 

P. J. Bellingham


© 2006 The Authors

Journal compilation 

© 2006 British 

Ecological Society, 


Journal of Ecology








, 1003–1010


& Wissel 1997). Disturbances may be of  many kinds,

for example wind, drought, flooding, fire, freezing and

herbivores, and attempts to generalize about how

ecosystems respond to disturbances, across different

disturbances and different ecosystems, are likely to

have many exceptions. If for no other reason, because

different disturbances cause different communities, for

example in an area of hundreds of  hectares, with access

to the same species pool, an ecosystem disturbed by wind

is likely to be very different to an ecosystem disturbed by

flooding. In addition, some recent experimental studies

of the relationship between diversity and ecosystem

properties have studied the first disturbance to affect

the ecosystem (Tilman 1996). It is likely that natural

ecosystems have recurrent disturbances and that

susceptible species will have become locally extinct, or

have small population sizes, by the time a study of a

particular natural disturbance event is made. Thus,

patterns in nature will probably be very different from

patterns discovered in relatively short-term experiments.

Finally, it has been shown that it is not diversity itself

that defines ecosystem responses to disturbance but the

properties of  the species in the ecosystem (Sankaran &

McNaughton 1999).

There are very few studies of how natural systems that

differ in diversity are affected by natural disturbances,

and even fewer of  forests despite the likelihood that

major disturbances determine the species composition

of large areas of forest (Oliver 1980). Tropical rain forests

are among the earth’s most diverse ecosystems, and many

are subject to major disturbances such as hurricanes,

which have major effects on their growth and composition

(Everham & Brokaw 1996). Research into the effects of

hurricanes shows that while some of the initial effects

are destructive, some of the longer term, decadal, effects

may be positive, for example increased growth of surviving

trees has been reported following hurricanes (Merrens

& Peart 1992; Bellingham 


et al


. 1995; Scatena 


et al


. 1996;

Batista & Platt 2003).

The effects of disturbance on diversity in tropical rain

forests have been less well studied. In Kolombangara,

plots with higher diversity suffered more canopy damage

and had higher recruitment and turnover of  the 12

commonest species (only 12 species were recorded after

1964) following two cyclones in the 6 years after the

start of  the study in 1964 (Burslem & Whitmore 1999).

In Nicaragua a comparison of  diversity between five

forests, four in one area hit by a hurricane and one in an

area not hit by the hurricane, showed higher diversity in

the hurricane damaged forests (Vandermeer 


et al


. 2000).

A 54-year study of a plot in Puerto Rico concluded that

the high diversity when the plot was established was due

to a hurricane 15 years previously, though a hurricane hit

in 1998 caused no increase in diversity in that plot by 2000

(Weaver 2002). None of these studies have unambiguously

shown increased diversity caused by a hurricane (or

cyclone) for different reasons: Burslem & Whitmore (1999)

had no complete species records after their cyclones;



et al


. (2000) had no pre-hurricane records;

and though Weaver (2002) had pre- and post-hurricane

records he found no increase in diversity.

Our study has both pre- and post-hurricane records

collected over a 30-year period from permanent plots

in four Jamaican montane rain forests. We address the

question of  whether there is a relationship between tree

species diversity (on the one hand) and resistance and

resilience (on the other), using a natural gradient of

diversity/fertility and a natural disturbance, a hurricane.

In addition to the between-forest comparison of diversity

and resistance and resilience we report on the overall

effect of the hurricane on diversity and growth, by using

the combined data set from the four forests. Thus, our

null hypotheses are: (i) that there was no relationship

between diversity on the one hand and resistance and

resilience on the other; (ii) that diversity was unchanged

by hurricane damage; and (iii) that growth was unaffected

by hurricane damage.




,   


We studied four Jamaican upper montane rain forests

within 300 m of  each other (18







 N, 76







 W; 1580–

1600  m): Col forest (Gap forest of Tanner 1977, 0.09  ha

sampled); Wet Slope forest (0.1 ha); Mull Ridge forest

(0.1 ha) and Mor Ridge forest (0.06 ha); hereafter

abbreviated to Col, Slope, Mull and Mor (the ‘forests’

in the current paper are the ‘sites’ of Tanner 1977). The

forests were selected in 1974 (for a study of  nutrient

cycling) as representative of forests in the western Blue

Mountains, the first 10 




 10 m plot in each forest was

subjectively positioned in ‘representative’ forest, the

other plots in each forest were contiguous. All stems



 3 cm d.b.h. (at 1.3 m) in permanent 10 




 10 m plots

were tagged, painted with a ring at 1.3 m, identified, and

measured in 1974, 1984, 1989, 1991, 1994 and 2004. Multi-

stemmed trees were recorded. In total we had 2745 stems,

2171 individuals and 68 species; nomenclature follows

Adams (1972), except where other authorities are listed.

Soil (0 –10 cm depth) total C, N (CHN analyser), P (acid

digest) and Bray 1-extractable P were determined by

Brookside Laboratory Association Inc. (Knoxville,

Ohio, USA) from cores collected and air-dried in July

and August 2004, from three randomly chosen plots

per forest.


 


Hurricane eyes pass over, or within 15 km of, the Blue

Mountains of Jamaica on average every 25 years. During

our study period (1974–2004) Hurricane Gilbert struck

in 1988; it was the most powerful hurricane recorded in

the Caribbean during the 20th century (Dodge 


et al



1999). The hurricane killed about 2% of stems in these

forests ‘immediately’ (Bellingham 


et al


. 1992) and 13%

over a 16-year period. It uprooted 5% of  stems, broke

crowns from 4% of stems and completely defoliated 19%




Resistance to 




© 2006 The Authors

Journal compilation 

© 2006 British 

Ecological Society, 


Journal of Ecology








, 1003–1010


of stems, thus increasing light availability at the forest

floor for up to 33 months after the hurricane (Bellingham


et al


. 1995, 1996). We report results over three decades

with respect to Hurricane Gilbert: pre-Gilbert decade

(1974 –84), Gilbert decade (1984 –94), and post-Gilbert

decade (1994 –2004).


 


For the tree stems (not the individuals) we report Fisher’s

alpha (








 ln(1 + 








) where 




 = number of species in

the sample, 




 = number of stems in the sample, 




 = Fisher’s

alpha) and the Shannon index, H



 (the negative sum over

all species of (




















), where 






 = abundance of 









 = total number of  stems in the sample). We

statistically compared H



 between (unreplicated) forests

following the worked example in Magurran (1988); in

our data cumulative H



 changed less than 5% after using

two plots in each forest. We carried out rarefaction

procedures using the ECOSIM program (Gotelli &

Entsminger 2006), in which we used the number of stems

(1716) per species in 1984 (Appendix S1 in Supplementary

Material) to estimate the number of species for 1716 stems

from the data sets for 1994 and 2004. For each decade we

calculated mortality, recruitment and turnover. Mortality

as percentage is: 100(




 = 1 – [1 – (
























), where






 = number at the beginning of a period, 






 = number

of survivors at the end of a period, and 




 = time in years.

Recruitment as percentage is: 100(




 = 1 – (1 –

























 = number of recruits during a period and 







the number at the end of the period. Turnover is the

average of mortality and recruitment. All comparisons

of mortality and recruitment were across equal decade

intervals so problems in assessing rates across different

census intervals were avoided (Kohyama & Takada 1998).

Mortality and recruitment of stems were compared

between forests using contingency tables. We used regres-

sion analysis to compare the number of  tree stems

per species in 1984 with those in 2004 and compared the

fitted line with a null hypothesis relationship of 1 : 1. We

compared the gains and losses of  species overall (all

forests combined) using a binomial test, with the null

hypothesis that the number of species recruited should

equal the number lost. We calculated Euclidian distances

to show departure from initial biomass composition as

in Lep






et al


. (1982), biomass calculated using equations

in Tanner (1980). The mean trunk diameter growth rates

were compared between forests within a decade by Mann–





-tests and between decades by paired 





To  discover whether differences in trunk growth rates

before and after the hurricane were related to trunk size,

we calculated, for the 977 trunks alive in 1974 and 1991,

the mean absolute growth in trunk diameter from 1974

to 1984 and divided it by 10, and the mean absolute

growth in trunk diameter from 1989 to 1991 and divided

it by 2. We plotted the difference between these two absolute

rates against trunk diameter in 1989 and fitted a linear

regression; we compared the intercept and slope of the

fitted line against a null hypothesis of  zero for both.




, ,   

      



Diversity was always highest in the three forests on more

fertile soils (Fig. 1) with lower C : N and C : P ratios

(Table 1); diversity was significantly lower in the Mor

with very infertile soil (comparisons of H



 for 1984: Mor

vs. Mull, 






 = 9.4, 




 < 0.001; Mor vs. Col, 






 = 10.3,




  <  0.001; Mor vs. Slope, 






 = 9.7, 




 < 0.001; all other

comparisons not significant). Diversity increased after

the hurricane in two of the three forests on more fertile

soils (Mull and Slope); in contrast, the least diverse Mor

forest showed no change in diversity throughout the

study (Fig. 1). Summed over the four forests there was

an increase in diversity in response to the hurricane

(Fig. 2, Appendix S1) due to the net recruitment of species

not previously recorded in the plots (but present, though

rare, in the surrounding forest). The increase in diversity

was a result of the increase in the number of stems because

rarefaction analyses, when the number of stems was kept

constant, gave no statistically significant increase in species

post-Gilbert. Between 1984 and 1994, eight species were

recruited and none were lost (exact binomial test, 





0.01); all these species were occasional or rare in 1994

(occasional and rare = lower 40% of species in terms of

rank abundance), except for the invasive alien tree 



rum undulatum


, which by 2004 was, in terms of individuals,

the 17th most abundant species of 65 (Appendix S1).

The changes in populations of trees resulted in a drift

Fig. 1 Fisher’s alpha and Shannon indices of  diversity for the

four forests: 

᭹ = Col; ᭢ = Slope; ᭺ = Mull; ᭞ = Mor.




E. V. J. Tanner & 

P. J. Bellingham


© 2006 The Authors

Journal compilation 

© 2006 British 

Ecological Society, 


Journal of Ecology








, 1003–1010


away from the initial biomass composition, but less so

in the least diverse Mor forest (Fig. 3).

Mortality rates did not differ between the forests in the

pre-Gilbert decade (Fig. 4, Appendix S2); they increased

in all four forests during the Gilbert decade but increased

less in the Mor and Slope than in the Mull and Col

(Fig. 4). Recruitment rates in the pre-Gilbert decade

were lower in the Mor than the Mull and Col forests

(Fig. 4, Appendix S2) and, like mortality, increased

during the Gilbert decade but much less in the Mor

(Fig. 4, Appendix S2). Thus, the three forests on more

fertile soils were the most dynamic; that is, they had

higher turnover rates (mean of mortality and recruit-

ment) than the Mor, in the Gilbert decade (Fig. 4,

Appendix S2).

Across the four forests, 38% of the stems present in

1974 had died by 2004, and 47% of the stems present in

2004 were recruits since 1974; 32% of  the basal area

present in 1974 was dead by 2004, but only 3% of the

basal area in 2004 was recruited since 1974.

Stem diameter growth was very slow in all forests,

but especially so in Mor (Mor 0.3 mm year





, Mull 0.4,

Slope 0.5, Col 0.6 from 1974 to 1984, Fig. 5); diameter

growth increased during the Gilbert decade and remained

Table 1 Soil (0–10 cm) nutrient concentrations, mean values from one core from each of  three randomly chosen plots per forest


N (%) 


C (%) 


P (%) 


C : N 


C : P 


N : P 


P (p.p.m.) 

‘Bray 1’

Bulk density 

(g cm





± 0.12


± 1.4


± 0.01


± 0.3


± 17.8


± 1.3


± 0.6


± 0.03



± 0.08


± 1.5


± 0.01


± 0.8


± 129


± 8.7


± 1.1


± 0.2



± 0.10


± 1.5


± 0.01


± 0.4


± 13.2


± 5.5


± 0.3


± 0.1



± 0.03


± 1.2


± 0.00


± 0.9


± 55


± 3.5


± 0.3


± 0.02

Fig. 2 The number of  stems per species (+1 to allow a log log

plot) in 1984 and 2004; solid line is best-fit line ( y = 0.80 (SE

0.055), x + 0.32 (SE 0.064)). The slope of the line is significantly

less than 1.00 (P < 0.05), and the intercept is significantly

greater than 1.00 (P < 0.01), the expectation if numbers of stems

per species were the same in 1984 and 2004 (dashed line); thus

species were recruited, rare and uncommon species increased

and common species were unchanged.

Fig. 3 Euclidean distance to show departure from initial

composition using data for biomass per species in each forest,

calculated as in Leps et al. (1982): 

᭹ = Col; ᭢ = Slope; ᭺ =


᭞ = Mor; arrow denotes impact of Hurricane Gilbert in


Fig. 4 (a) Percentage recruitment of  stems and (b) percentage

mortality of stems, in each of four forests in three decades;

mean per forest and standard error of the mean to show the

variation between plots within a forest. The variation between

forests (compared by contingency tables) was not significant

(P > 0.05) for mortality or recruitment from 1974 to 84, but was

significant for both mortality and recruitment in both 1984 –

94 and 1994–2004. 

᭿ = Col; ᮀ = Slope;   = Mull;   = Mor.


Resistance to 



© 2006 The Authors

Journal compilation 

© 2006 British 

Ecological Society, 

Journal of Ecology, 

94, 1003–1010

higher in the post-Gilbert decade 1994 –2004 (Mor

0.6 mm year


, Mull 0.6, Slope 0.8, Col 1.1, Fig. 5). The

increased growth was very marked in Mor and Col; pre-

Gilbert growth was significantly lower in Mor compared

with the other three forests, but in the last decade stem

growth in the Mor was not significantly lower than in

the others (Appendix S3). Furthermore, in those stems

present from 1974 to 1991, growth in the Gilbert and

post-Gilbert decades increased more in smaller than larger

stems (Fig. 6, Appendix S4). The stems recruited during

the Gilbert and post-Gilbert decades grew faster both

absolutely (compare Fig. 5a with Fig. 5b) and relative

to trunk size than those present pre-Gilbert.


     

 

Mor forest, the least diverse throughout the study, is on

the crest of the main ridge of the Blue Mountains and

was as exposed to the hurricane as any of the four forests.

Although we studied in detail only one Mor forest, which

limits the statistical tests we can carry out, it is very similar

in species composition and physiognomy to other Mor

forests within 5 km. Despite the fact that 10.1% of

surviving stems in the Mor forest were severely damaged

(defined as trees tipped up to > 40

° from their pre-

hurricane position and /or crown death) by the hurricane

(vs. 9.5% in the Mull forest, 7.0% in the Col forest, and

5.8% in the Slope forest), the Mor forest had lower

turnover than the other three forests. The low turnover

was probably a result of at least five factors. First, the

species in the Mor are less susceptible to being killed by

the hurricane. Evidence for this comes from the Mull,

where for the 277 individuals of species that also grow

in the Mor mortality was 16% (1984 –94); in contrast,

for the 245 individuals of species which grew in the

Mull but not the Mor, mortality was 35% (

 = 24.7,

P < 0.0001) (though there was no relationship between

mortality per species and wood density, just as there

was no correlation between wood density and damage

due to the hurricane, Bellingham et al. 1995). Secondly,

multiple-stemmed trees, which were more prevalent in

the Mor forest than the others (in 1974 38% of individuals

in Mor forest had multiple stems vs. 9–17% in the other

three forests; 

 = 93, P < 0.0001), have a lower chance

of dying than single stemmed trees (of 1126 individuals

with single stems 21% died between 1984 and 1994;

in contrast, of 227 trees with multiple stems only 8%


 = 14.4 P < 0.0001). Thirdly, there was a high

incidence of sprouting by most of the dominant species

of the Mor (Bellingham et al. 1994), which reduced indi-

vidual mortality. Fourthly, the Mor forest has shorter

trees; this probably resulted in less wind-caused mortality.

Fifthly, low mortality in the Mor forest was probably

partly a result of the weak, c. 50-cm deep, mor humus

that allowed the trees to flex in the wind, thus reducing

their canopy damage (although there was substantial

crown damage to one common tree in this forest, Cyrilla

racemiflora, Bellingham et al. 1995). The other half  of

‘turnover’ is recruitment, this was also lower in the Mor,

due to the very infertile, acidic (pH c. 3.0) mor humus,

which we hypothesize limits the number of species that

can grow in the forest (Grubb & Tanner 1976). Evidence

supporting this hypothesis comes from an experimental

clearing of 10 

× 10 m in the Mor forest; 10 years after

its creation the zone from which mor humus had been

removed, exposing the subsoil, had seedlings of 14 tree

Fig. 5 The absolute diameter growth rate of stems (mean 

± SE),

in each forest and in each decade; (a) stems alive in a particular

decade; (b) stems alive throughout, i.e. in 1974, 1984, 1994 and


᭿ = Col ; ᮀ = Slope;   = Mull;   = Mor.

Fig. 6 Difference between mean annual absolute stem diameter

growth 1989–91 and mean annual absolute growth 1974–84

(mm year


), by stem size classes, for those stems alive in both

1974 and 1991 (mean and SE).











E. V. J. Tanner & 

P. J. Bellingham

© 2006 The Authors

Journal compilation 

© 2006 British 

Ecological Society, 

Journal of Ecology, 

94, 1003–1010

species, and 13 seedlings m


; in contrast, undisturbed

mor humus had only six species and seven seedlings m


(Sugden et al. 1985). We conclude that characteristics

of the species dominating the Mor forest are the cause

of the high resistance and low resilience to hurricane

damage, not low diversity per se.

Mor forests are atypical in composition, structure and

response to Hurricane Gilbert (our Mor was resistant

not resilient) among Jamaican montane rain forests.

Why are resistant forests like the Mor forest not more

widespread? We think it is because in the steep Blue

Mountains there are very limited areas in which deep

acidic litter can accumulate and the process of feedback

begin and be maintained; even on ridge crests such

forests are confined to knolls. However, in places where

topography does not limit litter accumulation and where

hurricane disturbance is frequent, the ‘resistant’ syndrome

we found in Mor forest also appears to be uncommon (e.g.

lowland Nicaragua, Vandermeer et al. 2000). Nutrient-

limited forests in Hawaii were found to be more resistant

to hurricane damage and less nutrient-limited forests

more resilient (Herbert et al. 1999). We propose that at a

landscape scale, nutrient supplies have to be limiting (as

shown by low leaf nutrient concentrations in montane

rain forests), which promotes the colonization by trees

typical of  low nutrient systems (e.g. the Ericales), and

within such areas some sites will be especially nutrient

poor and by feedback mechanisms generate mor soils

and forests with their distinct species composition, which

are resistant to hurricane damage.

     


Forests on more fertile soils than Mor forests were

characteristically resilient to hurricane damage. In these

forests, diversity increased after hurricane disturbance

as a result of  an increase in abundance of uncommon

species and recruitment of species not previously recorded

in the plots. These two phenomena are the same; it is a

matter of  the scale at which the effect is measured. If the

plots had included all the forest in the area, the effect

would have been only an increase in abundance of the

rare  and uncommon species. The promotion of the

uncommon species was a result of  two patterns. First,

the canopy trees present in 1974 were mostly of more

light-demanding species that largely lacked seedlings in

the understorey. Of  the 10 most common trees, two had

no seedlings and five had few seedlings before Hurricane

Gilbert (Healey 1990). Secondly, the damage to the

canopy trees allowed more light to reach the lower

layers of  the forest, promoting germination of  seeds of

light-demanding species and the growth of the seedlings

and saplings of the advance regeneration into the recorded

minimum size class (3 cm d.b.h.). Most of the species

that increased in abundance during the Gilbert decade

were light-demanding species (for example Hedyosmum

arborescens and Alchornea latifolia), though one very

shade-tolerant tree also increased (Eugenia virgultosa).

Furthermore, we judge that most of the nine species that

were recruited after the hurricane were light demanding,

although we have little detailed information about their

ecology (almost by definition because they were mostly

rare): two are definitely light demanding (Brunellia

comocladifolia and Miconia dodecandra), four are likely

to be light demanding (Cestrum hirtumCritonia parvi-

flora,  Rhamnus sphaerospermus  and  Sapium harrisii),

and two we cannot judge. However, one common

recruit,  Pittosporum undulatum, which is an invasive

alien, has very shade-tolerant seedlings (Bellingham

et al. 2005). Thus, the hurricane mostly promoted light-

demanding species, but at least two shade-tolerant

species also increased in abundance.

We think it likely that many hurricane-affected forests

will show an overall increase in diversity after hurricanes.

Indeed, any forest where the canopy trees have a lower

diversity than the advance regeneration, a pattern true

for most tropical forests on a scale of hundreds of square

metres, is likely to show an increase in diversity after

major wind storms, which severely damage the canopy

and increase light and thus growth of the more species-

rich advance regeneration. Such a response follows general

predictive models of short-term responses to disturbance

(Connell 1978; Huston 1979) and agrees with predictions

from a modelling study of higher diversity in hurricane-

damaged forest in Puerto Rico (Doyle 1981).

The theoretical predictions that hurricane disturbance

will increase diversity find some support in the literature.

Hurricane-affected forests had higher diversity com-

pared with a less affected forest in lowland Nicaragua

(Vandermeer et al. 2000). In Kolombangara, plots that

were more damaged by cyclones had increased recruit-

ment and turnover of the 12 commonest species, but

effects on overall diversity were not recorded (Burslem

& Whitmore 1999). However, a 54-year study of diversity

(1946–2000) before and after hurricane damage in a

permanent plot in Puerto Rico did not show an increase

in diversity following two hurricanes, in 1989 and in 1998

(Weaver 2002). The lack of increase might be due to the

short interval, in which case later enumerations may

reveal an increase, or it could be due to the dominance of

two very light-demanding species (Cecropia schreberiana

and Psychotria berteriana) among the trees recruited after

the hurricane. If so, there may be areas in that forest

where increased light due to canopy damage was enough

to strongly promote the growth of pre-existing seedlings

but not so high that the seedling populations were

dominated by one or two strongly light-demanding

species. Short-term increases in diversity have also been

found after other kinds of disturbance in tropical rain

forests, such as logging, again due in part to recruitment

of light-demanding species (Molino & Sabatier 2001).

Recently it has been shown that in tropical rain forests

without large disturbances, populations of rare species,

in small quadrats, survive and grow better than those of

common species (Wills et al. 2006); however, note that

this study differs from the previous studies because

recruits were not included in the analysis by Wills et al.


Resistance to 



© 2006 The Authors

Journal compilation 

© 2006 British 

Ecological Society, 

Journal of Ecology, 

94, 1003–1010

Thus it seems that rare species fare better than common

species in a whole range of  disturbance regimes.

     

      


Despite the significant increase in rare and uncommon

species in Jamaica, and despite the fact that approxi-

mately half  of  the stems present in 2004 were recruits

since 1974 (and 32% of the basal area present in 1974 was

dead by 2004), there was a strong rank correlation in

species abundance between 1974 and 2004 (r

s 69

 = 0.86,

P < 0.001). The strong correlation in species composition

in Jamaica, across 30 years including a major hurricane,

is similar to the pattern across 54 years in Puerto Rico

(with hurricanes 45 and 52 years into the study, Weaver

2002), and similar to the pattern for the 12 most common

species across 30 years in Kolombangara (with four

cyclones, only two of  which damaged the forests, 3–

6 years after the study started, Burslem et al. 2000).

Thus, even in hurricane-damaged forests the major

pattern is for species composition to change little over


     

 ‒ 

Our results from Jamaica show more resistance to

hurricane damage in our least diverse forest. A similar

pattern was found across two forests in Puerto Rico, where

higher-altitude lower-diversity forest was more resistant

than higher-diversity lower-altitude forest (Walker et al.

1996). In Jamaica, resilience was also lower in the least

diverse forest when judged by recruitment of stems during

the Gilbert decade, though perhaps this is best seen as a

delayed resilience because recruitment increased in the

Mor in the post-Gilbert decade. Growth increase in the

surviving stems post-Gilbert, another form of resilience,

was also higher in Mor forest. Thus, our results do not

support the conclusion of  Tilman et al. (2001) that

‘communities with greater diversity tend to be stable’.

Hoping for a generalization across very different

ecosystem types (forest to grasslands, e.g. this study vs.

Kahmen et al. 2005), and across different disturbances

(wind to drought), and across natural systems previously

exposed to disturbances as compared with experiments

where the first disturbance is studied, is probably too

much, though it will probably be possible to generalize

for subsets of  these categories. Our results from forest

in one area of the world show that low diversity was

associated with high resistance to disturbance by wind.


In summary, although many initial effects of hurricanes

on forests are destructive, because trees are destroyed,

crown areas reduced and emergent crowns destroyed

(Brokaw et al. 2004), this very destruction increases light

for smaller trees, thus increasing growth, and in this

study, diversity. Comparisons between forest differing

in diversity show that if there is an emerging pattern, it

is that least diverse forests are more resistant to severe

wind damage.


Earlier papers have acknowledged the collaborators and

funders of  this long-term study. The 2004 enumeration,

funded by the Drummond Fund of Gonville & Caius

College, Cambridge, and by a Manaaki Tangata fellow-

ship from Landcare Research, New Zealand, was carried

out by the authors and Howard Beckford. N. Brokaw,

D. Burslem, D. Coomes, H. Griffiths, P. Grubb, D. Wardle

and an anonymous reviewer made valuable criticisms

of  the manuscript.


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Received 18 December 2005 

revision accepted 11 April 2006 

Handling Editor: Kyle Harms

Supplementary material

The following Supplementary material is available

online from

Appendix S1 The number of  stems of  each species in

1974, 1984, 1994 and 2004.

Appendix S2 The statistical tests comparing: mortality

(a), recruitment (b) and turnover (c) between forests in

the three decades..

Appendix S3 Mean diameter growth rates per decade for

stems in the four forests alive in 1974 and 2004.

Appendix S4 The difference between the absolute annual

stem diameter growth 1989–91 and that from 1974 to 84,

plotted against the stem diameter size in 1989.

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