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.