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11. REVIEW OF LITERATURE
REVIEW OF LITERATURE
Tropical d y evergreen f o r m
Tropical dry evergreen forests are dishibuted on the eastern (Commandel) coast of India
(Panhasamthy and Sethi, 1997), extending inland about 50 km (Mani and Parthasarathy.
2005), northeastern Sri Lanka (Blasco and Legris, 1973: Perera, 1975; Dittus, 1985),
northeastern Thailand (Bunyavejchewin, 1999), southwest China (Hongmao at 0 1 . 2002)
and south coast of Jamaica (Loveless and Asprey. 1957. Kelly rt a/.. 1988) and Bahamas
(Sm~th and Vankat, 1992). The dry evergreen forest formerly covered 80% of the island's
land area of Sri Lanka, and often has been dcscnbed ar "old secondary cltmax" In
rccogn~t~on of past disturbancc (Dlttus, 1985)
On the Coromandel cast coast of India, thc true troptcal dry evergrcen forests
occur presently in the form of small patchcsifragmcnts of 'sacred gruves' or 'temple
forests', such as that resource extraction from the grove would bring them thc wrath of
the deity (Parthasarathy and Katihikeyan. 1997a) Thc land use systems in many sacred
gro\ci arc nour threatened (Chandrashckara and Sankar. 1998). and must be studied
s~multaneously to understand and minimize the ccoioglcal Impact of humans on forest
ecosystem$ (Williams-Llnera. 2002) Small forcst fragments arc reported to provide a
safc!) net for a rtgnlficant number of spcc~es and t h c ~ r genetlc dtverslty (Turner czr ul.,
1994). a breathing space for conservationists to plan strategies for preventing the loss of
me spec~es concerned (Turner and Corlett, 1996).
Dry evergreen forcst formations are some what rare, expressing a habitat where
the molstwe supply shows no relattvely effective seasonal fluctuations, but is fairly
cons~stently madequate - having regard to particular cllmatic conditions - for the most
luxuriant growth. The tern is chosen for want of a better to express the idea of
sclerophyllous evergreen formations in a form popularly acceptable (Beard. 1944).
The climatic conditions of the tropical dry evergreen forests are two distinct
seasons in a year: the long dry season and the short rainy season. April, May and June are
the warmest months. Heavy rainfall occurs from October and December due to the
northeast monsoon (Balasubramanian and Bole, 1992) According to Beard (19441, the
development of dry evergreen forests and woodland 1s due usually to strong winds andlor
excessively freely draining soll, the soil moisture being thus inadcquatc to meet the
evaporatrng ability of the air.
In the dry evergreen forests, soils are mostly ferrallltrc sandy loam, forming
"Cuddalore sandstone" of Miocene period, the fossils of whlch contain genera presently
exrstlng rn the regron and also those of dry deciduous and wet evergreen forests rmplylng
a humid climate (Meher-Homjr. 1974). M e n comparcd with troplcal wet forests, thcy
recer\c less annual rainfall (<I200 mm), less basal area, buttrcsscs arc rare, caulifloly is
uncommon, tree boles are mostly 8-12 m in hcrgh~. herbaceous vascular epiphytes are
very rare and large vertebrate dispersers absent. Table I summarizes studres conductcd in
varlous trop~cal dry evergreen forests of the world, with main objectrves focused in those
studres
Plther and Kellman (2002) also consider that even vcry small forest patches (less
than 1 ha in size) could play a role in the maintenance of regional diversity by
augmentrng reglonal populat~ons, providing habitat and food for plant and animal spec~es
(Guidon. 1996; Lyon and Honvich, 1996). In smaller fragments specles oAen become
hypcrdlsturbed, leading to progress~ve changes in floristic composition (Laurance. 1997)
T~bk I Summary of main objectives focused in various studies on tropical dry evergreen
forests of the world. R e f e ~ n c e sources are arranged in ascending chronology.
SI. No. Site Focal objectives Cootributor(8)
The study deals with the non-
littoral formations as they Loveless & I. South coast of Jama~ca
occur on the limestone hills of Asprey. 1957
the south coast.
Detailed study on ecology, Point Calimere & Blasco & Legris.
2. phys~ognomy and dynamism in Marakkanam, India 1973
the two sites of south India.
Assessment of damagc caused
by a cyclone ~ncludes
Polonnaruwa, defoliation, brcakagc of t u ~ g s . 3. Dittus. 1985
Nonheast Sri Lanka branches and trunks, tree falls
and post-cyclone tree
mortality.
Comparison bctwcen flonstlcs.
4. Round Hill, Jamalca structurc an cnv~ronmcnt along Kelly e t a / . , 1988
with a rainfall gradlent.
The vegetation of fifteen
stands data analysed by North Andros Island, S m ~ t h & Vankat,
5 . dominance-type classificat~on Bahamas 1992
and deterended
correspondcncc analys~s.
Fruiting phenology of fleshy Balasubramanian
6. Point Callmere. lndia fmlted plants In relat~on to & Bole, 1993
cllmate.
Marakkanam & analysis along w ~ t h soil 7. Visalakshi, 1994
PuthupeL India propenies and the extent of
human disturbance
St. No. Site Focal o b J r ~ i v n ConMbutor(s)
Seasonal dynamics in fine Marakkanam &
8. roots mass and fine root Visalakshi. 1995 Puthupet, India
production.
The study deals spatial and
temporal variation of stomata1 Sakaerat, Norrheasten
9. conductance of the emergent Pitman. 1996 Tha~land
diptemcarp Hol~eo,ferrea in
Thailand
Description of rare and
Oorani & Puthupct, cndem~c liana species Dern~ Balachandtan & 10
India ovalifolra. rcd~scovercd from Gastmans. 1997
Pondlcheny
Puthupct, lnd~a
Investigation of population Parthasarathy &
structure and dispersion of all Srthl 1997 - - .. . . ,
trcc and llana spccics.
Parthasarathy & Kuzhantha~kuppam & Quantitatlvc inventory of plant
12. Karthikcyan, Thlrurnanlkkuzh~, India b~ud~vcrs~ty in two l-ha plots.
1997a
Quantltatlve asscssment of Sakaerat. Northcasten Bunyavcjchcw~n.
13. srructurc and stand dvnamlcs
In two ]-ha permanent plots
Plant biod~versity and
Ooran~ & Olagapuram, population structure and the Ramanujam & .I4
lndla role of bellcf systems in thc~r Kadamban, 2001
conservation
Xishuangba~a, Conserving plant biodiversity Hongmao er a/ . , 15
Southwest Chlna through traditional beliefs. 2002
Quantitauve ~nventoly of l~ana Reddy & Coromandcl coast,
16. lnA.2
d~venity and distribution in Parthasarathy. ... "." four l-ha plots. 2003
SL No. Site Focal objectives Contributorfs)
Stand structure, composition Venkateswaran Coromandel coast,
17 and human disturbance in five & Panhasarathy. lndia
I-ha permanent plots. 2003
Investigation of seasonal Pragasan &
Kuzhanthaikuppam & patterns In fine liner 18. Parthasarathy.
Oorani, lnd~a product~on and standlng crop 2005
of Ilttcr.
Community-level fruit Swamynathan & Kuzhanthatkuppam &
19. product~on and d~spersal Parthasarathy. Oorani, lndia
modes 2005
B~(dlvers~ty assessment of Man] &
Pudukotta~ district, trecs in five Inland tropical dry 20 Panhasarathy,
lndta cvergrccn forests of peninsular 2005
lndta
Inbert~gat~on of tree populat~on Venkateswaran
21 Puthupet, India changes over a decade ( 1 992- & Parthasarathy.
2002) 2005
Rcproduct~i,~ tra~ts of plants In
relat~on lo pollinat~on systems Selwyn &
Coromandel coast. and disport- dispersal modes 22 Panhasarathy.
lnd~a and flowering and fiuttlng 2006
phcnology In rclat~on to
variour rcproductiw traits
and such forest fragments can prov~de a brcath~ng space In which conservation strategies
can be developed to asalst the species conccmed.
Ideally, the establishment of large forest tracts as conservation areas would avoid
the necessity for such resuscitation from tlny fragments. but in some cases the nccesslty
has M y m v e d (Turner er 01. 1995). Thus d l forest remnants may have tmpotiant
b~ologtul. economlc and soclal aspects ( L a w . 1997) and chew fngmentcd patches
wrll cwtnbute most of the plant spccres avlulable for ncolonrzatton. emphastzlng that
the conservatlon of all landscape habrcats 1s cnttcal for the marntcnance of dtverstty
(Wrlltams-L~ncra era1 1998) However. Cores1 fragmencatlon does not occur alonc, but
IS always asroctated wtth other human-tnduced threats to trees, such as loggtng. forest
burntng and hunttng of key vertebrate seed drspmcn w~thtn forest remnants (Tabarellt rr
ol , 2004) Hence, rcsourcc planners should not Ignore or dlmrnlsh the potenttal rolc of
very small forest fragments In conservatton lnrttatrves but rathcr should uttlrae them as
contnbutrng components In rcgronal plans (Prther and Kcllrnan. 2002)
Sacred grows
Sacred groves are patche\ of natural cltmax vegetatron and protected by rclrgtou\ bcllef
of local people (Parthasarathy and Karthlkcyan. 1997a. Khumbongmayum er a / . 2005).
and are regarded a. the trearurc house of rare speclo (Upadhaya cr 01. 2003) In fact. the
nature of reltgtous cults ar\ocratcd wlth the wcred groves suggcbts that rhcsc cults date
from the huntrng agc before man had settled down to rarse ltvcstock or ttll the land The
detttes generally Ire open to the sky, and are known In many cases to bc offended of a
shelter be erected over them They are always sttuated at a dtslance from any human
settlement, all of whtch potnr to therr ongrn In the nomad~c stage ofsocrety (Cadgtl and
Vartak, 2004) The size ofthe sacred groves ranges from clumps of a few trees to a few
hectares (Chandrakanth el 0 1 , 2004)
Traditionally. ycd groves ernbady a rich npenotre of forest prewvr~uon
practices d share c ~ e r i r t i c s wllh common propmy mouree systems
(Chandrakanth er 01.. 2004). Sncred grove 1s a treat of virgin forest. harbouring rich
blcdiversity and pmtated tmdltronally by the local cornmun~tics as a whole
(Khumbongmayum el 01.. 2005). and I I holds potential for preserving not only
biodiversity and ecological funct~ons, but also cultural dlvcrs~ty (Gadgil and Vartak.
1975: Ciadg~l and Chandran. 1992. Ramakrishnan rr 0 1 . 1998). In contrast, the rel~gious
bel~efs and ntuals central to sacred grove preservatton arc now fast crodlng, and
therefore, these treasure houses of blodlvcrslty cannot be protcctcd indcfinltcly only
through re l~g~ous bellcfs (Tiwan er a/.. 1998). Due to current scarcity of varlous
resources. the old taboos arc less effect~vc and some sacrcd groves havc bcen destroyed
(Chandrakanth rr ul.. 1990).
Culture-based conservation has been a long tradlt~on of Ihc local community
practice, plant, and anllnals are closely assoc~ated with many soc~al customs and
r c l ~ g ~ o u s ntuals of local people In the rcglon The sacred plants, sacred an~mals, sacrcd
forests and holy rnountalna are common phcnomena In the mountain alcas of the rcgcon.
wh~ch have played an Important role and can bc cfTccttvcly lncorporatcd Into lntdern
conservation (Yang er a / . 2004) Cultural dlvers~ty has a close rclat~onsh~p with
blodlvers~ty conscrvatlon has recelved lncreastng altention (Ciadg~l cr 01.. 1993;
Augustine and Adrianc. 1999)
In Ind~a, the b~odiversity and cultural value of sacred groves have bcen
particularly well documented. Several studies have been carried out in India to assess the
biodlvers~ty of the sacred groves located in Kerala (Chandrashekara and Sanknr, 1998).
Mahanshha (Gadgl and Vsrtak 2004), Gujur t (Reddy el a / . 2004), northeast India
(Mrshra el a l . 2004. J a m and Pandey, 2003. Khumbongmayum el a/ 2005) and
Coromandel coast of Tamtl Nadu ( P m h a u n t h y and SeLt. 1997, Panhawrathy and
Karrhlkcyan. 1997a. Venkateswaran and Parhasamthy. 2003. Ramsnujam and Cynl.
2003) have demonstrated the btologtcal value of these sacred groves OAcn. tn populated
areas, sacred srtcs conserve t k only vegetatton wtthout radtcal human alteratton
(Ramaknshnan. 1996)
At the global level sacred groves have been r c p n e d from Asla (Koagne. 1986.
Gadgtl and Chandran, 1992. Ttwm rr a1 1998. Hongmao er ul 2002. Yang el a1 2004.
Anderson el a1 2005) and Afnca (Campbell, 2004) Although wrnc \upyx,nlng cultures
have been weakened by modern ~nfluenccs, sacred groves arc frcqucntly more acceptable
to local people than externally ~ m p s c d consmatlon poltcre\ (Nt~amao-Baldu. 1994)
There 15 Increased ~nternat~onal ~nterest In relrgtously based rc$tnct!ons on land and forect
stand use Howcvcr, the extent to whtch $0-callcd sacred groves rcprc\cnt carltcr forcjt
ccosystcmc, and thc~r pss tb le role In btod~vers~ty Lonservatton are tntcrrelatcd and
complex Issues, and whtch are belteved to he In transrtton from a forested past
(Campbell. 2004)
Sacred grove forest s~tes throughout the world arc Important for the prcservatlon
of plant and anlmal specres useful to local people (Wadley and Colfer, 2004) For
example, sacred groves become refuges for plants, blrds, mammals, and other forcst-
dwelling animals (Basu, 2000. Chandran and Huges. 2000. Stnha, 1995). and people
depend on them for vmous products used in evcryday ltfe (Burnan, 2003.
Chandrashekm and Sankar. 1998) Sacred groves may thus serve both local resource
needs and interntioral conmation goals (Decher. 1997; Lebbie and Guries, 1995;
McWilliam 2001; Swamy et aL, 2003).
In addition to pnssure put on them from daily use, these forest patch habitats are
fragile. and changes in traditional cultural and economical values may threaten their
existence (Byers era/. . 2001). Mishra er al. (2004) found that the cutting of mature trees
for timber, collection of fuel-wwd and cattle grazing were malnly responsible for the
community organization and altering the botanicallflonstic composition in Swer sacred
grove. Mehalaya, northeast India. The increaslng demand for land and wood and growing
disrespect for traditional values are paralleled by increaslng erosion of the forest edge,
cven in sacred forests (Hawthorne, 1993).
Examination of the contribution of the sacred forests to biodlverslty conservation
offers perspect~ve on the sacred forests as a model for environmental protection (Camara,
1994). Thus the role of natural sacred sites, particularly sacred groves, is attracting
lncreaslng Interest in international organizations and conservation organ~zations such as
UNSECO, the WWF and has significant relevance for the lmplementatlon of anlcle 8j of
the Conservation of Biological Diversity whrch stresses more on thc use of trad~tlonal
wisdom and practices for conservation and sustainable use of b~ological diversity
(Chandrashekara and Sankar. 1998).
Biodiversity of tropical forests
The introduction of the term biological diversity with its shon form biodiversity is rather
new, which emerged some twenty years ago (Lovejoy, 1980a; b; Wilson and Peters,
1988; Reid and Miller, 1989; McNeely el a!.. 1990; Chauvet and Oliver. 1993), but the
origins of the concept go far back in time. Ideas regarding the linkages and nlationships
between organisms and their environment, both b~otic and abiotic, wen developed from
the eighteenth century onwards, as nahmmlists such as Danvln, Humboldt and Wallace
observed the patterns of distribution of species and vegetation types in their natural
environments, but it was not until the early pan of the twentieth century that ecology
developed formal 1001s for the measurement and modeling of these relationships and their
d~verslty Palmer (1995) stated that specles diversity appears to be the most straight
forward concept of the components of biodivers~ty that the other two components namely
the genetic diversity and community diversity.
Tree diversity inventories in the tropics have employed a w~de range of sampling
protocols that vary in uee size threshold cons~dercd for sampling, and the number, size
and shape of the plots. Tree size i.e. girth or d~ameter at breast he~ght (gbhldbh; at 1.3 m
from the ground level) has been considered as a crlterlon for mensuration, There werc
studies includ~ng the enumeration of indlv~dual trees as small as 2.5 cm dbh (Knight,
1975) though 4.5 cm dbh (Bunyavejchewin, 1999). 5 cm dbh (Pel~ss~er and Riera. 1993;
Valencia ef a/.. 1994; Johnston and Gillman, 1995. Upadhaya cr a/.. 2003; Small pf 01..
2004), and 10 cm gbh (Parthasamthy and Kanh~keyan, 1997a; Parthasarathy and Sethi,
1997, Vcnkateswaran and Parthasarathy, 20031, 30 cm gbh (Kadavul and Palthasarathy,
1999a; b; Ayyappan and Parthasarathy, 1999; Ch~ttibabu and Parthasarathy, 2000, Sagar
era/.. 2003; Nath ef 01.. 2005; Muthuramkumar et a/., 2006), 91 cm gbh (Poore, 1968; Ho
el a/., 1987), to 152.4 cm gbh threshold (Fox, 1967). Limits of I cm dbh are rarely used
(e.g. Bongers er 01.. 1988), but has been gain~ng a momentum in the last decades
(Hubbell and Foster, 1983; Condit, 1995). The often used limits are 10 cm dbh or 30 cm
gbh (see Campbell er al.. 1986; 1992; Gentry, 1988s; b; Lieberman et a/.. 1996, Phillips
and Gentry, 1994; Phillips, 1996). This size class normally useful to forestry and they
supposedly play a major role in forest structure and function~ng than the lower sizes
(Newbery el 01.. 1992). Some studies havc also included lianas in the inventory (e.g.
Gentry, 1988b; Lieberman er al., 1996, Makana er 01.. 1998; Paahasarethy, 1999). Tree
herght has also been wnsidered as a criterion for tree diversity inventory. For instance, all
trees of ? I m height were inventoried In the lowcr Rro Negro, Amazonia (Rodrigues,
1961, Prance, 1979).
Floristic inventories and studres of forest dynamics usually rely on sampling plots
(Dallmerer, 1992). The effects of plot 51zc (e.g. K~lburn, 1966: Greig-smith, 1983) and
the influence of plot shape (Condrt ei a/ . , 1996. Laurance er a/.. 1998) on the estimates of
plant dlven~ty have been assessed, at least the former in detail, while the latter less
extensively, especially m the toplcs. Plot-less methods have also been employed for trec
dlvers~ty invcntory. For example, Balslcb er a/. (19x7) establ~shcd two -]-ha plots, one
each in terra firme and Varzea forests at Anangu, Ecuador, and enumerated all trces 210
cm dbh, employrng point-centered quadrat method.
Most studies have followed the plot method, lnclud~ng square plots (e.g. 100 m x
100 m; Campbell el al., 1986; Gentry. 1990) to rectangular plots (e.g. 80 m x 125 m;
Prance, 1990). to long belt transects (e.g. 10 m x 1000 m, Boom, 1986). Plot-based
research occurs within a range of plot s~zes from 0.1 ha plots (e.g. Gentry, 1988a), to I ha
plots (e.g. Black er al., 1950; Uhl and Murphy, 1981), 50 ha plot (in BCI, Panama
[Hubbell and Foster, 19831 and up to 52 ha plot in the Lambrr National Park, Malaysia
(Condit el a/ . , 2000). One-hectare plots have been widely used in tropical forests. In the
recent years the methodological emphasis in the study of tmpical forests has shifted to
large-scale permanent forest plots. The rationale is to provide suficiently precise
estimates of diversity, density, dispersion pattern, mortality, recnritment, g o w h and net
rates of change in structure and populations (Hall et a/.. 1998). The 50 ha plot at BCI.
Panama, established in 1982. was the first of the 'mega-plot' (Hubbell and Foster, 1983).
Since then, the number of permanent plots has increased rapidly in various
tropical forests of the world (Manokardn er 01.. 1990: Sukumar et a / , 1992; Condit. 1995;
Aiba and Kitayama, 1999; Ayyappan and Paflhasarathy, 2001. Nebel rr a/.. 2001; Sagar
and S~ngh, 2003). Such several large-scale forest plots with an area of 50 ha or more
(Condit. 1995; Condit er a / , 1996). howevcr, been established for the purpose of
analyzing the spatial pattern of populations and estlmatlng demograph~c parameters using
large sample s~zes (Kohyama and Takada. 1998). Small scale diversity is sometimes
independent of large scale divers~ty, rcflecting strong biotic and ab~otic interact~ons that
limtt small scale diversity (Cornell. 1985: 1993; R~cklefs. 19R7; Cornell and Lawton.
1992). Because the heterogeneity of an area at a small scale may be masked at a large
scale, the key factors for species diversity at the small scale might not be apparent at the
large scale (Ma, 2005)
Table 2 summanzes l~terature on quantltatlve ecological inventory of trees w ~ t h ~ n
the sample plots of mostly > I ha in the forests of the world. Tree species and family,
richness, stem density, and basal areas varied considerably across the tropics. It all began
with the pioneering study of Davis and R~chards (1934). They inventoried all trees ?I0
cm dbh in five 1.5 ha plot of tropical evergreen forest of Moraballi Creek, Guyana. The
floristic inventory of trees in the tropics had increased substantially after its status was
reviewed by Prance (1977), espec~ally on Amazon forest compos~t~on and suucture.
Genhy (3988b) opined that the highest alpha d rvm~ty of trees In the world occurred m
upper Amazonla He recorded 275-283 tree species ha'' for trees 210 cm dbh at
Yanomono and M~shana near Iqu~tos, Peru. lnventones In upper Amazoman Eucador
(Balslev er a/, 1987, Kom~ng er 01, 1991, Valenc~a er al, 1994) corroborate G e n y ' s
oplnlon Valencla el 01, (1994) encountered a srnlungly htgh tree specles nchness of 473
specles ha for 25 cm dbh This Inventory formed the world's h~ghest record of tree
specles nchness on a hectare basrs for stems >5 cm dbh
Although, tree dlverslty Inventory was l~ t l a t ed long back In Afncan troplcal
forests (Richards, 1939), only a handful of literature (at 21 ha scale ~nventones) are
ava~lable to date from the forests (Table 2) Some of available stud~es w~thln the African
forests, wh~ch sampled less then 1 ha area, include Gartlan et a1 (1986). Newbery et a1
(1986a), Taylor et a1 (1996) and Geldenhuys (1998) Recently four plots of 10 ha each
were established In the two forest types (two each In monodomrnant and mlxed forests)
of the ltun forest In northeastern Democratrc Republrc of Congo (Makana et 0 1 , 1998)
They reported only the results of 3 ha analys~s from a 10 ha block In each forest type
(Table 2) In therr study, all lranas 12 cm dbh were also Included In the Inventory
Quant~tatlve ecolog~cal mcntoly of AsIan trop~cal forests was lnlt~ated in the
second part of the 20'~ century (e g Ashton, 1964, Nicholson, 1965, Wyatt-Sm~th, 1966
etc ) Several stud~es have been conducted m the trop~cal lowland forests of south-east
Asla (Table 2) Newbery er a / (1992) oprned that forest lnventones of t h ~ s reglon often
use many small plots, and Invanably group specles, especially the non-commerc~al ones,
approx~mately to famlly or genus level only, so that the data are of l~mrted ecologrcal use
Notably, the D~ptero~xrpacac bave recaved the most detalled and accurate attention In
all the enumnatlons Many other famllles of cans~derable ecolo@cal, but little
commerc~al Interest (Euphorblaceae) are poorly known Of the total e~ght 'mega-plots'
establ~shed and malntalned In the troprcal evergreen forests of the world, four are located
In Asla (Kochummen et a1 1990, Cond~t et a1 2000) by the lnltlatlve of the centre of
Troplcal Forest Science network, whlch Include two plots In Malays~a (Pasoh Forest
Reserve and Lamblr Nat~onal Park) and one plot each In Sn Lanka (Slnharaja B~oshere
Reserve) and Tha~land (Hunr Kha Kaeng)
Slmllar stud~es on quant~tat~ve ecological ~nventory of trees In the lnd~an troplcal
forests are limited (Table 2) It was lnltlated by Ra1(1981), who [mentoned all trees ?I0
cm dbh In four plots of 2 7, 2 7, 2 63 and I 09 hectares respcctlvely at Dev~mane,
Mallmane, Kodkan~ and Katleken areas of the Western Ghats Most stud~es In the Ind~an
evergreen forests have been conducted durlng the last decade of the 20Ih century
Contemporaneous stud~es have been conducted ~n the seml-evcrgreen (Kadavul and
Panhasarathy, 1999a, b), evergreen forests of Western Ghats (Parthasarathy and
Kalth~keyan, 1997b, Snnlvas and Parthasarathy, 2000, Ganesan, 2001 Ayyappan and
Parthasarathy, 1999, Muthuramkumar el a1 2006) and Eastern Ghats (Chltt~babu and
Parthasarathy. 2000), and In a dec~duous forest (Sukumar el a1 1992, Sagar et al 2003)
and evergreen and moist dec~duous forests of the Western Ghats (Bhat et a1 2000) and
In the dry evergreen forests on the Coromandel coast (Parthasarathy and Seth], 1997,
Palthasarathy and Karthikeyan, 1997a, Venkateswaran and Parthasarathy, 2003) Ghate
et a1 (1998) studled different vegetation types (from evergreen forest to scrub) of the
Western Ghats totallng a 75 ha area, along a latltudlnal belt from 8" 21" N
Incidentally, the exponentla1 nse and geograph~c diverslficat~on In studres of
tropical trees IS lkely to continue though the next daade, w~th more than 10% of the
world's described tropical tree specles now being mon~tond in large, permanent forest
plots distributed throughout the Neotroplcs, Afnca and Asla (Burslem el a1 . 2001)
Species-area and spoeies-individual relationships
Species-area and spec~es-~nd~v~dual curves have been central to commun~ty ecology for
decades (F~sher el 0 1 . 1943, Preston, 1948, 1962a. b, Mac Arthur and Wilson, 1967,
Cond~t et a1 1996) The observat~on that the specles number tends to Increase,
continuously and monoton~cally wlth area was first puhl~shed In the work of Watson
( I 835) and latter ~t was re~terated The specles-area curve has been c~ted as one of the few
'laws' of community ecology (Schoener, 1976, Gould, 1979, McGuinness, 1984a, b) In
the tweenth century the emphas~s sh~fted from obscrvlng the relatlonsh~p to cxprcsslng 11s
from mathemat~cally (Anhenlus, 1921, Preston, 1960, 1962a, b, Gleason, 1922, 1925)
Bunge and F~tzpamck (1993), Colwell and Coddrngton (1994) and Chazdon er a1 (1998)
provided a broad overvlew of statlstlcal approaches for estlmatlng species r~chness from
sampies
The ~ncrease In tree species number w~th forest area been attnbuted to ecological
processes and also to sampllng effects, whereby larger forest fragments contaln more
plots that sample more of the communrty (HIII cr a/ 1994) Loss of dlvers~ty can only be
pred~cted uslng species-area relat~onsh~ps at the appropnate scale and ~n the correct place,
as trajectories of spec~es accumulafion dlffer according to forest type and disturbance
h~story (H111 and C u m , 2001)
Most models of commun~ty structure b d on hab~tat partrt~onrng suggest that
there will be an asymptote In the specres-accumulatron cwve Spec~es-area curves have
been constructed for several forests, and the specles numbers continue to nse over several
hectares (Wh~tmore, 1990, Richards, 1996) Cond~t et a1 (1996) analysed the specles-
area curves ra~sed for three slze classes for trees > I cm, ?I0 cm, 230 cm dbh in the three
50 ha plots, establ~shed one each In Pasoh (of Malaysra). BC1 (Panama) and Mudumalai
(Indra) They observed that specres cont~nued to accumulate mall the three Inventones up
to and beyond 50 ha Th~s contradicts a widely held belref among the aoprcal ecolog~sts
that tree specles rrchness reaches an asymptote at 1-3 ha (Boom, 1986, Gently, 1988b.
Tuom~sto rr a l , 1995)
Spec~es-area curves are st111 used to determtne the capac~ty of small forest
remnants to support specles d~versity (Rebelo and S~egfnd. 1990. Cowlmg and Bond,
1991. G~tay eral . 1991, Lugo er a1 1993. P~mm, 1998)
Species diversig and abundance
Troprcal forest 1s one of the most spec~es-rich vegetation format~ons on earth Typ~cally,
hundreds of tree specres coexrst In a single hcctarc of tropical forest (A~ba el a l . 2004)
One of the key goals of ecology IS to expla~n the d~strrbut~on and abundance of specres
(Hane er a / , 1999, Kunrn er a / . 2000) D~versity of a community can be assessed by the
proponional specles abundance data e~ther by uslng statlst~cal samplmg theory (F~sher ei
a / . 1943. Preston, 1948, 1962a) or by a variety of nonparametnc measures (S~mpson,
Shannon etc) Due to the complex nature and lack of theoret~cal jusc~ficat~on for
stat~strcal sampl~ng theory, the nonparametrlc measures have gamed a great deal of
popularity In the recent past (Msgurran, 1988; Krebs, 1989) Tmplcal fomts are
stmchually complex plant communltles (Phlll~ps and Gentry. 1994, Cond~t ef a / . 1996)
One of the charactenstic features of these forest. IS the11 h~gh specles nchness (Ayyappan
and Panhasarathy, 1999)
Ecosystems dlvmity on a spatla1 and areal scale 1s subd~v~ded Into alpha, beta,
gamma and delta d~vers~ty (Mac Arthur, 1965. Wh~naker, 1972) In forest ecosystems.
alpha dlvers~ty operates w ~ t h ~ n forest stands Beta dlvers~ty refers to h e vanatlon
between forests stands Gamma and delta dlvers~ty operate on large scales Vanous
lnd~ces have also been formulated for dep~ct~ng spec~es dlvers~ty The most common of
these are S~mpson's heterogene~ty index and the Shannon Index (Swmdel er a / , 1984)
Less informat~on IS ava~lable for beta drvcrs~ty, whlch descnbes how specles
composltlon vanes &om one area to another (Du~vcnvoorden er a / , 2002) Condit ef a/
(2002) present a new analys~s of beta d~verslty In whlch they compare the specles
composltlon of forest plots that are located at d~stances of 10 ' to 10' km apart In the
Neotrop~cs of Panama (southern Mesoamerica) and m Ecuador and Peru (western
Amazon)
R~chness and dlverslty patterns on elevat~on grad~ents are, however, l~ttle
understood, and have only been documented rccently (Rahbek, 1995, 1997, Vetaas and
Grytnes, 2002, Wang et a / . 2002, Bhattara~ and Vetaas, 2003), call~ng Into doubt the
hypothes~s that specles nchncss and dlverslty decrease monoton~cally w~th increasing
elevat~on (Sanchez-Gonzalez and Lopez-Mata, 2005) Pausas and Aust~n (2001) state that
the ma~n factors deteim~n~ng specles r~chness patterns at the local level are resource
availability and responses to environmental variables that have a t i t physiological
impact on plant growth or on resource availability.
Most divmity studies, especially for large extents, considered only one or two
components of diversity, species richness within local communities (u-diversity; e.g. Kerr
and Packer, 1997), species richness within a region (ydiversity; e.g. Cume, 1991; Caley
and Schluter, 1997; Lobo el a/.. 2001), or similarity between communities @-diversity;
e.g. Whittaker. 1972; 1977; Schmida and Wilson. 1985; Cody. 1991; Blackbum and
Gaston, 1996). Hamson and Buma (1999) also found that intermediate environments had
higher levels of a and y-diversity.
If the relative abundance of species In a particular plant or animal group in a
given community is some how measured, there will be some common species, and some
rare species and many species of varying degree of rareness (May, 1975). The concept of
dominance, i.e., the idea that certaln species so pervade the ecosystem that they exen a
powerful control on the occurrence of other species, is one of the oldest concepts in
ecology. McNaughton and Wolf (1970) opined that dominance is an expression of
ecological inequalities aris~ng out of different exploitation strategies. Possible
mechanisms which determine and maintain the dominance of one canopy tree species in
lowland tropical forests have been reviewed, again with a focus on wet or very wet
forests on well drained soils (Han et 01.. 1989; Connell and Lowman, 1989; Han, 1990)
and tolerance of soil fertility (Gentry et al., 1988a), type of mycorrhizal assoc~ation
(Grime et al., 1987; Janos, 1987), escape from predators (Janzen, 1974; 1984), and
succession (Connell, 1978; Hart, 1990).
It is well documented that species diversity of natural communities is often
strongly related to productivity (Palmer, 1992; Abnuns, 1995; Weiher, 1999; Braakhekke
and Hmftman, 1999; Brown, 20011, while an altihldinal gradient basically corresponds to
a temperate-precipitation gradient basically corresponds (Kilayama, 1992; Pendry and
Proctor, 1996; Mam el 01.. 1988). Wang et a/. (2002) found that productivity is measured
in terms of precipitation and temperature along an altitudinal gradient, a un~model pattern
with respect to the relation between diversity and productivity in Qilianshan mountains,
Gansu, China.
Local species rlchness can be estimated by extrapolating species-area curves
(Sagar el a/., 2003). The species-area relationships arise partly from an increase in habitat
d~versity with increasing area sampled (Diamond, 1988). These relationsh~ps are
important in ecological study because they provide insight into commun~ty structure
(Leps and Stursa, 1989), and the mathematical expressions of the models are used for
predicting species richness at larger scales, and extinction rates caused by hab~tat
destruction (Pimm el al.. 1995). The species-area relationship 1s a fundamental
component of conservation biology, and is often used to assess the long-term effects of
habitat fragmentation on biod~versity (Palmer, 1990).
A consequence of high specles dlversity that characterizes tropical forests is the
low density of most tree species and the large expected distances between the conspecific
trees. This phenomenon was noted a long back (Wallace, 1978). Rarity is a general
characteristic of tropical plant communities (e.g. Poore, 1968: Paijmans, 1970:
Thorington et al., 1982; Hubbell and Foster, 1986; Ho et al., 1987; Bongers er 01.. 1988;
Valmcia et a/., 1994; Pitman et a/., 2001). Rare trees normally would tend to be clumped
to ensure reproductive sunxss. This might be due to the d t y of the micmsite needed by
the species (Hubbell. 1979; Hubbell and Foster, 1986). Pwre (1968) found that at least
some of the rare species w m determined by soil differences. One of the major sources of
complexity in the analysis of rarity is the "sample" or "scale problem" (Soule, 1986).
It is generally recognized the species richness is positively associated with species
abundance (Denslow, 1995: Condit er al., 1998; Hayek and Buzas. 1997; Preston. 1962a),
and area and environmental heterogeneity have strong effects on species diverslty (Waide
er a/., 1999: Huston, 1979; 1994; 1980; Hubbell, 1997; 1998; Rosennveig, 1995:
Whitmore, 1998). The species-richness-abundance relationship suggests that large
populations are less prone to extinction than small ones (Preston, 1962a). Based on the
relationship between abundance and diversity, habitats supporting large numbers of
individuals can suppon more populations and more species than habitats supporting small
numbers of individuals (Huang er 01.. 2003).
Extrapolation using different models for species-area relationships can yield
different values of species nchness for a given area (Palmer, 1990; Soberon and Llorente,
1993), lntegrative measures such as diversity are valuable ecological metrlcs to simpl~fy.
characterize, and compare the complexit) of species assemblages (Christensen and Peet.
1984. Magunan, 1988) and are important topics in forest management studies (Hunter,
1990).
Spatialpatterns
Plant populations exhibit three patterns of spatial distribution - regular or uniform,
clumped or aggregated, and random. The individuals of a species is sald to be random ~f
the position of each individual plant is independent of that of all the others; aggregated
populations, arc those where there is a tendency for individuals of the species to occur in
clumps, and in regular populations the plants are more evenly spaced than they were
distributed according to chance (Pielou, 1960).
Clumped distribution is very common in nature and two causes have been
suggested for this (Feller, 1943). Firstly the seeds may fall at random over an area but if
the habitat is not homogeneous, the proportion of germinating and thriving will vary fiom
slte to site so that density is high in some sites are low in others. Secondly the hab~tat
may be homogenous, but the individual plants may occur In family groups, owing to the
fact that they reproduce vegetatively or by seeds with a small radius of dispersal.
Uniform d~stribution is extremely rare. According to Grcig-Smith (1983) a regular
pattern would be expected, if the members of a population were so abundant that they
compete with each other for available space. Most available published works ~nd~cate a
prevalence of clumped and a paucity of regular distribution patterns of tree species in
tropical forests.
Poore (1968) mapped trees t90 cm dbh m a 26-ha (620 m x 420 m) plot of
Malaysian lowland rainforest and found that 6 of the 13 most abundant species were
randomly distributed while 7 were clumped. Uniform distribut~on was found in only one
species. Forman and Hahn (1980) demonstrated that the spatial patterns of 16 most
species (>I0 cm dbh) in 4-ha forest plot at St. John, United States of Virgin Islands. They
found three quarters (I2 species) showed clumped pattern of distribution, but only one
species had uniform distribution, while the remaining 3 species showed random pattern.
Thorington el al. (1982) found that over dispmed seeds and seedlings of trees are
commonly eaten by animals in large quantity, but it is not a dominant factor for
conspecific species. The highest percentage (41%) of trees were clumped among five 1-
ha plots in the tropical moist forest of BC1, Panama. They mapped trees >60 cm dbh in
five I-ha plots. Of the 63 species 26 species showed significant clumping. Clumping of
tree species significantly acts as seasonal use by the animals (especially howler
monkeys), its survival and reproduction.
Spatial patterns may be determined by habitat, alternative population recruitment
strategies and differential competitive ability of seedlings (Janzen. 1970). In add~tion to
hlgh juvenile mortality caused by density or distance responsive predators, pathogens
etc , as postulated by the escape hypothesis imply disproportionate success for seeds that
escape the vicinity of the parent (Howe and Smallwood, 1982). Swan (1988) reported a
high degree of clumping in tree fall gaps due to the pressure of regeneral~on patches.
Poorter el al. (1994) determined spatial dtstnbution of canopy gaps In three sites
(total 71-ha) in tropical moist forest of Tai National Park, Ivory Coast. The study
revealed a clustered distribution of gaps for two of the three sltes. The mean number of
gaps per hectare and the percentage area in gap phase are remarkably similar for the threc
sttcs. Both direct and indirect evidence showed that gaps may tnfluence that spatial
d~stribution of trees in gaps but could be the result of clustered character of gap
distribution of trees in gap dishibution on its own. Sukumar er a/. (1992) at Mudumalai
tropical deciduous forest found that the most of the species showed clumped pattern of
distribution, with the exception of Gmelina arborea, which was randomly dispersed.
Acconli to Hubbell and Foster (1983) most of the tm species appear to have
their own individualistic dispersion patterns in their SO-ha plot study in BCI, Panama,
with only weakly developed associated by habitat in some species, and most of the
species being unevenly distributed and many exhibiting clumped dispersion.
Amesto el al. (1986) compared the spatial patterns of tree species (210 cm dbh)
in 2 north temperate, 2 south temperate, I sub tropical and 3 tropical forests. An Index of
dispersion based on distances to nearest conspecifics was used to infer tree spatial
patterns. Clumped patterns were predominant In all tropical forests (75 to 100% of
species) and in one north temperate forest. In 2 temperate forests and I nonh temperate
one, random panerns were common (50 to 88%). The difference in the proponions of
random versus clumped spatial panerns seems to be related to different histories of
disturbance in the forests compared. Webb eta/ . (1967) Ashton (1972) and Austin e t a / .
( 1972) indicated that in the absence of major disturbance, soil and water conditions play
major roles in controll~ng species distribution.
Paijmans (1970) study in New Guinea showed that most species were more or
less randomly distributed although instances of patterned distribution were found in each
plot, particularly in plot I where Sloanea sp., Lithocarpus and Euodia sp. were clumped
In the western part, while having few or no individuals in the remainder of the plot.
According to him the factors governing spatial distnbution such as variation in
environmental history and chance may all be operating at once.
In forest subjected to frequent disturbance (relative to the life cycle of dominant
species) the original resource patchiness is obliterated or reduced leading to random
pattern in the disturbed areas. lnitially colonization may occur in clumps due to seed
dispnsal patterns and vegetative spread (Hibbs, 1973). Later with the thinning of clumps
(Kershaw, 1973; Grieg-Smith, 1983) most trees should be randomly scattered
(Williamson, 1975).
As the forest matures clumping becomes common in association with gap
dynamics (Putz and Milton, 1982; Bmkaw, 1982; Runkle, 1982), but this is again
interrupted by frequent disturbances. Canopy gaps formed due to nee falls result in
Increased light mound upheaval and nutrient release (Bormann and Likens, 1979; Putz,
1983). Thls leads to predominantly clumped tree patterns with selection for gap-
dependent life cycle (Gmbb, 1977).
Population structure
Poore (1968) studied the population structure of trees 230 cm gbh for lndivldual species
as well as in two families: D~pterocarpaceae and Burseraceae. He found expanding 'L'
shaped tree distribut~on. Dipeterocarpaceae were represented by many more large trees
than the remaining families and the number of individuals of dipterocarps decreased more
slowly with increasing girth.
To determine population structure of trecs in tropical moist forest of BC1, Knight
(1975) established seven dbh size classes. The total number of ~ndlviduals in all the
quadrats belonging to each size class was determined for each species and was divided by
the total number of tree indlvlduals in all size classes of all species in the stand, thus
g~ving relative density data for each species size class, to have an estimate of population
structure and was used for evaluating the succession status of each species. According to
Hubbell (1979). if a population is composed largely or exclusively of adults than either
the population is declining or else reproduction in the species is episodic in nature.
Singh el 01. (1984) classified the girth distribution of species in Silent Valley
tropical rainforest, Kerala, as expandlng, decreasing, youthful, intempted and accidental.
According to him species-rich forests have a low percentage of expanding population
structures. They stated that the high specles richness in Silent Valley is made possible in
part by the species combination varying from one girth class to another. Thus species are
in constant flux in space and time and this is in turn is possible when suitable habitats of
sufficient size are available to encompass all the stages of growth of all species. Thus,
thcre occurs as a mosaic pattern or cycl~c change in regeneration giving rise to variation
in the combination of dominants.
Rai and Proctor (1986) studled the allometric distribution, girth and basal area at 5
cm interval in lowland rainforcst, Karnataka. Thcrc was a decrease in number of trees
from lower size class to higher size class. In the case of basal area, two sites showed
bimodal distribution and another two showed opposlte trend to the girth distribution.
Campbell er 01. (1986) studied the distribution of trees by class intervals of dbh in
terra tlrme and Varzea forests, which showed a typical, reverse ' J ' shaped curve. Swaine
el a/. (1987) in their study at Kade, Ghana showed that the distribution of size classes is
rypical of natural forest regenerating from seed w~th high numbers in the smaller size
classes and a more or less logarithmic decline in number with increasing size suggesting
a stable size and age distribution.
Balslev er al. (1987) studied the distribution of trees in different size classes form
both moist flooded and unflmded forests. Trees exhibited a reverse J-shaped curve for
und i shhd forest The inflooded fomt had higher proprtion of individuals with 525 cm
dbh when compared to flood plain forest. Tree distribution by high class formed a
skewed bell-shaped type.
Diameter distributions are commonly used to assess the disturbance effect within
forests (Hett and Loucks, 1976; Denslow. 1995) and to detect trends in regeneration
patterns (Pooner er ab, 1994). It can he used to gauge forest vitality with respect to
stocking of different age or size classes (Rollet, 1993; K~yiapi, 1998), and compare
recruitment of different forests (Kigomo el al., 1990). Moreover, tree density distribut~on
across different diameter classes ind~cates how well the growing forest is utilizing site
resources. Such a distribution IS ecologically more informative when accompanied with
data on spatial distribution of individuals (Krebs, 1989). A few small-to-medium sized
trees per hectare may Imply that land IS not being fully util~zed by the tree crop (Hitimana
rf a/., 2004).
Sukumar et a/. (1992) found that the deciduous forest of Mudumalai had a more
or less expanding population structure with a deficiency of indiv~duals in the smallest
size class. Some species had few or no Individuals in the smallest size class (<5 cm dbh)
as compared to the largest area.
Geldenhuys and Murray (1993) analysed the population structure of selected
dominant species. Size-class d~stribution (both he~ght and diameter) had been used to
infer development trends of species and forests. The population structure of most canopy
species exhibits a negative exponential or inverse J-shaped cure for a mixed evergreen
forest in South Africa.
Johnston and Gillman (1995) worked out the population structure of m s in fow
I-ha plots in the lowland fomt, Kuupukari, Guyana. The number of species and
individuals versus girth class exhibits a reverse I-shaped histogram. In the tropical
evergreen forest of Uppangala (Pascal and Pelissier, 1996) and Mylodai area of
Courtallum reserve forest (Parthasarathy and Karthikeyan, 1997b) the stand exhibited an
expanding population structure.
Monodominanf forests
The most abundant species constitute no more than 10% o i the total number of
individuals In various forest communities studied (Oliveira-Filho er al., 1994; Felfili.
1995; Silva Junior, 1995). However, there exist exceptions, where a single species shows
50 to IM)% of dominance. Such monodominant forests have been reported from different
continents (Davis and Richards, 1934).
Dominance is likely to result from the fortuitous and massive establishment of a
single, highly vagile, npidly growlng species (Hatl, 1990). Single-dominant forests differ
in structure from mixed forests and, as expected, the11 stratification is much more distinct
(R~chards, 1996). Monodominant forests actually contain many other specics, and are
found In immediate juxtaposition to d~verse forests. The accompanying species typically
arc also present in the adjacent higher-diversity stands. Apparently, dominance is typical
of the principle species throughout their ranges e.g. Gilbertiodendron deu'evrei in central
Africa (Hm et al.. 1989); Cvnomerra alexandrr in East Africa (Eggeling, 1947), hut data
on their autecology are especially scarce (Martijena and Bullock, 1994).
Read et a/. (1995) suggested that the mono-dominance may be more directly to
the disturbance regime rather than to a difference in site quality or soil variables in
Nothofagus dominated forests of New Caledonia. However, in some monodominant
forests, canopy disturbance is not restricted to one generation, but rather the dominant
species successfully recruits and replaces itself under its own canopy (Torti er of.. 2001).
Another interpretation of monodominance is successional ( C o ~ e l l , 1978; Hart, l990), as
a result of sequential replacement by one species which is the more resistant to stress, or
the best competitor in the pool of shade-tolerant species. The development of such
dominant populations, that characterist~cally have slow-growth and poor seed dispersal.
usually is impeded or dlverted by disturbance (Martoena and Bullock, 1994). Other
tropical monodominants have been explained as a sere in forest succession. These are
early successional species that are not able to reproduce underneath their own canopy;
hence, dominance persists for only one generation (Connell and Lowman, 1989; Hart.
1990; Read er a / . 1995). The dynamic equilibrium hypothes~s suggests that
monodominant forest could be either early success~onal (Huston, 1994) or, as speculated
for Guyana (Hammond and Brown, 19951, late success~onal forests (Han el a/., 1989,
Huston, 1994).
Moreover, shade-tolerant species that can establish species In the understory, but
need a canopy gap at some stage of their ontogeny, would also be a disadvantage (Clark
and Clark, 1992). Connell and Lowman (1989) suggested that the ectomycorrhizal
association may confer sufficient advantage to species establishing under exposed
conditions, e.g. following a large disturbance, that there may be a causal effect of the
rnycorrhizal type (ectomycorrhizal vs. endornycorrhizal) on the capacity of a species to
form monodominant canopies. Any biotic factor that promtes the interspecific
competitive advantage of the dominant trec species, or improves its ability to withstand
stresses of conspecific origin such as shade or a buildup of leachates, would promote
monodominance (Hart, 1990). It has been suggested that bopical nee species that form
species-specific ectomycorrhizal associations may have an enhanced ability to form
monodominant stands that will be self-replacing and able to invade neighboring mixed
forests (Connell and Lowman, 1989).
The abundance of saplings suggests the principal species is persistent and capable
of maintaing its dominance in subsequent generations, hut quantltatlbe data are few and
most of the descriptions regarding smaller size classes or regeneration processes have
been based on anecdotal evidence (Manijena and Bullock. 1994) For Gilhertiodendron
dewevrei. seed and seedling mortality is actually higher than for a common species in
adjacent m~xed forest; apparently lower mortality in larger size classes i~ more important
In determin~ng domlnance (Hart er ai., 1989) In the case of Dwohalanops aromatico in
Malaysia, more frequent reproduction and greater persistence of seedlings than In other
species may maintain its dominance (Whitmore, 1975).
Janzen (1974) postulated that a specles could anain domlnance if ~t experiences
greater recruitment by being a most fru~ter, thereby satiating seed predators. Mast fruiting
also may saturate all available germination sites by overwhelming other non-masting
species. A result of mast fru~ting is often a clumped distribution of individuals, whtch
would be a strong selective pressure for enhanced foliage defense agalnst specralized
herbivores and pathogens (Janzen, 1974). Hence, if the dominant species is endowed with
superior defences against local natwal enemies and therefore experiences low levels of
damage, it would gain sn additional gmwth andlor swvival advantage over other species.
Hart (1995) investigated the effect of mast fruiting on seed survival and canopy
dominance within the monodominant stands of Gilbertiodendron dewevrei in lturi forest
of the Democratic Republic of the Congo and found higher canopy dominance was not
assoc~ated with greater seed survival during one masting episode (Han, 1995).
Gross el a/. (2000) surveyed the rate of damage on young expanding leaves of
seedlings and saplings belonging to eight species within both monodomlnant
G~lbertiodendron dewevrei forests and adjacent mixed-species forests in eastern Congo.
Thew results showed that escape from herbivore and pathogen damage is not a
mechanism by which G. dewevrei achieves dominance, as it suffered the highest damage
levcl of any species surveyed. Similarly, other sub-dominant common species also
sufired high rates of damage. These results are discussed in relat~on to the phenolic.
fiber, and nitrogen content of leaves, and In the context of cuITent theories pertaining to
plant-herbivore interactions.
Forest dynamics: tree population changes over time and space
Understanding forest dynamics is fundamental to several aspects of rainforest ecology,
Including successful management of tropical forest for human uses. Due to the lack of
annual rings in many tropical tree species, the study oitropical forest dynamics has relled
heavily on the use of permanent plot data (Bunyavejchewln, 1999). One of the primary
purposes of permanent plots is to monitor forest diversity and processes otner time. As to
species diversity within a community, the permanent plot method is usually the most
accurav method to study its dynamics (Zang el al. , 2005).
Potentially, permanent tree plot could be employed as psn of an early-warning
system to detect the possible effects of environmental change on forests (Risser el a/..
1993). Repeated censuses of permanent forest plots can be used to describe the dynamics
of tree populations, i.e, to determine recruitment rate and mortality Wohyama and
Takada, 1998). Since the mid-20" century, a substantial body of data has been gathered
on the rate of tree mortality and recruitment ("turnover") in humid tropical forests
(Phillips and Gentry, 1994). Long-term studies are needed to determine whether changes
that actually occur over time within individual sites conform to successional patterns
described in chronosequence studies (Sheil. 1999). as well as to elucidate the processes
associated with these patterns (Capers er 01.. 2005).
Brokaw (1985a), Martinez-Ramos (1985) and Denslou (1987) emphasized the
importance of canopy openings (gaps) in the structural and compositional dynamics of
many troplcal forests. Research on troplcal forest dynamics has focused an characteristics
of gaps, including definition on boundary, frequency of creation, size of canopy
openings, species dependence, internal heterogeneity, physiologic requirements, species
packing and equilibrium or non-equilibrium status, among others (e.g. Hartshom, 1978:
1980; Denslow, 1980; Brokaw, 1982; 1985b; Orians, 1982; Augspurger, 1984; Bazzaz,
1984, Hubbell and Foster, 1983; 1987; Brandani er al. . 1988).
Tree mortality may be higher at the forest edge because of lack of protection
against strong winds or heavy storms (Williams-Linera, 1990a; b). This may result in an
Increase in the frequency of tree fall gaps in forest remnants (Lovejoy et al. . 1986; Kapos
el ui.. 1997), Interest in tree mortality and forest dynamics has increased recently
becaus forest dynamics is thought to be involved in determining tree species diversi~
(Phillips el aL, 1994), and also thought to be related to global climate change, in
particular (Phillips and Genhy, 1994).
Variation in forest dynamics can already be found on a small spatial scale and be
related to the differences in soil richness (e.g. van Schaik and Minmanto, 1985). Burslem
and Whitmore (1999) reported an annual mortality rate of 1.14% to 2.32% and
recruitment rate of 1.12% to 7.70% in the evergreen forest of Kolombangara Soloman
Islands. On a global scale, Phillips er a1 (1994) correlated a whole range of variables
(e.g, soil quality, rainfall seasonality) to forest turnover and found that forest tumover
explained most of the variation in tree species. In general, forests with higher rates of tree
tumover have a higher number of tree species. Table 3 summarizes the results of tree
dynamics monitored in various tropical forests of the world.
Forest degradation, human impacts and conservation
Tropical forests throughout the world are disappearing or deteriorating to a very rapid
pact. Among the most significant causes 1s the change in land use by the conversion of
land use to agriculture, pasture, or urbanization. Selective logg~ng and natural factors also
account for a large portion of nop~cal forest loss (Alvarado and Sanberg. 2001).
Tropical deforestation is a major concern on several fronts. It is significant to
global climate warming and regional climate change (Houghton el al., 2000); global
losses in biotic diverslty and net primary product~vity (DeFries eta/ . , 1999; Vitosuek el
a/., 1997); local-to-regional land deforestation (Barrow, 1991); and threats to ecosystem
services and other variable functions (Daily et al., 2000; Kremen er al.. 2000).
Singh el a/. (1984) described the vegetation of Silent Valley. Kerala and stressed
the need for conservation by making a plea to consider as it a Biosphere Reserve. Hubbell
and Foster (1986) studied the implications for tropical tree conservation in BCI, Panama.
Salwasser (1990) enumerated the major factors affecting biological diversity that include,
pollution, fragmentation of habitats, overuse of resources, conservation of wild areas to
agriculture and indushy and other human uses. Denich (1991) observed that deforestation
causes secondary succession of tropical forests. To preserve and sustain worldwide
tropical forest ecosystems, it has become critical to find remedial solutions to ameliorate
the deforestation and deteriorat~on of these ecosystems (Alvarado and Sanberg, 2001).
Yet, species extinctions are an inadequate measure of biodiversity loss and do not
prowde information about changes in the capacity of particular species to contribute to
the functioning of ecosystems (Luck er al.. 2003). Hubbell and Foster (1992) cmphaslzed
that the basic and applied ecological research have a vital and cost effecthe rolc to play
In trop~cal forest conservation and management of natural tropical forest is not possible
w~thout a better hol~stic understanding of low such forests actually work ecolog~cally and
~ntcract with humans. Thc increase ~n the proportion of declining species with increase in
disturbance intensity ind~cated that local anthropogenic pressure is responsible for the
dcplet~on (Sagar and Singh. 2003).
Gentry (1992) studied two major patterns of diversity and ~mportant conservation
plann~ng - the distribution of diversity forests occur in al three continents and are
concentrated in lowland but with greatest areas with high and evenly distributed iainfall,
in South American forests. Tree and liana species richness is greatest In upper Amazon,
non-tree species richness greatest in north Andean foothills and south Central America,
underlying conservation need of these areas.
Hobbs and H w m e k e (1992) stated that disturbance is an important component of
man ecosystems. They revived the effects of disturbances such as fire, grazing, soil
disturbance, trampling, fragmentation, and nuwient addition, on plant species diversity
and invasion could change the native species and enhance the invasion of other natural
communities. Costa el a!. (2002) found that selective logging causes several changes in
the structure and functioning of forests, which are due to the opening of variously sized
gaps and consequent changes in microclimatic conditions.
Ledig (1992) stated that humans have converted forcst for urban and agricultural
uses, which cause degraded environment with atmospheric and so11 pollutants.
Dcforcstation causes a vast scale of forest areas to reduce diversity by direct elimination
of both locally and adapted populations, which will bring atmospheric pollution and
global warming as a major threat in near future, especially because of the fragmentation
forest arcas. Yadav el al. (2003) observed that shifting cultivation practiced in Tirpura
hill slopes cause destructton of forest wealth and accelerate land degradation. The u-
dnersity and its components decreased with increasing dlsturbancc Intensity, reflecting
cnhanced pressure with tncreasing disturbance (Sagar and Singh, 2005).
Valkenburg and Ketner (1994) studied the floristic changes that occur following
human disturbance in mid-montane forest in Papua New Guinea. Of the two forests,
disturbed forest Mt. Kaindi and undisturbed forest MI. Mission, Norholagus pule] was
locally dominant in Mt. Kaindi. Change of floristic composition was observed between
18~0-2000 m. The survival of seedlings and saplings of Nothofigus increased by light
penetration. In addition natural events such as landslides, drought, tire and storms and
also human activities are overruling the natural disturbances which cause the
disappearance of species such as Nothofagus and Costanopsis due to lack of seed.
In forest areas traditional agriculture did not appreciably erode divers~ty where the
population density remained low, however overall, in many trop~cal forests pressures like
shifting cultivat~on, gathering forest products, monoculture etc, diminish the forest
biodiversity. Land-use may alter the habitat spat~al pattern, increasing the distances
among habitat patches (Hansen et ol., 2001). An important consequence of this habitat
fragmentation is a reduction in habitat connectivity, which could constrain the abil~ty of
many species to move across the landscape In response to cllmate change (Primack and
Miao. 1992).
Loreau (2000) revealed that biod~versity and ecosystem funct~oning has grown
from concerns about the potential ecological consequences of the present and future loss
of b~odiversity caused by thc increased impact of human activities on natural and
managed ecosystems. Foody and Cutler (2003) indicatcd that remotely sensed data may
be used as a source of information on biodiversity at the landscape scale that may be used
to ~nform conservation science and management.
The information on tropical plant diversity 1s needed because of its potential
usefulness and implication for conservation and management across the protected area
(Kadavul and Parthasarathy, 1999a). Further, an understanding of forest processes is also
fundamental to the management of natural and disturbed vegetation (Congdon and
Herbohn, I 993) palticularly in tropical forests.