Seed dispersal in changing landscapes

Biological Conservation 146 (2012) 1–13

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Biological Conservation

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Review

Seed dispersal in changing landscapes

Kim R. McConkey a,⇑, Soumya Prasad b, Richard T. Corlett c, Ahimsa Campos-Arceiz d, Jedediah F. Brodie e,Haldre Rogers f, Luis Santamaria g

a A.V. Rama Rao Research Foundation, 7-102/54 Sai Enclave, Habshiguda, Hyderabad 500 007, Indiab Centre for Ecological Sciences, Indian Institute of Science, Bangalore 560 012, Indiac Department of Biological Sciences, National University of Singapore, 14 Science Drive 4, Singapore 117432, Singapored School of Geography, University of Nottingham Malaysia Campus, Semenyih 43500, Selangor, Malaysiae Wildlife Biology Program, University of Montana, Missoula, MT 59812, USAf Department of Biology, University of Washington, Seattle, WA 98195, USAg Mediterranean Institute for Advanced Studies – IMEDEA (CSIC-UIB), Miquel Marqués 21, E07190 Esporles, Mallorca, Balearic Islands, Spain

a r t i c l e i n f o

Article history:Received 17 May 2011Received in revised form 5 September 2011Accepted 10 September 2011Available online 27 December 2011

Keywords:Biological invasionsClimate changeFragmentationHuntingOverharvestingSeed dispersal

0006-3207/$ - see front matter � 2011 Elsevier Ltd. Adoi:10.1016/j.biocon.2011.09.018

⇑ Corresponding author. Tel.: +91 40 27117175; faxE-mail addresses: [email protected] (K.

(A. Campos-Arceiz), [email protected] (J.F. B

a b s t r a c t

A growing understanding of the ecology of seed dispersal has so far had little influence on conservationpractice, while the needs of conservation practice have had little influence on seed dispersal research. Yetseed dispersal interacts decisively with the major drivers of biodiversity change in the 21st century:habitat fragmentation, overharvesting, biological invasions, and climate change. We synthesize currentknowledge of the effects these drivers have on seed dispersal to identify research gaps and to showhow this information can be used to improve conservation management. The drivers, either individually,or in combination, have changed the quantity, species composition, and spatial pattern of dispersed seedsin the majority of ecosystems worldwide, with inevitable consequences for species survival in a rapidlychanging world. The natural history of seed dispersal is now well-understood in a range of landscapesworldwide. Only a few generalizations that have emerged are directly applicable to conservation man-agement, however, because they are frequently confounded by site-specific and species-specific varia-tion. Potentially synergistic interactions between disturbances are likely to exacerbate the negativeimpacts, but these are rarely investigated. We recommend that the conservation status of functionallyunique dispersers be revised and that the conservation target for key seed dispersers should be a popu-lation size that maintains their ecological function, rather than merely the minimum viable population.Based on our analysis of conservation needs, seed dispersal research should be carried out at larger spa-tial scales in heterogenous landscapes, examining the simultaneous impacts of multiple drivers on com-munity-wide seed dispersal networks.

� 2011 Elsevier Ltd. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22. Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23. Disturbances to seed dispersal processes: identifying gaps in our knowledge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

3.1. Habitat fragmentation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2. Overharvesting . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

3.2.1. Frugivores . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43.2.2. Plants. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

3.3. Biological invasions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53.4. Climate change . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63.5. Interactions between drivers of biodiversity change . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

4. Incorporating seed dispersal knowledge into conservation management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7
4.1. Habitat fragmentation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74.2. Overharvesting . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8
ll rights reserved.

: +91 40 27179149.R. McConkey), prasadsoumya@ gmail.com (S. Prasad), [email protected] (R.T. Corlett), [email protected]), [email protected] (H. Rogers), [email protected] (L. Santamaria).

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2 K.R. McConkey et al. / Biological Conservation 146 (2012) 1–13

4.2.1. Frugivores . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84.2.2. Plants. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

4.3. Biological invasions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84.4. Facilitating adaptation to climate change . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

5. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9Appendix A. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

1. Introduction

Linking ecological research and conservation practice is a majorchallenge in conservation biology (Knight et al., 2008). Ecologicalresearchers usually look in detail at only a few components or pro-cesses within a system while conservation practitioners makedecisions that affect entire ecosystems (Gardner et al., 2009). Un-less the lessons from detailed academic studies are synthesizedin a useable form, practitioners will continue to base their deci-sions on an understanding of a subset of the ecosystem. The ecol-ogy of seed dispersal is an excellent example of this failure ofcommunication, with a growing understanding of the processes in-volved having so far had little influence on conservation practice.Although seed dispersal research is increasingly being conductedat larger spatio-temporal scales, it has not been synthesized in auseable form for conservation practice and many conservation-rel-evant gaps exist.

Seed dispersal interacts with all the major drivers of biodiver-sity change in the 21st century: habitat fragmentation, overhar-vesting, biological invasions, and climate change (Pereira et al.,2010) (Fig. 1). In fragmented landscapes, seed dispersal has a majorinfluence on both plant species persistence (Farwig et al., 2006;Herrera and Garcia, 2010; Sansevero et al., 2011) and vegetationrecovery when disturbance is reduced (Howe and Miriti, 2004).Hunting has an obvious impact on seed dispersal by large verte-brates (Terborgh et al., 2008; Vanthomme et al., 2010) and over-harvesting of plants may also disrupt dispersal mutualisms(Ticktin, 2004). The patterns of spread of invasive plants are largelydetermined by their dispersal vectors (Buckley et al., 2006; Gosperet al., 2005). Finally, anthropogenic climate change adds to theimportance of seed dispersal, since most plant populations onEarth will need to move long distances over the next 50–100 yearsif they are to keep pace with moving climates (Chen et al., 2011).However, large-bodied animals capable of dispersing seeds to dis-tances required for plants to maintain this pace, are often threa-tened by hunting or habitat loss (Corlett, 2009). Failure to keeppace with changing climates may lead to major biodiversity de-clines and a significant reduction in carbon fixation.

There is evidence for a growing, global, seed dispersal crisis,which has so far been masked, by the long life-span of perennialplants. Plant populations that are neither being dispersed, norregenerating in situ, may persist for decades in an apparentlyhealthy state (Guimarães et al., 2008). These ‘living dead’ are, atleast for now, still living, so their rescue is low down on the long listof conservation priorities. At the same time, the role of ‘assistedmigration’—artificial dispersal—as a solution for expected futureproblems is being debated in the conservation literature withoutany clear understanding of the potential for unassisted migration(Hoegh-Guldberg et al., 2010; Sax et al., 2009; Vitt et al., 2010). Itis apparent that improved communication between seed dispersalresearchers and conservation practitioners could benefit both sides.

Our aim here is to synthesize current understanding of theinteractions between seed dispersal and the major drivers of global

change in order to identify key gaps that require further researchand to provide useful guidance to conservation practitioners. Westart by reviewing existing knowledge and identifying knowledgegaps, and then follow this by suggestions for how our currentunderstanding of seed dispersal processes can be incorporated intoconservation planning and management.

2. Methods

The process of seed dispersal (i.e., the movement of seeds awayfrom parent plants), links the reproductive stage of adult plants(pollination, fruit set) with processes regulating the establishmentof their offspring, including post-dispersal seed survival, seedlingestablishment, and sapling growth (Wang and Smith, 2002). Ourreview encompasses recent (mostly published post-2000) researchaddressing the impacts of four major drivers of global change (hab-itat fragmentation, overharvesting, biological invasions, climatechange) on the seed dispersal process itself, i.e., from fruit removalto seed deposition. Impacts on upstream (flowering, pollination,seed set, etc.) and downstream (seed predation, germination,etc.) processes are important, but outside the scope of this review.Within the scope, we focus largely on findings of direct relevanceto conservation management and on conservation-relevant gapsin existing knowledge. The review is targeted at both conservationpractitioners and seed dispersal researchers.

The aim of this review was to achieve a broad review of theproblem. This breadth means that each part is covered in a concisemanner, highlighting important insights. The impact of each driveron seed dispersal was reviewed thoroughly to identify the maintrends that have been established and the gaps in our knowledge.In the final review, we present recent references that address thevarying results and opinions. Where there was contradictory evi-dence for a particular effect, we include each scenario that has suf-ficient support. We used these published results to identifygeneralizations that may have some applicability in conservationmanagement. Our suggestions for improved conservation manage-ment include novel ideas that are apparent when maintenance ofseed processes is considered, restating the objectives of some lead-ing (but unfortunately rare) work directly using seed dispersalknowledge in conservation management, and rethinking someclassic conservation techniques.

3. Disturbances to seed dispersal processes: identifying gaps inour knowledge

3.1. Habitat fragmentation

In human-dominated landscapes, fragments of the original veg-etation cover often remain, with varying size, quality and connec-tivity (Lindenmayer et al., 2008). Seed dispersal is highly variablewithin and between these habitat fragments and is dependent onfactors that determine plant and frugivore persistence within

Fig. 1. Effects of five drivers of global biodiversity loss on dispersal vectors (wind and animals) and the subsequent impact these changes have on seed dispersal at populationand community levels. Links proven by empirical studies are indicated by a solid line, and unproven links by a hatched line.

K.R. McConkey et al. / Biological Conservation 146 (2012) 1–13 3

fragments, and the mobility of seeds between fragments (Brudviget al., 2009; Cramer et al., 2007; Moran et al., 2004). Under theinfluence of the island biogeography theory, seed dispersal re-search in fragmented landscapes has been founded on the idea offragments as islands of native vegetation within a sea of human-modified land-uses. Most studies have focused on patch character-istics, examining the influence that remnant area and/or isolationhave on various stages of the seed dispersal cycle. Declining habitatarea and increasing isolation have been found to reduce both thequantity of fruit removal and seed dispersal distances within andbetween fragments in most systems (Cordeiro et al., 2009; Kirikaet al., 2008; Lehouck et al., 2009; Rodriguez-Cabal et al., 2007;Uriarte et al., 2011), although exceptions exist (Farwig et al., 2006).

The nature of the matrix surrounding habitat fragments influ-ences seed movement between fragments through its effects on fru-givore abundance and behavior, and on wind dynamics (Damschenet al., 2008; Prevedello and Vieira, 2010). The abundance and diver-sity of native plant and frugivore species in fragments decreasesalong a coarse gradient of matrix types, from secondary forest, agro-forestry, exotic plantations, arable crops, to pasture (Gardner et al.,2009). However, few seed dispersal studies have explicitly consid-ered the matrix as a conduit for movement of seeds and frugivores(Breitbach et al., 2010; Gillies et al., 2011; Watling et al., 2011).Frugivores differ in their ability to move across a matrix and theirtolerance to habitat degradation (Prevedello and Vieira, 2010), andit is the multiple combinations of these factors that determine thefinal impact on seed dispersal. Matrix attributes alter frugivorebehavior and movement patterns (Breitbach et al., 2010; Gillies etal., 2011; Uriarte et al., 2011; Watling et al., 2011) with variable re-sults reported, even for species with similar diets and social organi-zations (Chapman et al., 2003) and across seasons (Naoe et al., 2011).These effects may result in altered seed dispersal patterns, althoughthis link has rarely been explored (Magrach et al., in press; Serio-Sil-va and Rico-Gray, 2002). How frugivores use the matrix will be influ-enced by bio-physical characteristics (Breitbach et al., 2010; Gillieset al., 2011), its disturbance history, changes in food quantity orquality, and altered interactions with other wildlife (e.g., predation,competition; Watling et al., 2011) and with people (e.g., hunting).

Seed movement across and into the matrix is also influenced bythe presence of structural elements such as linking corridors, livefences, paddock trees and remnant vegetation (Tewksbury et al.,2002; Levey et al., 2005; Pizo and Dos Santos, 2010). Most researchidentifying elements that promote seed movement across and intothe matrix has focused on interactions driven by fragment-tolerantdisperser species (Galindo-González et al., 2000; Levey et al., 2005),rather than forest-dependent frugivores (Gillies et al., 2011).

Wind circulation patterns are also altered in fragmented habi-tats (Damschen et al., 2008). Matrix permeability may be criticalfor wind-dispersed species in some ecosystems, but studies assess-ing landscape connectivity for wind-dispersed species have pro-duced variable results, and no studies have investigated theimpact matrix composition has on wind dynamics. Habitat corri-dors can act as wind breaks in open matrices and lead to increasedaccumulation of wind-dispersed seeds (Damschen et al., 2008).Similarly, canopy gaps can increase the deposition of seeds byaltering wind speeds, seed aerodynamics and by trapping seedsin eddies (Schupp et al., 1989). Conversely, pasture matrices couldenhance patch isolation, since interspersed paddock trees createobstacles to wind dispersal while failing to provide suitable habitatfor establishment (Rolim and Chiarello, 2004; Magrach et al., inpress).

Persistence of plant species in fragmented landscapes is influ-enced by how fruit and seed characteristics interact with disperserattributes. Larger-seeded wind- and animal-dispersed species of-ten experience reduced dispersal in habitat fragments comparedto smaller-fruited species in both tropical and temperate land-scapes (Cramer et al., 2007; McEuen and Curran, 2004). Thesefruit-size effects are probably driven by declines in frugivorescapable of handling large fruits in habitat fragments (Moran etal., 2004) and the shorter average dispersal distances of largerwind-dispersed fruits (Greene and Johnson, 1993). Plant speciesthat are dispersed by a few specialized dispersers (e.g., large fruitsize requiring large-gaped dispersers, species dispersed by deep-forest frugivores) experience reduced fruit removal (Cordeiro etal., 2009; Lehouck et al., 2009) and shortened dispersal within hab-itat fragments (McEuen and Curran, 2004). In contrast, plant


species that are dispersed by generalist frugivores or those thatstill have several functional dispersers are less affected by frag-mentation (Martínez-Garza and Howe, 2003; McEuen and Curran,2004; Farwig et al., 2006). Although generalizations about fruit sizeand dispersal capacity can be useful in the absence of other infor-mation (particularly for wind-dispersed plants; e.g. Thomson et al.,2011), there is no substitute for knowledge of local dispersal natu-ral history. We have a poor understanding of disperser redundancyeven in intact communities and there are exceptions to most gen-eralizations (Jordano et al., 2007; Moran et al., 2004).

Our limited understanding of seed dispersal across heteroge-neous fragmented landscapes reflects a major research gap. Few at-tempts have been made to address seed movement acrossfragmented landscapes at landscape scales (Brudvig et al., 2009;Damschen et al., 2008), to identify plant and fruit traits that charac-terize dispersal-limited species in fragmented landscapes, or toidentify the factors influencing the spatio-temporal variation in seeddispersal patterns, which can have important consequences for sub-sequent stages of the dispersal cycle (Herrera and Garcia, 2010).

3.2. Overharvesting

3.2.1. FrugivoresWildlife hunting, for subsistence or trade, is a major cause of de-

cline for many frugivores, particularly mammals and large birds(Corlett, 2007; Wright, 2003). The impacts are now most obviousin the tropics, where hunting pressures are rapidly increasing,but most temperate systems have previously lost species as aresult of hunting (Ellsworth and McComb, 2003; Martin andSteadman, 1999). Hunting in tropical forests is a pervasive threatin even the largest and most remote reserves, since it involves bothcommercial harvesting of high-value species as well as harvestingof common species for local use (Peres and Terborgh, 1995). Sincemany hunted vertebrates eat fruit, their decline alters disperserassemblages, with negative consequences for seed dispersal(Corlett, 2007; Redford, 1992; Wright, 2003; Vanthomme et al.,2010). These effects are intensified by the fact these animal com-munities are already truncated systems, following thousands ofyears of hunting pressure (Corlett, 2007; Wright, 2003). The eco-logical consequences of hunting have been investigated mostwidely in the Neotropics and island ecosystems (e.g., Chimeraand Drake, 2010; Kelly et al., 2010; Meehan et al., 2002; Peresand Palacios, 2007; Wright, 2003), with few studies in Asia orAfrica despite unprecedented hunting levels in some regions(Corlett, 2007; Fa and Brown, 2009; Vanthomme et al., 2010).

Selective removal of frugivores decreases overall fruit removalfrom fruiting trees and seed dispersal (Wang et al., 2007; Wrightet al., 2000), ultimately causing a decrease in the reproductive suc-cess of plants (Forget and Jansen, 2007) and reduced populationgrowth rates for animal-dispersed plant species (Brodie et al.,2009). Because our understanding of disperser redundancy is lim-ited, however, it is difficult to predict the impact the decline of asingle hunted vertebrate may have. The few studies addressingredundancy in hunted systems have generated inconsistent results,with the effect probably dependent on traits of the plant species(Sethi and Howe, 2009) and the diversity and evolutionary historyof local frugivore guilds. Network studies are useful for determiningthe robustness and redundancy within dispersal systems and iden-tifying the species that are most important for maintaining networkstructure (i.e., ‘hubs’ are species with a large number of interac-tions, ‘connectors’ bind together different subgroups in the net-work; Mello et al., 2011b). These studies suggest dispersalnetworks are robust to loss of some frugivores (Bascompte et al.,2003; Mello et al., 2011a), but more community-wide studies arerequired (Donatti et al., 2011) and the ecological and behavioral re-sponses of surviving species are largely unknown. The lost disperser

may be partially compensated by changes in the activities of non-hunted species as a result of reduced competition. Compensatorypopulation increases have been widely reported (Wright, 2003),but compensatory replacement of ecological roles has not yet beendocumented (and, in some cases, it has been shown not to takeplace, e.g. Staddon et al., 2010).

Hunters usually target the largest vertebrates and have done sofor millennia (Corlett, 2007; Crowley, 2010; Wright, 2003). Largevertebrates are often associated with the dispersal of large-fruitedspecies (Gautier-Hion et al., 1985; Kitamura et al., 2002) and aremore central components of dispersal networks (Donatti et al.,2011). Consequently, large-fruited, animal-dispersed, tree speciesdecline, often quite drastically, in overhunted forests (Peres andPalacios, 2007; Terborgh et al., 2008). This pattern is most visibleon islands, where historic and prehistoric hunting pressure has re-sulted in highly simplified frugivore assemblages, with often one(or no) animals capable of dispersing large seeds (Chimera andDrake, 2010; Hansen and Galetti, 2009; Meehan et al., 2002), andeven these species are frequently hunted (Walker, 2007).

Differences in generation time between vertebrates and foresttrees means that there will be a time lag between animal extirpa-tion and tree decline; forests that have been overhunted for decadesmay develop an ‘extinction debt’ for their vertebrate-dispersedtrees (Brodie et al., 2009). Such debts are probably nearly universalin tropical forests today as a result of the explosion of hunting pres-sure in recent decades. Defaunated forests may exhibit slow shiftsin species composition, as animal-dispersed trees are replaced bywind- or gravity-dispersed species (Brodie et al., 2009) and large-seeded trees by those with smaller seeds (Terborgh et al., 2008).Thus, even if at some later time we could control hunting, the rein-troduction or natural recolonization of extirpated frugivores maybe hindered by the reduction in the food trees upon which the ani-mals fed. Empirical studies testing these ideas are limited by thelong time scales involved, but systems that lost their dispersers ata more distant time period could provide adequate test cases. Forexample, some island systems that have suffered frugivore extinc-tions within the last few 100 years have plant species with fruit dis-playing characters that are not suitable for extant (or exotic)frugivores, and these plant species appear to be suffering recruit-ment failure (Chimera and Drake, 2010; Griffiths et al., 2010).

Even if animals are not completely extirpated, hunting-inducedreductions in frugivore density (‘half-empty forests’) can haveimportant effects on seed dispersal. We have little understandingof the functional relationship between frugivore density and seeddispersal, but the context-dependent nature of seed dispersaleffectiveness (Schupp et al., 2010) suggests that the relationshipmay often be non-linear (Redford and Feinsinger, 2001). Thisnon-linearity is due to at least three factors. First, animals may ex-hibit differences in foraging efficiency within populations, imply-ing that the loss of particular individuals may disproportionatelyaffect dispersal services (Redford and Feinsinger, 2001). Second,foraging behavior may be density-dependent. While modest de-clines in frugivore numbers might have minimal effects on seeddispersal, certain species may have thresholds beyond which ex-tant frugivore populations cease to provide effective disperser ser-vices. For example, flying foxes cease to be effective dispersers atlower population densities due to reduced intra-specific competi-tion (McConkey and Drake, 2006). Third, reduced competitive ef-fects following the loss of a species from a frugivore communitymay alter visitation patterns, fruit removal and dispersal distancesby other frugivore species due to altered inter-specific interactionsat fruiting trees (cf. McConkey and Drake, 2006).

3.2.2. PlantsLocal markets across the tropics display an ever-changing array

of wild-collected fruits (R. Corlett pers. obs.). Wild fruit harvests


are less diverse and more seasonal in the temperate zone, but thegrowing demand for ‘organic’ products may increase pressure onfavored species (Willer et al., 2008). A majority of the harvestedwild fruits are fleshy, so this harvest may compete with vertebratefrugivores (Ticktin, 2004). The impact of fruit and seed harvests onthe recruitment dynamics of resource plant populations has beenexamined at several sites (reviewed by Ticktin (2004)), but the im-pacts of fruit harvests on frugivores remains poorly researched.

Studies that have examined the demography of harvested plantpopulations using matrix population models suggest that thesepopulations might persist even under very high levels of fruitand seed harvest (Ticktin, 2004). However, these results might bea consequence of the low transition probability from seeds toadults in long-lived plant populations. Just as an ‘‘extinction debt’’exists between plant species and hunted animals, the impact ofplant harvesting may take many years to become apparent. Severalharvested plant populations have exhibited lowered seedling den-sities at intensively harvested sites, which has been attributed tolowered dispersal, fire, grazing or harvesting techniques (Ticktin,2004). However, seedling abundance may have little effect onoverall population dynamics for long-lived trees (Brodie et al.,2009; Pfister, 1998). Either detailed population models or verylong-term monitoring is required to determine whether reducedseedling abundance will translate into lower adult tree density.

Fruit and seed harvests reduce food supplies for frugivore pop-ulations. The limited available evidence suggests that frugivorepopulations can decline and frugivore communities may be alteredby fruit harvests (Galetti and Aleixo, 1998; Moegenberg and Levey,2003). Harvesting of palm fruits in Amazonia and the Atlantic for-ests of Brazil has reduced the abundance and species richness ofavian frugivores (Moegenberg and Levey, 2003) and frugivores al-tered their diets when palms were removed (Galetti and Aleixo,1998). The likelihood of resource redundancy within the diets offrugivores, however, or the conditions under which it may occur,is virtually unknown. Wide-ranging species that are able to trackresources are probably less susceptible to harvesting than thosewith a more restricted range (Rey, 1995), unless the harvested spe-cies fruits during periods of low community-level fruit availability(Kinnaird, 1992). In less-diverse habitats, fruit may be a limitingfactor for breeding or frugivore survival (Bronstein et al., 2007),and harvesting-induced reductions in fruit supplies could severelyimpact frugivore populations. In contrast, exotic plant speciesthat are planted in rural and urban landscapes for fruit harvests,can lead to increased fruit availability for native frugivores(McDonald-Madden et al., 2005) and altered fruit–frugivore inter-actions (Galetti et al., 2010; Nelson et al., 2000).

Harvesting of non-fruit plant products is also widespread andmay impact fruit supply and frugivores. Commercial logging canreduce fruit abundance either through direct targeting of fruit-producing species for extraction (Potts, 2011) or indirect removalof critical food resources such as strangler figs (Shanahan et al.,2001) that tend to grow on commercial timber trees of harvestablesize (Lambert and Marshall, 1991), and thus may significantly re-duce dietary diversity of important seed dispersers (Felton et al.,2010). Small-scale extraction of large trees reduces the numberof cavities available for nesting for some avian frugivore species(Cockle et al., 2010). At most sites, the dependence of frugivorespecies on harvested tree species is poorly understood.

Long-term studies are needed to address the impacts of plantharvests on the abundance and structure of resource populationsand frugivore communities. The consequences of harvests on thespatial patterns of resource populations remain largely unknown,despite having potentially critical impacts on frugivore move-ments. Spatial patterns of fruit availability determine which indi-vidual plants frugivores choose to feed from, as well as theirmovement patterns and consequent dispersal (Carlo and Morales,

2008). Changes in spatial patterns of fruit availability have the po-tential to influence such dynamics across all diet species, yet areentirely unstudied with respect to resource harvesting.

3.3. Biological invasions

Invasive species can alter native dispersal mutualistic networksdirectly, through the establishment of novel seed dispersal interac-tions with the native biota; or indirectly, by influencing the abun-dance, distribution or behavior of native species. From aconservation viewpoint, the outcome of these interactions is oftennegative (e.g. enhancing plant invasion) although positive out-comes may also occur (e.g. enhancing dispersal of native species)and are largely understudied (Schlaepfer et al., 2011). The impactof invasive species in pollination networks are increasingly docu-mented and indicate an overall erosion of connectivity amongstnative species (Aizen et al., 2008). Similar analyses would be ben-eficial in identifying the impact of invasive species in seed dis-persal networks (Gleditsch and Carlo, 2011).

Direct interactions have been well documented, reflecting theirincreasing frequency, and researchers are now attempting to pro-vide a more generalized framework highlighting the conditionsunder which alien species can become established in fruit–frugi-vore networks (Moran et al., 2004). It is important that suchframeworks be developed in more regions and in a greater varietyof habitats. Alien plant species often co-opt the services of nativedispersers, while alien frugivores make use of the resources pro-vided by native plant species. The establishment of these directinteractions between alien and native species can occur in fourways:

(a) Native frugivores dispersing alien plant species: this can be amajor determinant for the establishment and spread of analien plant (Buckley et al., 2006; Gosper et al., 2005; Aslan,2011; Gleditsch and Carlo, 2011). Frugivore movementsoften have a direct effect on the dispersal patterns and suc-cess of the alien plant species (Wilson et al., 2009), creatingconservation conflicts in invasive control (Gleditsch andCarlo, 2011), particularly when the native disperser is anendangered species (Campos-Arceiz and Blake, in press;Westcott et al., 2008). Invasiveness may be further facili-tated by increased investment by the plant in seed dispersalstructures regardless of the dispersal mode (wind or animaldispersed) (Murray and Phillips, 2010).

(b) Alien frugivores dispersing native plant species: the introduc-tion of alien dispersers can be beneficial if they complementthe dispersal service provided by native dispersers or restoredispersal services that were lost when a native disperserbecame extinct (Hansen et al., 2010). However, it may havenegative effects for the plant if alien dispersers consumefruits that would be otherwise available to more efficient,native dispersers (Castro et al., 2008).

(c) Alien frugivores dispersing alien plants: alien plants can bedispersed by previously-established alien dispersers, andsuch interactions may be one of the triggers of invasionmeltdown processes that facilitate the entrance of new alienspecies in a self-accelerating way (Simberloff and Von Holle,1999; Green et al., 2011). Positive feedbacks between aliendispersers and plants include the dispersal of alien plantsby introduced wild (Constible et al., 2005) or domestic(Campbell and Gibson, 2001) herbivores and alien frugivo-rous birds (Mandon-Dalger et al., 2004).

(d) Native and alien animals dispersing native and alien plants: theestablishment of novel seed dispersal mutualisms betweenalien and native biota often takes place in a multi-species con-text. These complex, yet probably very common, scenarios are


characterized by seed-dispersal networks that include bothnative and alien plants and dispersers (e.g. Davis et al.,2010; Lopez-Darias and Nogales, 2008).

Alien species can also alter seed dispersal networks indirectlyby influencing the abundance, distribution or behavior of speciesfrom the native dispersal network (Traveset and Richardson,2006). Indirect interactions between alien and native species re-main poorly understood for most animals and plants, with theexception of alien ants. Argentine ants (Linepithema humile) can se-verely reduce native ant communities. Seed dispersal by thesealien ants is often less effective, with decreases in seed removalrates and distances (Christian, 2001; Gómez and Oliveras, 2003)or they may preferentially disperse alien plant species (Rowlesand O’Dowd, 2009). Invasive ants can also alter the behavior of ver-tebrate dispersers: for example, endemic frugivores from islandshave been documented avoiding plants occupied by aggressiveinvasive ants (Davis et al., 2010; Hansen and Muller, 2009).

The introduction and expansion of alien vertebrates (such asrats, cats, snakes, and mustelids) that are predators of native frugi-vores has been regularly documented (particularly for oceanic is-lands, Courchamp et al., 2003), but reports of effects on thenative flora triggered by the loss of native dispersers are confinedmainly to anecdotal evidence and occasional observation (Travesetand Richardson, 2006), despite having potentially devastating con-sequences for seed dispersal mutualisms (Rodríguez-Pérez andTraveset, 2009). Alien plants in a disperser-limited habitat couldalso disrupt native mutualisms by ‘‘stealing’’ dispersers from na-tive plants, but they could also enhance the consumption of nativefruits situated in the immediate neighborhood (Gleditsch andCarlo, 2011; Neilan et al., 2006), or they could have a neutral effect(Gosper et al., 2006). Finally, alien animal species are frequentlyboth seed dispersers and seed predators, changing the dispersaldynamics of plant species across multiple recruitment stages(Wotton and Kelly, 2011).

3.4. Climate change

Predicting changes in plant distributions, community composi-tion, and ecosystem function in response to climate change is a keypriority in current ecological research, since these responses couldhave a large impact on the future of both biodiversity and carbonstorage (Purves and Pacala, 2008), but the relevance of dispersalbiology to these predictions has often been ignored. A meta-analysis of species range shifts associated with warming overrecent decades estimated that the geographical ranges of specieshad moved to higher latitudes at a median rate of 16.9 km perdecade, which matches expectations if species are trackingrecorded changes in temperature (Chen et al., 2011). No landplants were included in this study, but unless plants are less sensi-tive to warming, they would need to move a similar distance tokeep up with the movement of thermal isotherms. This medianis large in comparison with most known dispersal capabilities forplants (Vittoz and Engler, 2007; Aitken et al., 2008; Corlett,2009), suggesting that seed dispersal may limit the ability of plantsto respond to climate change, even without an expected accelera-tion of warming (IPCC, 2007). As a result, both species losses fromcommunities and immigration to communities may lag behindclimate change, leading to both extinction debts and ‘immigrationcredits’ (Jackson and Sax, 2010).

The data on which these estimates of latitudinal range shiftswere based is all from the temperate zone: in the tropics latitudi-nal temperature gradients are almost flat, so poleward migrationdoes not provide an escape route from rising temperatures. How-ever, altitudinal temperature gradients in the tropics are similarto those in the temperate zone. A global meta-analysis of altitudi-

nal range shifts found a median rate of 11.0 m increase in altitudeper decade (16.0 m for the seven plant studies included), in con-trast to the 50 m per decade needed to track rising temperatures(Chen et al., 2011). This lag in elevational response was unex-pected, given the relatively short distances involved, but maypartly reflect the topographic and microclimatic complexity ofmontane landscapes. In steep topography, these observed and ex-pected vertical rates translate into horizontal movements 2–5� asgreat, which are within the dispersal capabilities of many species,suggesting that altitudinal escape, where available, is more likelythan latitudinal escape.

Most studies predicting plant distribution under changing cli-mates use the ‘climate envelope’ approach, which couples informa-tion on current species distributions in relation to climate withfuture climate scenarios, ignoring the complexity of the processesinvolved when vegetation changes. In particular, they do not takeinto account the ability of plant species to migrate over the dis-tances required in the time available, or the impact of habitat frag-mentation on this ability (Corlett, 2009); they predict potentiallysuitable habitat for plant species rather than the potentiallycolonizable habitat (Engler and Guisan, 2009). Incorporating seeddispersal distances into predictive models will be particularly dif-ficult in species-rich vegetation where seed dispersal distancescan vary over four or five orders of magnitude between plant-vec-tor combinations (Vittoz and Engler, 2007; Corlett, 2009). In thecommonly used grid-based models, for example, a high resolution(fine-grained) model is needed to incorporate short-distance dis-persal (e.g. by ants or rodents) within habitat patches, while alow resolution (coarse-grained) model is needed to incorporatethe influence of long-distance dispersal (e.g. by elephants or fruitpigeons) in broad-scale, spatially heterogeneous landscapes.

Climate change may also affect the dispersal capabilities ofplants. Wind-dispersal will be impacted directly by changes inwind speeds, particularly the frequency of extreme winds (Soonset al., 2005; Nathan et al., 2011), and in some cases may be affectedby changes in plant morphology (Zhang et al., 2011). Plant re-sponses relevant to seed dispersal by animals would includechanges in fruit quantity and quality, and in the phenology of itsproduction. Outside the lowland tropics, warming may lift temper-ature constraints, leading initially to increased fruit production andearlier fruiting in spring-flowering species (Gordo and Sanz, 2010).In tropical lowland rainforests, annual tree growth increments de-cline significantly with small increases in temperature or decreasesin rainfall (Clark et al., 2010), but the sensitivity of plant reproduc-tive phenology is unknown. Relevant animal responses would in-clude range shifts, and changes in the phenology of breeding(and thus the seasonal pattern of dietary needs) and migration,all of which have been widely reported in birds (Jones andCresswell, 2010), while range shifts have been reported for mam-mals (Chen et al., 2011). Range shifts may be a problem when dis-persal agents track temperature changes, while plants lag behind.Phenological changes will be most important if they lead to a mis-match in the timing of fruit production and frugivore presence at alocation (Gordo and Sanz, 2010; Jones and Cresswell, 2010).

3.5. Interactions between drivers of biodiversity change

There is widespread concern that different drivers of global bio-diversity change will act synergistically to compound threats toecosystem functioning (Brook et al., 2008). However, few studieshave attempted to determine the nature (synergistic, additive orantagonistic) or quantify the strength of interactions between driv-ers (Schweiger et al., 2010). Given that multiple drivers frequentlythreaten fruit–frugivore interactions this is a priority area for fu-ture research. Invasive species, for example, can more easily invadedisturbed systems than those with an intact fauna and flora (King

Table 1Potential outcomes of two-way interactions between drivers of global biodiversity change on seed dispersal. Outcomes are noted as to whether they are additive, antagonistic orsynergistic. Research that supports the interaction effects are referenced: ⁄ indicates this has been observed for other ecological interactions and may be relevant to seed dispersal.

Drivers of biodiversitychange

Interactions

Type Potential interaction pathways

Climate change (CC) �Fragmentation Synergistic CC alters phenology and frugivores may be unable to track changes due to FR (⁄Tylianakis et al., 2008; Hegland et al., 2009)

Additive Altered wind patterns and frugivore movement between fragments (Damschen et al., 2008; Prevedello and Vieira, 2010)Biological invasion Synergistic CC gives advantage to BI phenology and crop outputs, and can increase establishment or distribution of invasive plants

(Crossman et al., 2011) and animals (Hellman et al., 2008); Morphology of invasive plants can adapt rapidly to CC, improvingdispersal capabilities (Zhang et al., 2011)

Frugivore loss Synergistic Plant species dependent on large-bodied frugivores unable to disperse to suitable habitats under CC (⁄Corlett, 2009)Fruit harvesting Antagonistic CC alters fruit abundance (Gordo and Sanz, 2010) which may make FH unviable

Fragmentation (FR) �Biological invasion Synergistic FR makes conditions more suitable for BI, which disrupt native dispersal networks (Chimera and Drake, 2010)

Antagonistic FR may slow spread of BI when long distance dispersal is rare (Alofs and Fowler, 2010)Frugivore loss Synergistic FR increases access for hunters (Kupfer et al., 2006); large-bodied frugivores decline more in smaller FR, resulting in reduced

dispersal of large-seeded plant species (Wotton and Kelly, 2011)Fruit harvesting Synergistic FR increases access for harvesters (Kupfer et al., 2006), who prefer large fruit (S. Prasad, pers. obs)

Biological invasion (BI) �Frugivore loss Synergistic BI more likely due to reduced competition following FL (⁄Chapin et al., 2000); Invasive seed predators target undispersed seeds

accumulated under parent plants (Wotton and Kelly, 2011)Antagonistic Invasive animal species replace native animals targeted for hunting (Desbiez et al., 2011)

Fruit harvesting Synergistic Invasion by alien plants more likely due to reduced seed rain of native species (⁄Chapin et al., 2000); Invasive plant species mayimpede regeneration of harvested species (Ticktin et al., 2006; Rist et al., 2010)

Additive Invasive plant species may reduce fruit production (Setty et al., 2008)

Frugivore loss (FL) �Fruit harvesting Additive Plant species dispersed by large-bodied frugivores are often key forest produce subject to intensive harvests (Rai, 2004,

S. Prasad, pers. obs)


and Tschinkel, 2006) and the effects on seed dispersal interactionsmay be compounding (Wotton and Kelly, 2011). Fragmentationmay also act synergistically with other drivers (Table 1), placing in-creased stress on fruit–frugivore interactions. Intensification of hu-man use is usually associated with increased levels of animal andplant harvesting, providing more opportunities for invasive speciesto enter ecosystems. Perhaps the most significant potentially syn-ergistic interaction involves climate change, since it may com-pound the effect of all other drivers of global change.

4. Incorporating seed dispersal knowledge into conservationmanagement

Knowledge of seed dispersal processes in human-modifiedlandscapes can be applied in order to slow-down habitat degrada-tion and biodiversity decline, accelerate the recovery of degradedareas, manage biological invasions and facilitate the adaptationof plants and animals to climate change. A functional understand-ing of the seed dispersal vectors within a community can providean invaluable tool for addressing a range of conservation problems– from deciding where to direct eradication methods for a biolog-ical invasion, to determining which frugivores are critical for eco-system maintenance and safeguarding their abundance andmobility across landscapes. Knowledge of fruit–frugivore relation-ships will expose gaps in the dispersal system where active man-agement, such as re-introduction of an animal, is required forecosystem maintenance.

4.1. Habitat fragmentation

Understanding of dispersal interactions in a system can aid con-servation managers tasked with mitigating the effects of fragmen-tation and restoring degraded lands. Because plant recruitment infragmented and degraded landscapes is often dispersal-limited(Cordeiro et al., 2009; Herrera and Garcia, 2010; Lehouck et al.,2009), the key actions required must increase dispersal of desirableplant species (Appendix A). Facilitating seed movement across

fragmented landscapes or into degraded areas can accelerate eco-system recovery and can reduce the costs of conservation effortsaddressing these concerns.

In fragmented landscapes, the preservation or creation ofhabitat corridors is probably the best way to enhance landscapeconnectivity for plants and their dispersal agents, particularly fordispersers that are habitat-specialists or have low mobility(Gilbert-Norton et al., 2010; Levey et al., 2005). However, corridorcreation often involves huge economic and social costs (Naidooand Ricketts, 2006) and less expensive options, such as steppingstones created by live fences, remnant trees, windbreaks may beadequate to enable movement of more mobile species (Estradaand Coates-Estrada, 2001; Galindo-González et al., 2000; Gilbert-Norton et al., 2010; Pizo and Dos Santos, 2010). Remnant treesand live fences also act as foci for seed rain into the matrix anddegraded areas (since they provide perching and roosting sites)and may provide suitable conditions for seedling establishment(Verdu and Garcia-Fayos, 1996; Pausas et al., 2006; Méndezet al., 2008; Herrera and Garcia, 2009), thereby creating a feasible,low-cost option for large-scale forest restoration in fallows andpastures (Martínez-Garza and Howe, 2003; Rodrigues et al.,2009). Perches and isolated trees could serve as ‘‘catchment areas’’for seeds for restoration projects and reduce costs involved in seedcollection. However, small-seeded wind and bird-dispersed speciesfrequently dominate recruitment in areas where these strategieshave been implemented (Martınez-Garza et al., 2009; Cole et al.,2010) and active management may be required for large-seededspecies.

Active introduction of keystone or generalist plant species(which attract a wide range of dispersal agents) into the matrixor degraded areas can accelerate recovery of degraded regions(Martínez-Garza and Howe, 2003; Sansevero et al., 2011). Networktheory could aid the identification of keystone plant species whichact as hubs or connectors in seed dispersal systems (Donatti et al.,2011; Mello et al., 2011b). Plant species that fruit during periods offood scarcity, such as figs, are also important targets for activeplanting (Lambert and Marshall, 1991). The diversity and quantity


of seed dispersal into restoration sites could also be accelerated byinnovative techniques such as the use of essential oils from fruitsto attract frugivorous bats (Bianconi et al., 2007) or the deploy-ment of artificial roosts (Kelm et al., 2008).

Restoration efforts should prioritize dispersal-limited plant spe-cies which may fail to arrive on their own. A functional classifica-tion of dispersal interactions within a community (Dennis andWestcott, 2006) can help identify such dispersal limited plant spe-cies. In the absence of information on community-wide fruit–frugi-vore relationships, late-successional or deep-forest species shouldbe prioritized for restoration planting as a precautionary measure.Introducing these species before they would arrive on their owncould attract vertebrate dispersal agents that can accelerate seedmovement into or across these regions (Martínez-Garza and Howe,2003).

Habitat restoration could also be accelerated by protecting,attracting, restoring or re-introducing frugivore populations(Appendix A), given that the quantity of animal-dispersed seedsis positively linked to frugivore abundance (Garcia et al., 2010).Particularly important are frugivore groups with large gape sizeand high mobility, thereby playing a disproportionate role in theestablishment of plant populations across breaks in habitat conti-nuity (e.g., large fruit pigeons from Madagascar through Asia andthe Pacific, e.g. Dew and Wright, 1998; Oliveira et al., 2002; Corlett,2009; frugivorous bats in tropical forests, e.g. Muscarella andFleming, 2007; Corlett, 2009; toucans in South America, Holbrook,2011; hornbills in Africa and Asia, e.g. Lenz et al., 2010; thrushes inEurope and the Americas, e.g. Jordano, 1993). Such key frugivorespecies are often hunted, and might exist below functional densi-ties required to maintain dispersal interactions. Because these fru-givores are usually wide-ranging they have rarely been assignedinternational or national protection status. Conservation prioritiesneed to reflect ecological function as well as the global risk ofextinction, and these large, mobile frugivores can provide key linksin fragmented systems.

Incorporating an understanding of seed dispersal of target plantspecies into spatial planning of restoration efforts can provide use-ful insights (van Loon et al., 2011), such as identification of areas ofwith high seed rain (which could be protected from livestock graz-ing and other threats to seedling recruitment) and areas with lim-ited dispersal where seeds need to be actively introduced.

4.2. Overharvesting

4.2.1. FrugivoresAn understanding of seed dispersal processes within communi-

ties can enable conservation managers to identify those speciesthat are disproportionately important for habitat maintenance.The use of network analyses can aid identification of frugivoresthat are most important (hubs and connectors) for maintainingthe structure of seed dispersal networks. Recent network studiesindicate that large frugivores, which consume a diverse range offruit, are frequently the strongest interactors (Donatti et al.,2011; Mello et al., 2011a,b). In the absence of community-wideinformation on fruit–frugivore interactions, large-bodied, large-gaped, and wide ranging frugivorous taxa should be targeted forconservation since these animals frequently have a disproportion-ate impact on ecosystem functioning (Fritz and Purvis, 2010;Campos-Arceiz and Blake, in press).

Instead of focusing solely on their extinction risk, species prior-itization for conservation action needs to reflect ecological func-tion, and population levels for prioritized species must bemaintained at levels which can conserve their ecological function(particularly given the risk of non-linear thresholds leading tothe cessation of ecological functioning at intermediate abun-dances; e.g., McConkey and Drake, 2006) (Appendix A). As our

knowledge of links between population abundance and ecologicalfunction increases, we should seek ways to predict sustainablepopulation densities from ecological indicators, which can be costand time effective (e.g., predicting appropriate densities of largefrugivores from seedling densities of large fruited plants).

In areas where functionally important seed dispersers have be-come extinct, we may need to re-introduce the species from otherregions. In situations where species have become globally extinct,alternative frugivore species have been introduced as ‘‘surrogate-dispersers’’ (e.g., tortoises in Mauritius; Hansen et al., 2010). How-ever, such approaches need to be considered judiciously becauseintroduced animals may provide much less effective dispersal thantheir native counterparts, may aid the dispersal of invasive plants,or may themselves become invasive (Christian, 2001; Hansen et al.,2010). Surrogates from the local or regional fauna are likely to bethe best candidates for introduction and they could be identifiedby considering functional roles at the community level.

4.2.2. PlantsHarvest levels for fruits collected by humans should be re-eval-

uated to ensure harvesting does not negatively impact frugivoreabundance and seed dispersal patterns (Appendix A). There arethree main scenarios in which fruit harvests could have criticalconsequences for dispersal function: regions with a low diversityof plants producing fruit, harvested species that fruit during overalllow periods of community-wide fruit availability and species thatmay offer potentially keystone resources for frugivores. In thesesituations, we need to find ways of ensuring that critical fruit re-sources are not driven to extinction by actively involving localstake holders to develop sustainable harvest practices (e.g., regu-lating harvest levels, improving harvesting techniques and settingaside populations free from harvests; Ticktin, 2004).

4.3. Biological invasions

Targeted programmes for the control of some invasive speciescan be developed from a functional understanding of commu-nity-wide frugivore-plant interactions (Appendix A). This knowl-edge can be used to facilitate predictions on which plants oranimals are most likely to establish within an ecosystem, whichvectors will promote the spread of the species (e.g., potential seeddispersers for invasive plants, and food sources for invasive frugi-vores) and where the species is likely to spread to (using dispersalkernels based on dispersal vectors) (Gosper et al., 2005; Buckleyet al., 2006; Murphy et al., 2008; Gosper and Vivian-Smith,2009). Most efforts to integrate seed dispersal research into inva-sive species management have taken place in Australia (Murphyet al., 2008; Westcott et al., 2008; Gosper and Vivian-Smith,2009), although invasive species are a widespread global concern.

In some cases invasive plants and animals become integratedinto native dispersal networks, and their role in these networksmay balance or even surpass their detrimental effects on nativebiota. In regions that have lost most native dispersers, invasive fru-givores are frequently the main seed dispersal agents of nativeplants (Chimera and Drake, 2010). Widespread invasive plantsare often integral parts of native frugivore diets, and removingthese might have negative effects on native frugivore populations,unless replaced by functionally-similar native plants (Gosper andVivian-Smith, 2009). It is therefore important that invasive speciescontrol is prioritized, according to the impacts (positive, neutral ornegative) the invasive species has on ecosystem functioning; andthat their potential, negative side effects are anticipated and cor-rected for (e.g. by introducing native fruiting plants or frugivores)during the control or eradication efforts.


4.4. Facilitating adaptation to climate change

We need to explicitly incorporate seed dispersal patterns intomodels predicting vegetation responses to global climate change.These models are needed for developing more robust predictionsof future carbon stocks and thus carbon-cycle feedbacks to the cli-mate system. They could also help managers mitigate impacts ofclimate change by identifying critical vector species, dispersal cor-ridors or keystone sites that facilitate plant migration (Williamset al., 2005) and develop better assisted migration programs forspecies that may require long-distance translocations (Hoegh-Guldberg et al., 2010; Vitt et al., 2010).

Policy makers and resource managers will have to considerproactive measures that are robust to the range of plausible cli-mate scenarios in anticipation of changes that will occur decadesin the future (Heller and Zavaleta, 2009; Appendix A). The speedof movement needed to track predicted temperature changes inthe 21st century has a global mean of 1.69 km a year or 169 kmper century (Chen et al., 2011), which is beyond the capacity ofmost long-lived woody plants as well as many plants of otherlife forms (Corlett, 2009; Nathan et al., 2011). Much faster move-ments will be required in the lowland tropics, where the tem-perature gradient is almost flat. In these cases, assistedmigration may be the only option (Hoegh-Guldberg et al.,2010; Vitt et al., 2010). However, many protected areas areestablished in rugged topography, where steep temperatureand rainfall gradients mean that cool and/or moist refuges mayexist within potential natural dispersal distances (which are typ-ically 100s to 1000s of metres per generation). In such land-scapes, a key strategy to facilitate seed movement is tostrengthen landscape connectivity (using approaches outlinedabove) across temperature and rainfall gradients (Appendix A).Managers can utilize knowledge of seed dispersal vectors to mi-mic, assist or enable species movement across gradients, therebyallowing species to naturally rearrange their distributions as re-quired by climate change (Millar et al., 2007).

Given the overlap of ecological, economic and social concernsthat are promoting the planting of trees at large scales (underthe Kyoto Clean Development Mechanism, REDD), multi-purposeplantations that address biodiversity concerns, augment localeconomies and sequester carbon need to be designed and imple-mented (Paquette and Messier, 2010). Incorporating an under-standing of seed dispersal (Appendix A) into the design of suchlarge scale reforestation projects can aid in promoting ecologicalfunction in these plantations, and may also accelerate the pace ofcarbon sequestration (Jansen et al., 2010). This calls for active col-laborations between dispersal ecologists, foresters and resourcemanagers.

5. Conclusions

The natural history of seed dispersal is now well-described ina range of landscapes worldwide (e.g., South Africa, Farwig et al.,2006; Lenz et al., 2010; Australia, Dennis and Westcott, 2006;Brazil; Almeida-Neto et al., 2008; Donatti et al., 2011; Thailand,Kitamura et al., 2002; Hong Kong, South China; Corlett, 2011;Spain, Herrera and Garcia, 2010; Jordano et al., 2007). As willbe apparent from this review, however, the idiosyncrasies ofthe plant and animal species involved have, at least so far, pre-cluded many robust generalizations that can be usefully appliedfor conservation management. Our recommendations for conser-vation management are therefore, at best, rules of thumb thatwill need to be supplemented by local natural history informa-tion. Fortunately, such information is often available, at leastfor large and charismatic species. Even where it is not, an aware-

ness of seed dispersal and its limitations will almost always leadto better conservation management than the alternatives ofassuming either that plants are immobile or that they are dis-persed everywhere. The needs of conservation managers shouldalso inspire better and more relevant ecological research. Seeddispersal ecologists need to expand their field of view in orderto study how seeds move—or fail to move—across real (frag-mented, heterogeneous) landscapes on the scales (kilometers)that matter for the long-term survival of plant and animal diver-sity in a changing world.

Acknowledgements

We thank Chanpen Wongsriphuek and Andres Link for discus-sions and comments on the manuscript. We thank the organizersof the 23rd Annual Meeting of the Society of Conservation Biologyat Beijing, China, where these ideas were first discussed. We alsothank three anonymous reviewers for their valuable commentson the manuscript.

Appendix A

Incorporating seed dispersal into conservation management foreach driver of biodiversity change.

Conservation action R
ecommendations
Habitat fragmentation
Enhance
connectivity
a) Linking corridors to facilitate
movement of dispersal vectorsand seeds (Levey et al., 2005; Gil-bert-Norton et al., 2010)

b) Where corridors are not feasible,maintain or create stepping-stones(Estrada and Coates-Estrada, 2001;Galindo-González et al., 2000; Gil-bert-Norton et al., 2010; Pizo andDos Santos, 2010).

c) Introduce plant species that playimportant functional roles in seeddispersal networks (e.g., keystoneand generalist plant species)(Mello et al., 2011a).

Conservationpriorities

Pm

rotect wide-ranging frugivores that canove across fragmented landscapes

Restoration ofdegraded lands

a) Accelerate succession by stimulat-ing visits by frugivores

b) Prioritize dispersal-limited plantspecies and generalist plant speciesfor restoration (Martínez-Garzaand Howe, 2003).

c) Protect sites with a potentially highdensity of seed rain (identifiedusing dispersal kernels) from live-stock grazing

Overharvesting of animals
Conservation
priorities
a) Prioritize conservation of animals
providing unique dispersal services,especially long-range dispersers

b) Identify minimum threshold densi-ties for key dispersers, below whichtheir ecological function declines

(continued on next page)


Appendix A (continued)

Conservation action R
ecommendations
Frugivorereintroductions

a) Re-introduce frugivores that havegone locally extinct, especiallythose which are functionally unique

b) Consider introduction of function-ally-similar frugivore species whenthe frugivore species has becomegloballyextinct (Hansen et al., 2010)

Overharvesting of plants
Harvest policies a) Re-evaluate harvest levels in which
frugivore populations may beaffected (e.g., in low-diverse regionsand or for key stone fruit resources)(Galetti and Aleixo, 1998).

b) Adopt better harvesting techniquesand set-aside source populations toensure maintenance of dispersalfunction (Ticktin, 2004)

Biological invasions
Eradication D esign eradication programs using seed
dispersal kernals to identify sites ofpotential infection sites of invasive plants(Murphy et al., 2008)

Identification S
creen fruit traits of introduced plants toassess their potential for invasiveness(Buckley et al., 2006)
Replacement S
creen fruit traits to identify native plantsthat can replace widespread invasivespecies (Gosper and Vivian-Smith, 2009)
Climate change
Enhance landscape
connectivityFi

ollow recommendations given above tomprove connectivity across temperature

and rainfall gradients within fragmentedlandscapes to enable movement of plantdispersal vectors

Halt frugivoredeclines

Fe

ollow recommendations given above tonsure functional densities, especially for

key frugivores providing unique dispersalservices

Afforestationpolicies

Id
ncorporate understanding of seedispersal processes in ongoing large-scale
afforestation programs(a) choice of species - prioritize dis-

persal-limited plant species andlate-successional plant species

(b) choice of sites - strengthen corri-dors and connectivities across tem-perature and rainfall gradients

Assisted migration W
hen migration capacity is exceeded,introduce species into potentiallysuitable habitats
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Seed dispersal in changing landscapes

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