ECOL250; Tropical Ecosystems and Biodiversity (2007)
Measures of resilience through functional diversity as displayed in tropical ecosystems with varying degrees of anthropogenic disturbance: Conservation implications Nils Harley Boisen Master of Science student in Management of Natural Resources, Institute for Management of Natural Resources, Norwegian University of Life Sciences, P.O.Box 5003, NO-1432 Ă…s
Author correspondence: firstname.lastname@example.org
ABSTRACT: The theoretical connection between ecosystem function and biodiversity is, albeit a fundamental topic in ecology, nevertheless a topic of controversy and recent intensified debate. This theory in its most simple definition postulates that a reduction in biodiversity results in a reduction in ecosystem level processes, and therefore more complex ecological communities should be comparatively more stable, i.e., resistant -/- resilient to perturbations. To date, only a minority of studies have tested this hypothesis under realistic levels of diversity, let alone levels of relatively high diversity as is the case in the tropics. There are obvious associations between the biodiversity /- ecosystem-function hypothesis and concerns within conservation biology. Thus, a common consensus is that this area of research may permit science to predict the consequences of extinctions on ecosystem properties, and the case for conserving biodiversity is strengthened if observed reductions in ecosystem function are the observed result of biodiversity loss. This paper thus addresses the validity of applying the biodiversity -/- ecosystem-function theory in the context of conservation. After giving a thorough description of the theory itself, the findings from six biodiversity-anthropogenic disturbance studies are discussed in light of their applicability to conservation. To be inclusive of common diverse ecosystems, this paper includes studies that have addressed major tropical- saltwater, aquatic, and forested terrestrial ecosystems. The conclusion of this paper is that the hypothetical mechanisms and underlying components relating to the theory of biodiversityâ€™s facilitation of ecosystem resilience are clearly relevant to planning of conservation policies when utilized in a proper context.
KEY WORDS: Biodiversity, conservation, ecosystem, functional, insurance, redundancy, resilience,
INTRODUCTION The theoretical connection between ecosystem function and biological diversity (from here on: biodiversity) is, albeit a fundamental topic in ecology, nevertheless a topic of controversy and recent intensified debate. This theory in its most simple definition postulates that a reduction in biodiversity results in a reduction in ecosystem level processes, and has deep academic and philosophical roots (Srivastava & Vellend 2005). The conventional perception is that the more complex an ecological community, the more interactions within, thus more ecological pathways available (Leveque 1995). With more ecological pathways, the blockage of one due to a disturbance will be compensated by the opening of another until the previous blockage has “healed” (Leveque 1995).The earliest foundations for this line of thought were laid by Elton (Elton 1958) and MacArthur (MacArthur 1955), and somewhat later by Tilman (Tilman 1982) who theorized that complex ecological communities are more stable, i.e., resistant -/- resilient to perturbations (Leveque 1995; Srivastava, et al. 2005). Perturbation in this context refers to stochastic environmental change, judged on a scale from small (slight random disturbance) to large (ecological “catastrophes”) (Leveque 1995). Stability in this context is thus a measurement of a given ecosystem’s, community’s, population’s, specie’s, etc, detection of and response to perturbations, in relation to the given perturbation’s scale (Leveque 1995).The magnitude of resilience could be measured by recording the amplitude of the largest stress from which a community or population can reattain previous equilibrium, though defining equilibrium is a difficult task as most communities and populations have a nature of spatial and/or temporal fluctuation. In light the recent increase in studies debating this topic, many review authors (e.g, Schwartz, et al. 2000; Srivastava, et al. 2005) state that the conventional theory has yet to hold water, and the topic has otherwise been fervently debated through its history (Huston 1997; Cameron 2002; Hooper, et al. 2005; Leigh 1975; May 1973, May 1975). There are many studies to date, both theoretical and experimental that vary in their support of this theory. Most are however criticized for being unusually simple with a restricted set of ecosystem functions and functional groups, low levels of diversity, and largely artificial conditions such as laboratory microcosms (Slade, et al. 2007). Such studies are unlikely to reflect the true conditions of natural communities as the community structure is misrepresented by highly artificial experimental design (Larsen, et al. 2005, (Schwartz, et al. 2000, (Slade, et al. 2007).
To date, only a minority of studies have tested this hypothesis under realistic levels of diversity, let alone levels of relatively high diversity as is the case in the tropics (Slade, et al. 2007). As changes in species composition triggered by human activity are predicted to influence ecosystem function, field research that measures ecosystem response to species removal or exclusion play an important role in the conservation value of this theory (Hector, et al. 2001; Loreau, et al. 2001; Schwartz, et al. 2000; Slade, et al. 2007; Srivastava, et al. 2005), as there are obvious associations between the biodiversity -/- ecosystem-function hypothesis and concerns within conservation biology (Srivastava, et al. 2005). Logically, the interconnection of ecosystem components (guilds, communities, populations, etc.) and ecosystem function makes the preservation of such components important for ecosystem function (Hector, et al. 2001; Nyström 2006; Nyström & Folke 2001; Schwartz, et al. 2006; Schwartz, et al. 2000; Srivastava, et al. 2005). There is a common consensus that this area of research may permit science to predict the consequences of extinctions on ecosystem properties, and the case for conserving biodiversity is strengthened if observed reductions in ecosystem function are the observed result of biodiversity loss (Srivastava, et al. 2005). This paper thus addresses the validity of applying the biodiversity -/- ecosystem-function theory in the context of biological conservation by giving a thorough description of the theory itself, and an introduction to the conservation debate before presenting the finds from six tropical biodiversity -/- anthropogenic disturbance studies and discussing them in light of their applicability to conservation. To be inclusive of common tropical ecosystems, this paper included studies that have addressed major tropical- saltwater, aquatic, and forested terrestrial ecosystems.
Possible Mechanisms in the Relationship between Biodiversity and Ecosystem Function Ecosystem function and ecosystem services are terms commonly used synonymously, the latter defined by Daily (1997) as “the conditions and processes through which natural ecosystems, and the species that make them up, sustain and fulfil human life” (Srivastava, et al. 2005). Such services could be crop pollination, protection from climatic disturbances, or even maintenance of biodiversity (Daily 1997), which is theoretically circular for ecosystem function (Srivastava, et al. 2005). The concept of ecosystem function is according to Srivastava (2005) best defined to date by Pacala & Kinzig (2002), who assign “ecosystem function” three separate categories that many distinct functions easily apply to; i) “stocks of energy and materials (e.g., biomass)” ii) “fluxes of energy or material processing (e.g., productivity, decomposition)” iii) “stability of rates or stocks over time”. Srivastava (2005) 3
lists potential mechanisms that underpin the possible positive relationship between biodiversity and ecosystem function. These are niche complementarity, functional facilitation, sampling effect, and dilution effect, described as relating to stock and flux functions, while the insurance effect, and portfolio effect are described as relating to stability functions (Srivastava, et al. 2005). Before defining these mechanisms of ecosystem function further down, it is necessary to understand the ecological concepts; keystone species, niche, functional group, key group, functional redundancy, and response diversity. Originally introduced in a predator study of a rocky intertidal zone by Paine (1969), the concept of keystone species states that in relation to its relative proportion of the community’s biomass, such a species is essential to maintaining the structure and diversity of its ecological community through their life history traits and interactions with the rest of the community. Solbrig (1991) recognizes three general classes of keystone species; i) species that through their trophic position maintain the diversity of competing organisms under them by controlling the abundance of potentially dominant species thus preventing competitive exclusion; ii) species that enter into different forms of mutualisms that link the fate of many partner species, and by means of this have a large influence on community structure and diversity; iii) species that provide keystone resources (ecosystem engineers) that are essential to reliant populations (Leveque 1995). In ecological terms, a niche refers to the collection specific needs and tolerances in relation to environmental conditions and resources, biotic and abiotic, that a given population requires to maintain sustainable numbers. A common way of putting it is that a species’ niche is its ecological “occupation”. A realized niche is the space of ecological factors a species occupies in the presence of competitors, predators, etc, while the fundamental niche is the space a species potentially occupies in the absence of such. Also called guilds, functional groups are an assembly of species, (often taxonomically similar) within the same community that perform similar ecological roles, i.e have similar niches. This concept was already suggested by Charles Darwin (Darwin 1859) who identified that “species of the same genus have usually, though by no means invariably, some similarity in habits and constitution”. The preservation of ecological processes is assumed to rest on such groups (Nyström 2006). Individual species within a functional group may occupy different trophic positions, and by possessing different traits in relation to different environmental conditions and resources overlap several different functional groups with other species (Nyström 2006). A key group is a particularly important sub-assembly of species within a functional group that play a dominant role regarding a particular function (Nyström 2006). 4
Niche complementarity is described as the greater ability of diverse communities or populations to efficiently exploit resources than depauperate ones due to a higher degree of niche differentiation between species or genotypes, resulting in higher ecosystem productivity and retention of nutrients (Bolger 2001; Loreau 2000). Loreau (2000) states that the effect of complementarity creates the foundation for permanent species associations that enhance collective performance, and according to Jonsson (2003) this is generally the most common and accepted explanation for positive species richness effects on ecosystem function. In contrast, Loreau (Loreau, et al. 2001) mentions that in unstable systems characterized by regular disturbances, ecosystem function is likely to be driven by the colonization ability and growth rate of individual species rather than niche complementarity. Functional facilitation is described as the increased ecosystem function associated with more diverse communities by possessing more species that have a positive effect on (facilitate for) the functional capabilities of another (Bruno, et al. 2003; Cardinale, et al. 2002; Srivastava, et al. 2005). The strength of the positive interaction between facilitation and ecosystem function is dependant on the functional characteristics of the species involved in relation to the context of the environment (Hooper, et al. 2005). Bruno (2003) states that recent research has uncovered striking influences of facilitation on communities, and that the theoretical framework of modern ecology has not kept pace with this. Sampling effect, also known as positive selection effect (Loreau 2000), combines species sorting mechanisms with probability theory, positing that the inclusion of a dominant, functionally important species increases with diversity when there is a positive covariance between the competitive ability of a species and its per capita influence on ecosystem function (Huston 1997; Srivastava, et al. 2005; Tilman, et al. 1997). As diversity increases, complementarity, facilitation, and sampling effects are likely to display saturated responses (Hooper, et al. 2005). When species differ significantly in functional traits, compatibility and/or facilitation, the weakening of these mechanisms will expectedly affect ecosystem function after species loss has resulted in highly impoverished communities (Hooper, et al. 2005). The Dilution effect hypothesizes that because of frequency dependant selection among species or genotypes, specialized pathogens -/- predators may have less per capita effect on species and/or genotypes that occur in low densities in highly diverse communities and/or populations (Mitchell, et al. 2002; Srivastava, et al. 2005). Regarding mechanisms that contribute to the stability of ecosystem functions, the insurance effect, known better as the insurance hypothesis, states that redundancy in functional roles (functional redundancy) allows a given speciesâ€™, populationâ€™s, or 5
communityâ€™s net functional contribution to the ecosystem to be maintained after a perturbation (Baker, et al. 2004; Srivastava, et al. 2005; Yachi & Loreau 1999). Thereby, possible negative effects of disturbances are buffered because of the compensation among species (Walker, et al. 1998; Yachi, et al. 1999). However, the concept of response diversity is also important to the insurance effect. Response diversity simply implies a diversity of possible responses between species to potential perturbations within a functional group (NystrĂśm 2006). It is theorized that such diversity is essential for ecosystem resilience by insuring that a portion of a functional group survives a possible large scale disturbance and thereby continue performing the groupâ€™s given function while other species rebound. Finally, the portfolio effect states (without requiring species interactions) that compared to the fluctuations of any one species, the fluctuation of many individual species can display less unpredictability as a whole (Srivastava, et al. 2005; Tilman, et al. 1998).
The Conservation Relevance of the Biodiversity -/- Ecosystem-function Theory Research addressing biodiversity and ecosystem function emerged at a time when Western public interest in conservation was at a peak. Conservation biology is according to Swartz (2000) characterized by enthusiastic support for key ideas, such as the biodiversity- ecosystem function theory, that fail to meet expectations as a conservation framework when empirically tested. A fundamentally relevant question for conservation pertaining to biodiversity -/ecosystem-functioning is if species loss relates to reduced ecosystem function (Srivastava, et al. 2005). The loss of one species is likely to alter the probability of further species loss, particularly if the species lost displays an important trophic role or other interaction (Hector, et al. 2001). Srivastava (2005) cautioned the conservation community from endorsing biodiversity -/- ecosystem functioning linkages as a promotion model in conservation. Though 71 of 100 biodiversity -/- ecosystem-functioning studies examined by Srivastava (2005) demonstrated a positive correlation between diversity and at least one ecosystem function, 53 of the positive relationships were log linear. Many authors (e.g., Swartz, 2000) dispute that this implies numerous species can be lost from a system before there is noticeable decrease in function. Selective loss of apex species may nonetheless have particularly strong ecosystem repercussions (Srivastava, et al. 2005). Srivastava (2005) points out that not only a given extinction or set of extinctions may be the major degrader of ecosystem function, but in fact the nature of the stressor that has caused the extinction/s depending on the covariance in species tolerances. Moreover, Srivastava (2005) maintains that though conserving diversity 6
ensures stability through a range of response to future change, the argument for conserving diversity to maintain function is often more feeble regarding stocks and fluxes than it is for conserving stability. Based on their findings this particular meta-study concluded that biodiversity -/- ecosystem-function theory is not likely to constructively aid conservation managers in practical decisions (Srivastava, et al. 2005). Hector’s (2001) meta study argues the opposite, stating that though an asymptotic relationship between biodiversity and ecosystem functioning is observed on short term local scales, at larger spatial and temporal scales the loss of biodiversity is likely very important to function. Previous to the millennium shift most studies on the topic were rather small scale (<100m2), thus irrelevant for conservation policies commonly implemented at far larger scales. Moreover, it seems that most studies on the topic did not realistically address community dynamics (Hector, et al. 2001), and in concurrence with Swartz (2000) and Srivastava (2005), Hector (2001) states that generalization is obscured between most studies due to differences between systems, functions measured, level of diversity manipulated, in addition to experimental approaches. Nonetheless, Hector (2001) argues that positive relationships between biodiversity and ecosystem functioning appear to be relatively common and thus a strengthening argument for conservation despite the relationships not always being strong linear ones with equal species contribution. “We expect species richness to be more important when ecosystem stability is dependent on the functioning of a wide range of different processes that are influenced by different sets of species.” (Hector, et al. 2001). Arguably, though different species potentially contribute disproportionately to ecosystem processes, the identification of all important species is unfeasible, especially in highly diverse environments (such as the tropics) (Hector, et al. 2001). Moreover, it cannot be refuted that such species are themselves intrinsically dependant on other species of their community in ways that have yet to, or cannot be, described (Hector, et al. 2001). Ecosystem functioning arguments for conservation correspond to other conservation arguments that are justified by utilitarian, ethical, or aesthetic attitudes, producing a strong case for conservation (Hector, et al. 2001). It seems that since Hector’s publication in 2001 a number of studies, such as five of the six below described, tend to concur with Hector and patch the holes in this theory’s relevance for conservation (Lévêque’s (1995) was pre- millennium shift, though still relevant).
Tropical Biodiversity-Disturbance studies; Saltwater Systems Published in 2005, a study from 1998 titled “Changes in four complementary facets of fish diversity in a tropical coastal lagoon after 18 years: a functional interpretation”, was conducted in Terminos lagoon, located in the Southern Gulf of Mexico. This study attempted to identify the spatial and temporal changes in species composition of the piscatorial inhabitants using 4 diversity indices linked to 4 independent biodiversity components. The second objective was to discuss the functional implications of diversity changes, by linking possible changes in the fish community with published information on observed environmental variation (Miranda, et al. 2005). Terminos lagoon is described by Miranda (2005) as being a lagoonal estuarine ecosystem consisting of critical habitat for larval and juvenile finfish species, and supporting commercial fisheries. The basin surface area is 1661.50 km2 and its average depth 3.5 m (Miranda, et al. 2005). Despite being a protected area since 1994 through national recognition of its ecological and economic importance, adequate management is made difficult due to urban growth on the main island, continued oil extraction on the continental shelf, as well as development activities for the oil infrastructure and agriculture performed on the watershed (Miranda, et al. 2005). By means of sampling conducted in 1980 to 1981, and again ca. 18 years later in 1997 to 1999, a significant decrease was displayed in lagoon’s nekton richness and evenness (Miranda, et al. 2005). Between the two study periods there was observed a significant increase in the salinity and water temperature of Terminos lagoon suggesting a degradation of estuarine conditions and a shift from hypohaline to euhaline/hyperhaline conditions. Likewise, during the final survey some fish families were in fact better represented than in the first. As a result of the large reduction in diversity, two randomly chosen individuals or species from within a fish assemblage were more likely to be closely related in the final survey than in the first, indicating according to the author a homogenization of functional diversity (Miranda, et al. 2005). This is based on the study’s shared notion with Darwin (Darwin 1859) that taxonomically related species are more likely to share some ecological attributes than taxonomically unrelated ones (Miranda, et al. 2005). As a result of the hypothetical reduction in functional diversity, Miranda (Miranda, et al. 2005) postulates further that the this loss will undermine Terminos lagoon’s ecological responsiveness to disturbances by means of a degraded insurance effect, in addition to an impeded functioning of the ecosystem. An example of a possible effected function given by the author are the lagoon’s phytoplanktivorous and predatory guilds, the first considered 8
complementary with the latter, providing a secondary production as prey whereas the latter guild is usually migratory, thus exporting biomass and thereby nutrients into the lagoon (Miranda, et al. 2005). Regarding resilience, Miranda (Miranda, et al. 2005) mentions an increases susceptibility of invasion by non-natural species as a result in openings in resource availability caused by local niche extinctions. Published in 2006, a review article by Magnus Nyström (2006) titled “Redundancy and Response Diversity of Functional Groups: Implications for the Resilience of Coral Reefs”, analysed the role of diversity within functional groups in securing vital ecosystem processes that contribute to the resilience of coral reefs in a human dominated biosphere. Three functional groups (zooxanthellae (symbiotic micro algae in reef-building corals), reefbuilding corals, and herbivores) were used to illustrate two major components that confer ecosystem resilience in coral reefs, being functional redundancy and the diversity of responses to change within functional groups. Coral reefs are described by Nyström (2006) as being keystones of economic and societal development throughout tropical coastal areas, with corals being a key group within the functional group ‘‘habitat builders”, and states that these are simultaneously being degraded at a rapid pace. Human activity correlates strongly with the observed regional disruption of coral reef ecosystems (Nyström 2006). According to the author, studies show that coral reefs are complex and non-linear like other ecosystems with smooth change being easily and abruptly interrupted, in contrast to the common notion that coral reefs in the face of gradual environmental changes are expected to respond in a smooth linear fashion (Nyström 2006). Nyström (2006) states that it is the loss of ecosystem resilience that facilitates such abrupt transitions, e.g., coral dominance to macroalgal dominance. Even subtle disturbances within coral systems can result in impeded ecosystem resilience that is only recognized after a reef fails to properly compensate in the aftermath and thus shifts to a dramatically different state that the system could have otherwise withstood (Nyström 2006). Zooxanthellae are a genetically divers group made up of many different clades that are a very important functional group on coral reefs through their symbioses with coral (Nyström 2006). It has been demonstrated that different species of reef building corals can be populated by specific clades, or that individual coral structures can host a multi-clade community simultaneously (Nyström 2006). A large scale reduction in clades due for example to prolonged temperature increase can result in coral mortality and extensive reef-system damage. A recent study (Baker, et al. 2004) demonstrated that a specific clade of Zooxanthellae shown to be more tolerant of elevated temperatures (Rowan 2004) was more 9
abundant in corals after a harsh mass bleaching event. Thus this clade may provide adequate redundancy for host survival while the more fitting cladeorclades re-establish themselves in the community (Nyström 2006). A variety of tolerance between different clades to different types of perturbations illustrates response diversity according to Nyström (2006) helping to insure the survival of coral systems. A broad variety of functional roles are provided by reef building corals (Bellwood, et al. 2004), and the provision of habitat is perhaps the most essential. The diversity of coral morphology, and distribution of such determines a reef’s degree of structural heterogeneity and thus habitat quality for other organisms (Nyström 2006). Despite loss of structural heterogeneity due to shifts in community structure resulting from disturbances, the diversity of coral reefs can insure the continued provision of habitat construction (Nyström 2006). When Belizean staghorn (Acropora cervirornis) reefs suffered extensive mortality in the 1980s as a result of disease and high temperature, an uncommon lettuce coral (Agricia tenuifolia) became the dominant reef builder (Aronson, et al. 2004), illustrating functional redundancy despite other corals not being sufficient substitutes for habitat created by branching species (Nyström 2006). The morphology of coral can be critical for community reestablishments after destruction, displaying an important aspect of response diversity which can for example make branching coral species essential due to their traits of rapid recolonization once conditions have stabilized (Nyström 2006). Compared to western Pacific reefs, Caribbean reefs have potentially less ecological insurance due to a lesser degree of functional diversity and species diversity within functional groups making the Caribbean more vulnerable to catastrophic shifts in community structure (e.g., to algal dominance) (Nyström 2006). As a result, Caribbean reefs may be more vulnerable to human impacts such as pollution and global warming, and the functional redundancy and response diversity in Caribbean coral reefs may be more easily eroded by degraded availability, ‘health,’ and diversity of coral larvae and selection for particular functional traits (coral types) (Hughes, et al. 2003). By facilitating succession, reef herbivory is considered an underpinning mechanism of the continuity of dominant reef states (Bellwood, et al. 2004; Nyström 2006). Though herbivore species diversity is relatively low compared to other functional groups of fish, reef herbivore fish possess a wide variety of feeding strategies, preferences, and behaviours (Nyström 2006). Depending on species and size for example, reef herbivores can individually have distinct functional impacts on reef ecosystem processes through algae control (Nyström 2006). Aspects of herbivore functionality in this context can be for example the type of algae 10
they remove, or how they feed (Nyström 2006). Based on feeding strategy, three functional herbivore groups have been recognized; grazers that prevent algae overgrowth, scrapers that facilitated coral recruitment, and bioeroders that remove coral (Bellwood, et al. 2004; Nyström 2006). The functional redundancy of these functional groups can vary to large degrees, but a reduction in their numbers has nevertheless displayed major changes in the balance between corals and algae (Bellwood, et al. 2004). Fisheries for example often target grazing herbivore species that remove fleshy algae, subsequently resulting in a loss of coral reef resilience (Nyström 2006).
Tropical Biodiversity-Disturbance studies; Fresh Water systems in Africa Titled “Role and consequences of fish diversity in the functioning of freshwater African ecosystems”, this review from 1995 by Lévêque, analysed previously published data from freshwater ecosystems in Africa in order to address the biodiversity -/- ecosystem functioning concept. According to Lévêque (1995) at the time of this study the concept of keystone species had not been investigated for African freshwater fish, though keystone predators had been demonstrated. The best known demonstration of proven keystone predators is the case of the introduction of the apex predator Nile perch (Lates niloticus) in Lake Victoria which had a cascading effect in the lake’s trophic structure resulting in the extinction or rarity of numerous endemic haplochromine cichlids (Haplochrominae spp.) and a possible increase in phytoplanktonic production (Goldschmidt, et al. 1993). Haplochromine cichlids have over evolutionary time undergone an unparalleled species diversification that has displayed an incredible degree of functional morphology with only slight morphological differences resulting in profound differences in species feeding habits (Leveque 1995). Ecologists have long pondered over how the cichlid communities of the African Great Lakes which apparently share very similar niches can coexist without competitive exclusion. One theory regarding Lake Victoria’s diverse assembly of piscivorous haplochromines is that functionally redundant species are individually localised and rarely coexist. In addition, piscivorous haplochromine cichlids are shown to be specialized predators with strong prey preferences despite being fundamentally capable of possessing a more diverse diet (Leveque 1995). Although able to feed on a range of organisms, the various species have a strong preference for certain food items. Moreover, Lévêque (1995) states that (at the time of the study) niche segregation had been demonstrated also among the zooplanktivorous haplochromines of Lake Victoria in spite of apparent niche overlapping. 11
Published in 2006, a study by Swartz et al. from 1999 titled “Effects of Nile perch, Lates niloticus, on functional and specific fish diversity in Uganda’s Lake Kyoga system”, was conducted in seven lakes in the Lake Kyoga satellite system located downstream from Lake Victoria, as the title states, in Uganda. This study’s research attempted to elucidate the effects of Nile perch on Kyoga system fish communities, and to describe the utility of the Kyoga system as a natural reference experiment for fisheries management. In combination with increased eutrophication, the population boom of the artificially stocked apex predator Nile perch in Lake Victoria, Lake Kyoga, and many satellite lakes has with increasing speed eradicated hundreds of endemic fishes and drastically reorganized the food web structure while creating the worlds largest freashwater fishery (Schwartz, et al. 2006). Entire functional groups of fish and even some birds and thus their related ecological services began disappearing in the 1980s on to the present. “Many species of shell-crushing molluscivores, epilithic algae scrapers, higher plant eaters, detritivores paedophages, planktivores, piscivores, scale-eaters, and specialized insectivores all became extremely rare, and a myriad of species are apparently extinct”. By examining δ15N (the difference between sampled 15N/14N ratios and a standard) signatures from the epaxial muscle from a subsample of catch (n= 361) to study the trophic structure of the and Lake Kyoga ecosystem, Swartz (2006) found that the perch free lakes possessed larger δ15N variance and more fluctuating fish lengths than the Nile perch lakes which suggested a significant relationship between species diversity and functional diversity. Lakes with Nile Perch were less diverse, had larger fish, and a less complex trophic structure (Schwartz, et al. 2006). This study suggests that the eradication of highly specialized fish species from the lakes has reduced the number of pathways from primary production to apex consumers in the fish community and therefore likely reducing the potential long term yields for the fishery (Schwartz, et al. 2006). Swartz (2006) hypothesizes that functional and behavioural plasticity can potentially mediate species loss.
Tropical Biodiversity-Disturbance studies; Arthropods in Tropical forest Systems Titled “Experimental evidence for the effects of dung beetle functional group richness and composition on ecosystem function in a tropical forest”, by Slade et al. (2007), this study looked at the effects of dung beetle (Scarabaeinae spp.) functional group richness and composition on two interlinked and functionally important ecological processes, by utilizing a combinatorial experimental design and manipulating species at natural densities in an evergreen tropical forest in Danum Valley Conservation Area, Sabah, Malaysian Borneo. The 12
two interlinked and functionally important ecological processes in question were dung removal and secondary seed dispersal. This study is believed by the author to be the first experimental field manipulation of functional group richness for a terrestrial animal group that quantifies the contribution of functional groups to ecosystem function at realistic levels of diversity. There is relatively little hunting pressure in the Danum Valley Conservation Area, and all of Borneo’s large mammals are found there. Based on dial activity, size, and method of dung removal, dung beetles can be separated into distinct functional groups. This study observed Paracoprid nesters (tunnellers) that bury dung below the dung pile, and telecoprid nesters (rollers) that make a dung ball and roll it from the source pile before burying it elsewhere. Two other guilds (endocoprids and kleptocoprids) were considered as having minimal effects on dung removal during the study and were not included. In particular, Slade’s (2007) study results emphasized that not only functional group identity, but also a given functional group’s species composition is important in determining the ecological consequences of extinctions or altered patterns of species abundance. Moreover, the composition of functional groups in addition to richness was key. Dung beetle functional group richness correlated positively with both dung and seed removal across the study with no evidence of competitive exclusion (Slade, et al. 2007). The degree of ecosystem functioning however was functional-group-specific, which indicated the importance of functional group composition (Slade, et al. 2007). The study demonstrated evidence that indicated significant complementarity as well as facilitation among the groups, though complementarity was limited to specific group combinations. Dung removal was reduced by ca. 75 percent in the absence of the large nocturnal tunnellers giving this group the largest (most disproportionate) influence on ecosystem function (Slade, et al. 2007). Thus, maintenance of complete and proper ecosystem functioning requires intact functional groups composition in the given ecosystem (Slade, et al. 2007). The other groups would likely fail to compensate for the loss of the nocturnal tunnellers (Slade, et al. 2007). Published in 2006, a review study by Philpott (2006). titled “Biodiversity in Tropical Agroforests and the Ecological Role of Ants and Ant diversity in Predatory Function”, was carried out for the purpose of addressing, among other things, what patterns and mechanisms drive the loss of ant (Formicidae spp.) diversity under agricultural intensification, and the effects of ant diversity on the functional role of ants as predators through control of pests agents and fungal diseases. Philpott (2006) states that ants are important predators in tropical agroforestry ecosystems and that the species diversity of this insect group generally declines with agroforestry (e.g., coffee and cacao) intensification. How this loss of ant diversity may 13
affect the functional role of ants as predators is not a topic addressed among the literature (Philpott & Armbrecht 2006). Structurally diverse agroforestry systems are well known for their advantages to the facilitation and conservation of biodiversity (Philpott & Armbrecht 2006). Typical tropical examples are the shade cultivation regimes of coffee (Coffea spp.) and cacao, (Theobroma cacao) (e.g., Beer, et al. 1997; Moguel & Toledo 1999; Perfecto, et al. 2005; Reitsma, et al. 2001). The conversion of the traditional shade regimes to sun exposed monocultures has been shown to result in a of loss of arthropod biodiversity (Moguel, et al. 1999; Perfecto, et al. 2003; Perfecto, et al. 1997; Perfecto, et al. 2005), negatively affecting the potential role arthropods, and ants as such, as biological control agents (Philpott, et al. 2006). Many groups of ants such as tramp species, forest specialists, and predatory specialists are sensitive to habitat changes such as agricultural intensification (Andersen, et al. 2002; BrĂźhl, et al. 2003), and many groups show a strong negative response to coffee and cacao intensification (Perfecto, et al. 1997). Specifically regarding agricultural intensification, nest site limitation has received the most attention as a mechanism that limits ant diversity (Slade, et al. 2007). Ants are often perceived as pests, and therefore it is important for agriculture that their true ecological role is explored as they are extremely diverse and omnipresent in tropical forests. More genera of ants are predators than any other tropical trophic guild, making predatory ants a functional group that can display a range of responses to unpredictable pest outbreaks (Philpott, et al. 2006). According to Yachi (1999), predator functionality can through the insurance hypothesis be increased by means of intraspecific differences in foraging or behaviour. Though homopterans are predominantly seen as an agricultural pests, Way (1992) suggests that this groupâ€™s relation to ants can be beneficial as ants will be attracted to specific foraging sights. Other studies concur, demonstrating increased plant growth and fitness under circumstances of occupant homopterans attracting ants that then deter other more detrimental herbivores (e.g., Dixon 1971; Messina 1981; Room 1972). In Vietnamese agroforestry two ant species (D. thoracicus and O. smaragdina ) that tend homopteran trophobionts are nonetheless included in pest management programmes (Van Mele & van Lenteren 2002). This is not only because of increased ant predation, but the ant-induced increase in sweeter honeydew produced by homopteran trophobionts may facilitate the attraction of hymenopteran or dipteran parasitoids. Vandermeer et al. (Vandermeer, et al. 2002) found Azteca ant activity correlated with green scale (Coccus viridis) concentration on plants in shade coffe regimes and under such circumstances ants significantly reduced damage by 14
introduced lepidopteran larvae thus outweighing the slighter potential damage caused by their homopteran “livestock”. Facultative ant- homoptera mutualism can be beneficial in coffee production as observed by Perfecto (Perfecto & Vandermeer 2006) who found that coffee plants occupied by fewer homopterans suffered significantly higher damage to the coffee berry borer (Hypothenemus hampei) du to attracting fewer Azteca ants. Philpott’s (2006) conclusion from the review was that agroecosystem function and conservation goals will be advanced in systems where ant diversity is high and facilitated for by humans such as in traditional agroforestry regimes.
CONCLUSIVE DISCUSSION The above described studies demonstrate how anthropogenically driven fragmentation of habitat results in reorganization of relative species composition as well as loss, with major potential repercussions on associated ecosystem functioning and resilience. This change may not be visible in a short temporal perspective, but manifest itself slowly over longer time is shown from Miranda’s (2005) study. Change can also be spatial. When similar though highly separate ecosystems differ in their retention of functional- and species diversity within functional groups, certain ecosystem may be more vulnerable to disturbances and change as is likely the case between the Caribbean and Pacific coral reef ecosystems. Local communities in regions with high levels of species richness should logically experience a saturation of immigrating species, thus regional extinction events should result in reduced ecosystem resilience by robbing local communities of their species saturation. This seems very relevant especially for Lake Victoria’s once highly diverse assemblage of niche overlapping haplochromines that kept themselves segregated enough to coexist. Rare interbreeding sometimes made new species, while local extinctions were compensated for by the immigration from the nearest comparable species. The regional extinctions of many of these species due to the Nile perch has made it impossible for the diversity of many haplochromine assemblages to repair itself. Or, as Philpott’s (2006) study, agroecosystem functioning (at least regarding ant depredation) and conservation goals can be optimized if species rich heterogeneous cultivation systems such as agroforestry are maintained at a regional level in highly fragmented landscapes. Predicting the ecological consequences of local extinctions requires consideration of species biomass as well as abundance in regards to the identity and composition of a given functional group. The eradication of highly specialized fish species from the freshwater regions of Lake Victoria and a number of surrounding lakes has reduced the number of 15
pathways from primary production to apex consumers in the fish community and therefore likely reducing the potential yields for the fishery (Schwartz, et al. 2006). Swartz (2006) hypothesizes that functional and behavioural plasticity can potentially mediate species loss, which is theoretically possible since specialized predators that display strong prey preferences (such as the piscivorous haplochromine cichlids) are still fundamentally capable of possessing a more diverse diet. However, as observed from this area a reduction in predatory species has damaged the predominantly top-down regulation within the community due to reduced functional diversity at this trophic level. Increased predator diversity increases the probability of the (in this case piscatorial) community containing key predators, thereby enhancing this trophic levelâ€™s function (Hooper, et al. 2005). Conserving a functionally diverse predator community in behaviour and diet is important to ensure the long term resilience to environmental change (Philpott, et al. 2006). Though the functional redundancy of functional groups can vary to large degrees, a reduction in their numbers can result in major changes in ecosystem balance such as has been displayed by a reduction in herbivore groups on the balance between corals and algae. As shown by Slade (2007), some functional groups may contribute more than others and therefore the intact composition of functional groups in a given ecosystem must be maintained for the sake of the ecosystemâ€™s continual functionality, as the loss of a major contributing group may not be easily compensated for. This is also true for species within a functional group. Certain species can give the impression of being more important than others regarding specific functions and their loss be detrimental to ecosystem function, however this does not mean that those species not giving the same impression are unimportant as such species may be the new drivers of ecosystem function when the environment changes. Moreover, though species belonging to a specific functional group are functionally redundant, they may in fact have differing responses to environmental change and/or disturbances (Hooper, et al. 2002; Larsen, et al. 2005), demonstrating an important issue for the potential reactions of ecosystem function to such. The response diversity of symboints can be extremely important, as shown from among the Zooxanthellae clades on Coral. If it were not for such response diversity among symbionts many biomes would look quite different regarding the large controlling effects of many types of microscopic symbionts such as mycorrhiza, nitrogen fixing bacteria, luminescent bacteria in deep sea fish, etc. In conclusion it seems that the hypothetical mechanisms and underlying components relating to the theory of biodiversityâ€™s facilitation of ecosystem resilience is clearly relevant to planning of conservation policies when utilizing a proper perspective. A proper perspective will apply the conservation implications of this theory in a correct spatial and temporal 16
context, with realistic regards to community structure as well as ecosystem threats. The ecosystems of the world’s tropical biomes possess the highest comparative amounts of biodiversity, and are simultaneously experiencing the highest rates of biodiversity loss. As the functioning of these ecosystems comes under increasing stress it is clear that their biodiversity must be conserved to preserve long term functioning and resilience. These important implications for conservation also spill over to the preservation of ecosystem services that are essential for humanity.
LITERATURE CITED ANDERSEN, A. N., HOFFMANN, B. D., MULLER, W. J. & GRIFFITHS, A. D. 2002. Using ants as bioindicators in land management: simplifying assessment of ant community responses. Journal of Applied Ecology 39(1):8-17. ARONSON, R. B., MACINTYRE, I. G., WAPNICK, C. M. & NEILL, M. W. 2004. Phase shifts, alternative states, and the unprecedented convergence of two reef systems. Ecology 85(7):1876-1891. BAKER, A. C., STARGER, C. J., MCCLANAHAN, T. R. & GLYNN, P. W. 2004. Corals' adaptive response to climate change. Nature 430:741. BEER, J., MUSCHLER, R., KASS, D. & SOMARRIBA, E. 1997. Shade management in coffee and cacao plantations. Agroforestry Systems 38(1):139-164. BELLWOOD, D. R., HUGHES, T. P., FOLKE, C. & NYSTRÖM, M. 2004. Confronting the coral reef crisis. Nature 429:827- 833. BOLGER, T. 2001. THE FUNCTIONAL VALUE OF SPECIES BIODIVERSITY—A REVIEW. Biology and Environment: Proceedings of the Royal Irish Accadamy 101B(3):199–224. BRUNO, J. F., STACHOWICZ, J. J. & BERTNESS, M. D. 2003. Inclusion of facilitation into ecological theory. Trends in Ecology & Evolution 18(3):119-125. BRÜHL, C. A., ELTZ, T. & LINSENMAIR, K. E. 2003. Size does matter – effects of tropical rainforest fragmentation on the leaf litter ant community in Sabah, Malaysia. Biodiversity and Conservation 12(7):1371-1389. CAMERON, T. 2002. 2002: the year of the `diversity - ecosystem function' debate. Trends in Ecology & Evolution 17(11):495-496. CARDINALE, B. J., PALMER, M. A. & COLLINS, S. L. 2002. Species diversity enhances ecosystem functioning through interspecific facilitation. Nature 415(6870):426-429. DAILY, G. C. 1997. Nature’s Services: Societal Dependence on Natural Ecosystems. Island, Washington, DC. DARWIN, C. 1859. The origin of species by means of natural selection. Murray, London. DIXON, A. F. G. 1971. The Role of Aphids in Wood Formation. II. The Effect of the Lime Aphid, Eucallipterus tiliae L. (Aphididae), on the Growth of Lime, Tilia x vulgaris Hayne The Journal of Applied Ecology 8(2):393 - 399. ELTON, C. S. 1958. The Ecology of Invasions by Animals and Plants. Chapman and Hall, London. GOLDSCHMIDT, T., WITTE, F. & WANINK, J. 1993. Cascading Effects of the Introduced Nile Perch on the Detritivorous/Phytoplanktivorous Species in the Sublittoral Areas of Lake Victoria. Conservation Biology 7(3):686-700.
HECTOR, HECTOR, A., JOSHI, JOSHI, J., LAWLER, LAWLER, S., SPEHN, SPEHN, E. M., WILBY & WILBY, A. 2001. Conservation implications of the link between biodiversity and ecosystem functioning. Oecologia 129(4):624-628. HOOPER, D., SOLAN, M., SYMSTAD, A., DIAZ, S., GESSNER, M., BUCHMANN, N., DEGRANGE, V., GRIME, P., HULOT, F., MERMILLOD-BLONDIN, F., ROY, J., SPEHN, E. & VAN PER, L. 2002. Species diversity, functional diversity, and ecosystem functioning. Loreau, M., Naeem, S. & Inchausti, P. Biodiversity and Ecosystem Functioning : Synthesis and Perspectives. Oxford University Press. HOOPER, D. U., CHAPIN, F. S., EWEL, J. J., HECTOR, A., INCHAUSTI, P., LAVOREL, S., LAWTON, J. H., LODGE, D. M., LOREAU, M., NAEEM, S., SCHMID, B., SET, H., SYMSTAD, A. J., VANDERMEER, J. & WARDLE, D. A. 2005. EFFECTS OF BIODIVERSITY ON ECOSYSTEM FUNCTIONING: A CONSENSUS OF CURRENT KNOWLEDGE. Ecological Monographs 75(1):3-35. HUGHES, T. P., BAIRD, A. H., BELLWOOD, D. R., CARD, M., CONNOLLY, S. R., FOLKE, C., GROSBERG, R., HOEGH-GULDBERG, O., JACKSON, J. B. C., KLEYPAS, J., LOUGH, J. M., MARSHALL, P., NYSTROM, M., PALUMBI, S. R., PANDOLFI, J. M., ROSEN, B. & ROUGHGARDEN, J. 2003. Climate Change, Human Impacts, and the Resilience of Coral Reefs. Science 301(5635):929-933. HUSTON, M. A. 1997. Hidden treatments in ecological experiments: re-evaluating the ecosystem function of biodiversity. Oecologia 110(4):449-460. JONSSON, M. & MALMQVIST, B. 2003. Mechanisms behind positive diversity effects on ecosystem functioning: testing the facilitation and interference hypotheses. Oecologia 134(4):554-559. LARSEN, T. H., WILLIAMS, N. M. & KREMEN, C. 2005. Extinction order and altered community structure rapidly disrupt ecosystem functioning. Ecology Letters 8(5):538-547. LEIGH, E. G. 1975. Population fluctuations, community stability and environmental variability. The Belknap Press of Harvard University Press, Cambridge. 51-73 pp. LEVEQUE, C. 1995. Role and consequences of fish diversity in the functioning of African freshwater ecosystems. AQUATIC LIVING RESOURCES 8(1):59. LOREAU, M. 2000. Biodiversity and ecosystem functioning: recent theoretical advances. Oikos 91(1):3-17. LOREAU, M., NAEEM, S., INCHAUSTI, P., BENGTSSON, J., GRIME, J. P., HECTOR, A., HOOPER, D. U., HUSTON, M. A., RAFFAELLI, D., SCHMID, B., TILMAN, D. & WARDLE, D. A. 2001. Biodiversity and Ecosystem Functioning: Current Knowledge and Future Challenges. Science 294(5543):804-808. MACARTHUR, R. 1955. Fluctuations of Animal Populations and a Measure of Community Stability Ecology 36(3):533-536 MAY, R. M. 1973. Stability and Complexity in Model Ecosystems. Princeton University Press, Princeton. MAY, R. M. 1975. Patterns of species abundance and diversity. The Belknap Press of Harvard University Press, Cambridge. 81-120. pp. MESSINA, F. J. 1981. Plant-protection as a consequence of an ant – membracid mutualism – interactions on goldenrod (Solidago sp.). Ecology 62(6):1433 - 1440. MIRANDA, J. R., MOUILLOT, D., HERNANDEZ, D. F., LOPEZ, A. S., CHI, T. D. & PEREZ, L. A. 2005. Changes in four complementary facets of fish diversity in a tropical coastal lagoon after 18 years: a functional interpretation. Marine Ecology Progress Series 304:1-13. MITCHELL, C. E., TILMAN, D. & GROTH, J. V. 2002. Effects of grassland plant species diversity, abundance, and composition on foliar fungal disease. Ecology 83(6):1713 - 1726.
MOGUEL, P. & TOLEDO, V. M. 1999. Biodiversity Conservation in Traditional Coffee Systems of Mexico. Conservation Biology 13(1):11-21. NYSTRÖM, M. 2006. Redundancy and Response Diversity of Functional Groups: Implications for the Resilience of Coral Reefs. AMBIO: A Journal of the Human Environment 35(1):30-35. NYSTRÖM, M. & FOLKE, C. 2001. Spatial Resilience of Coral Reefs. Ecosystems 4(5):406417. PACALA, S. W. & KINZIG, A. P. 2002. Introduction to theory and the common ecosystem model. Pp. 169 -174 Kinzig, A. P., Pacala, S. W. & Tilman, D. In Functional Consequences of Biodiversity: Empirical Progress and Theoretical Extensions. Princeton Univiversity Press, Princeton, NJ. PAINE, R. T. 1969. A Note on Trophic Complexity and Community Stability. The American Naturalist 103(929):91-93. PERFECTO, I., MAS, A., DIETSCH, T. & VANDERMEER, J. 2003. Conservation of biodiversity in coffee agroecosystems: a tri-taxa comparison in southern Mexico. Biodiversity and Conservation 12(6):1239-1252. PERFECTO, I. & VANDERMEER, J. 2006. The effect of an ant-hemipteran mutualism on the coffee berry borer (Hypothenemus hampei) in southern Mexico. Agriculture, Ecosystems & Environment 117(2-3):218-221. PERFECTO, I., VANDERMEER, J., HANSON, P. & CARTÍN, V. 1997. Arthropod biodiversity loss and the transformation of a tropical agro-ecosystem. Biodiversity and Conservation 6(7):935-945. PERFECTO, I., VANDERMEER, J., MAS, A. & PINTO, L. S. 2005. Biodiversity, yield, and shade coffee certification. Ecological Economics 54(4):435-446. PHILPOTT, S. M. & ARMBRECHT, I. 2006. Biodiversity in tropical agroforests and the ecological role of ants and ant diversity in predatory function. Ecological Entomology 31(4):369-377. REITSMA, R., PARRISH, J. & MCLARNEY, W. 2001. The role of cacao plantations in maintaining forest avian diversity in southeastern Costa Rica. Agroforestry Systems 53(2):185-193. ROOM, P. M. 1972. The fauna of mistletoe Tapinanthus bangwensis growing on cocoa in Ghana: relationships between fauna and mistletoe. The Journal of Animal Ecology 41(3):611 - 621. ROWAN, R. 2004. Thermal adaptation in reef coral symbionts. Natur 430:742. SCHWARTZ, J. D. M., PALLIN, M. J., MICHENER, R. H., MBABAZI, D. & KAUFMAN, L. 2006. Effects of Nile perch, Lates niloticus, on functional and specific fish diversity in Uganda's Lake Kyoga system. African Journal of Ecology 44(2):145-156. SCHWARTZ, M. W., BRIGHAM, C. A., HOEKSEMA, J. D., LYONS, K. G., MILLS, M. H. & VAN MANTGEM, P. J. 2000. Linking biodiversity to ecosystem function: implications for conservation ecology. Oecologia 122(3):297-305. SLADE, E. M., MANN, D. J., VILLANUEVA, J. F. & LEWIS, O. T. 2007. Experimental evidence for the effects of dung beetle functional group richness and composition on ecosystem function in a tropical forest. Journal of Animal Ecology 76(6):1094-1104. SOLBRIG, O. T. 1991. A Research Agenda for Biodiversity. Pp. 123. From Genes to Ecosystems. Cambridge; International Union of Biological Science, Paris. SRIVASTAVA, D. S. & VELLEND, M. 2005. BIODIVERSITY-ECOSYSTEM FUNCTION RESEARCH: Is It Relevant to Conservation? Annual Review of Ecology, Evolution, and Systematics 36(1):267-294. TILMAN, D. 1982. Resource Competition and Community Structure. Princeton University Press, Princeton.
TILMAN, D., KNOPS, J., WEDIN, D., REICH, P., RITCHIE, M. & SIEMANN, E. 1997. The Influence of Functional Diversity and Composition on Ecosystem Processes. Science 277(5330):1300-1130. TILMAN, D., LEHMAN, C. L. & BRISTOW, C. E. 1998. Diversity-stability relationships: statistical inevitability or ecological consequence? The American Naturalist 151(3):277 282. VAN MELE, P. & VAN LENTEREN, J. C. 2002. Survey of current crop management practices in a mixed-ricefield landscape, Mekong Delta, Vietnam - potential of habitat manipulation for improved control of citrus leafminer and citrus red mite. Agriculture, Ecosystems and Environment 88:35-48. VANDERMEER, J., PERFECTO, I., IBARRA NUĂ‘EZ, G., PHILLPOTT, S. & GARCIA BALLINAS, A. 2002. Ants (Azteca sp.) as potential biological control agents in shade coffee production in Chiapas, Mexico. Agroforestry Systems 56(3):271-276. WALKER, B., KINZIG, A. & LANGRIDGE, J. 1998. Plant Attribute Diversity, Resilience, and Ecosystem Function: The Nature and Significance of Dominant and Minor Species. Ecosystems 2:95-113. WAY, M. J. & KHOO, K. C. 1992. Role of Ants in Pest Management. Annual Review of Entomology 37(1):479-503. YACHI, S. & LOREAU, M. 1999. Biodiversity and ecosystem productivity in a fluctuating environment: The insurance hypothesis. Proceedings of the National Academy of Sciences 96(4):1463-1468.