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== ''' A review of biodiversity-ecosystem function research ''' ==
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== Introduction ==
  
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In recent years, the recognition that species may play important roles in ecosystems and the rapidly emerging interest in the biodiversity conservation have prompted ecologists to ask new questions on the relationships between `diversity' and `ecosystem function' (for example, Walker, 1992<ref>Walker, B.H., 1992. Biodiversity and ecological redundancy. Conserv. Biol. 6: 18-23.</ref>; Schultze and Mooney, 1993; Jones and Lawton, 1995; Johnson et al., 1996).
  
Human needs and actions have, and will continue to, extensively alter ecosystems and biodiversity on a global scale [1]. Predictions of changes in biodiversity, not only in marine, but also terrestrial and freshwater ecosystems [2], have raised substantial concern over the consequences of biodiversity loss on ecosystem processes and function, which subsequently affect the provision of ecosystem goods and services, and ultimately affect human well-being [3].
 
 
Since the early 1990’s a portfolio of evidence obtained from the development of theory, laboratory experiments, field experiments and observational studies has shown that, irrespective of the system under study, increasing biodiversity tends to have positive effects on ecosystem properties, although the pattern of response may vary depending on the ecosystem and species investigated.
 
  
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== Why it is important? ==
  
[[Image:Flow diagramm.jpg|thumb|right|400px|Figure 1: Summary of research approaches adopted to address the relationship between biodiversity and ecosystem function in the peer-reviewed scientific literature. Yellow boxes represent theoretical studies, green boxes represent experimental manipulations of species diversity, red boxes represent the observational studies and the purple boxes represent those studies in which BEF concepts are linked to, or applied, in the real world.(modified from [7])]]
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One reason for the interest in the functional role of biodiversity (rather than structural) in ecosystems is that society might be more likely to take action to preserve biodiversity if it could be shown that there was some direct economic gain by doing it (Bengtsson, 1998). 
 +
Over the last fifteen years, an increasing number of studies have focused on biodiversity. This is principally because the world’s flora and fauna are disappearing at rates greater than during historical mass extinction events (Chapin et al, 2001). As recently suggested by Thomas et al. (2004), there is an 18 to 35% risk of species-level extinction resulting from climate changes by the year 2050. Moreover, other processes, for example, agricultural expansion in response to an increasing demand for food, have a negative impact on biodiversity as a result of habitat destruction (Tilman et al., 2001; Humbert and Dorigo, 2005).
  
== Emergence of a new paradigm ==
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Biodiversity and Ecosystem function are central to both community and ecosystems ecology and need to be understood to predict, for example, how communities and ecosystems respond to environmental change (Bengtsson, 1998) and on understanding how declining diversity influences ecosystem services on which humans depend (Duffy, 2003).  
 
The central thesis that guided early community ecology saw patterns in the distribution and abundance of species merely as an expression of the abiotic (chemical and physical) and biotic (species interactions) components of the environment, giving a predictive understanding of species distribution and abundance within ecosystems [4]. In the early 1990’s, however, an increasing number of ecologists began to challenge this view and, instead, started to examine, how ecosystem properties are mediated by the biota [5]. A wide range of hypotheses were developed describing the form of the biodiversity-ecosystem function relationship and which collectively formed a framework within which this relationship could be tested experimentally [6].  
 
  
In a series of phases, biodiversity-ecosystem function research has steadily improved to make experimental designs and model predictions more realistic (Figure 1). The timing of publications from each phase shows that, although different approaches have been used within the biodiversity-ecosystem function framework almost since the first influential paper was published in 1994 [8], there has been a general trend for experimental and theoretical studies to incorporate more natural environmental complexities (Reality filter, Figure 1).
 
  
== Phase 1 - Laboratory experiments ==
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== Research on Ecosystem Functioning ==
Initially, biodiversity-ecosystem function relationships were investigated by manipulating biodiversity under controlled conditions in the laboratory (Figure 2a, b) [12, 13]. In such studies, simple, single species communities are randomly assembled and their effects on ecosystem function determined. Then diversity is increased by constructing multi-species assemblages comprising the single species that have already been characterised and the effects of these multi-species assemblages on ecosystem function is determined (Figure 2c). If the observed response of the multi-specific assemblages differs from the response predicted by summation of the single species responses, then it is concluded that diversity has had an effect.
 
  
[[Image:Figure 2a.jpg|thumb|left|300px|Figure 2a: Mesocosms used to investigate the effect of invertebrate biodiversity loss on sediment nutrient release, e.g. 9.]]
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[[A review of biodiversity-ecosystem function research|Research on Biodiversity - Ecosystem Functioning]] (the BEF agenda) has stimulated a new and highly productive intercourse between population, community, ecosystem, and conservation ecology (Kinzig et al. 2002; Loreau et al. 2002; Duffy, 2003).
[[Image:Figure 2b.jpg|thumb|300px|none|Figure 2b: Mesocosms used to investigate the effect of biodiversity loss on bioturbation and nutrient release, e.g. 10, 36.]] 
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Most experimental evidence for biodiversity effects on ecosystem functioning has come from terrestrial ecosystems, particularly grasslands (Naeem et al. 1994, Tilmann et al. 1997a, Hector et al. 1999, Schmid et al. 2001; Giller et al., 2004). These studies have shown that changing biodiversity in natural ecosystems is likely to have much more complicated impacts on ecosystem functioning than predicted from changes in plant diversity alone (Duffy, 2003). For example in trophic levels of plant communities, as diversity is lost from a system, impacts will also depend from the loss of predators which will evoke change in the structure of all trophic levels (Hairston et al. 1960; Power 1990; Estes et al. 1998; Duffy, 2003).  
  
 +
The mosaic of habitat patches in aquatic systems often is more spatially compact than in terrestrial environments, presenting more tractable experimental systems at the landscape scale (Schindler and Scheuerell 2002). Because each aquatic ecosystem is composed of multiple habitat types, assessing the effects of biodiversity changes on the functioning of aquatic ecosystems requires experimental designs that allow a scaling up from individual homogenous patches to large scale, often highly heterogeneous areas (Giller et al. 2004).
  
Although these experiments successfully articulated biodiversity-ecosystem function hypotheses and allowed for the unambiguous interpretation of cause-effect relationships, critics were quick to assert that such studies lack realism because they tend to only include a few species (representing only a subset of the total community), often only from one trophic level, and they invariably assume that species loss is random. In addition, these types of experiments attracted further criticism because ecosystem function is measured infrequently and in the absence of the appropriate environmental context, thus making applicability to the real world questionable [14].
+
The most influential empirical research on biodiversity-ecosystem functioning linkages has been the series of experiments manipulating diversity in grasslands (reviewed by Tilman et al. 2002) and in aquatic microbial microcosms (reviewed by Petchey et al. 2002). Typically these have tested how ecosystem-wide biomass accumulation or metabolic rates change along gradients of species richness achieved by randomly assembling experimental communities from a pool of species. The grassland experiments have manipulated plant species richness, and sometimes also
  
[[Image:Figure 2c.jpg|thumb|centre|400px|Figure 2c: Experimental design commonly adopted in biodiversity-ecosystem function experiments. Red boxes indicate treatments with one species (Single species), two species and three species (Species mixtures). Total biomass between treatments is kept constant.]]
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functional group richness. These studies have demonstrated significant positive correlations between species richness and plant biomass. Loreau et al. (2002) provide a global overview of concepts and debates concerning the relationships between biodiversity and ecosystem functioning (Humbert and Dorigo, 2005).
  
Despite debates over experimental designs and applicability, syntheses of the available evidence suggest that, irrespective of the system studied, increased biodiversity tends to have a positive effect on ecosystem properties, such as primary productivity, nutrient flux and decomposition, but the pattern of response varies depending on the ecosystem and species investigated [15, 16]. This variability between results suggests that ecosystems respond differently to biodiversity loss, as natural ecosystems are complex, open systems that are composed of interconnected gradients, patches and networks between, and within which, organisms move and interact [17]. The high degree of control required in experiments and the short time periods under which they are conducted means, however, that the environmental variation and biological interactions that occur in natural systems are largely controlled for. The inherent complexity of environmental systems resulted in increasing calls to incorporate more environmental realism into experimental designs [13, 18] as it is still unclear whether the currently observed patterns in the biodiversity-ecosystem function relationship will hold, for example, for realistic extinction scenarios, in multi-trophic communities, and over larger spatial and temporal scales (14, 19).
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It has been clearly established that ecosystem functioning depends both on biotic factors and/or processes (such as the diversity and functions of the species, and interactions between species) and abiotic factors (such as climate or geology). However, what relative contribution these factors make is still a central question in the debate about diversity and ecosystem functioning (Huston and McBride, 2002; Humbert and Dorigo, 2005).
  
== Phase 2 – Inclusion of a subset of environmental variation ==
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Species deletion stability can also be linked easily to removal experiments that address the consequences of species loss for ecosystem functioning (Thébault, et al. 2007). With a few exceptions, theoretical work on the direct impact of species loss has focused on the study of secondary extinctions but has not considered associated changes in ecosystem properties (see King and Pimm 1983, Petchey et al. 2004).  
[[Image: BEF_Seagrass.jpg|thumb|left|250px| Figure 3: Outdoor mesocosms used in experiments investigating the effects of gazer diversity on ecosystem fucntioning in seagrass beds, [20].]]
 
  
Diversity manipulations in outdoor mesocosms [Figure 3, 20] are subjected to natural fluctuations in light, temperature and rainfall and thus represent better analogues of natural systems than more controlled laboratory conditions.
 
  
In-situ experiments or species addition/removal field experiments have been particularly valuable for determining biodiversity effects under naturally fluctuating environmental conditions (Figure 1)[21-23]. These experiments largely involve synthetically assembled communities in plots, mesocosms (Figure 4) or mesh-bags, in which biodiversity is controlled, but natural fluctuations in, for example, tidal cycles, rainfall or temperature, occur (Figure 4c). Although such studies incorporate more environmental variation, the spatial and temporal scales of the experimental designs largely do not relate to the dynamics of the study organisms, particularly in terms of their size or generation time. One way to incorporate such population dynamics at small and large scales is to adapt elements of meta-population theory [24]. In such studies, the importance of dispersal, the mechanism that regulates and maintains species coexistence (especially in fragmented habitats), in mediating the biodiversity-ecosystem function relationship, is recognised. At the same time, theoretical models have been extended to include environmental fluctuations [25].
+
Many of the studies that dealt specifically with the mechanisms involved in the relationships between biodiversity and ecosystem functioning investigated the niche complementarity mechanism, stimulating both theoretical and experimental approaches (e.g., Naeem et al., 1994; Loreau, 1998). The sampling effect, difficult to distinguish from the niche complementarity, is defined as the greater likelihood of finding species with a strong impact on ecosystem functioning in highly diversified communities (e.g., Huston, 1997; Hector et al., 1999; Wardle, 1999). These are not either-or mechanisms, but may be viewed as concomitant processes (Naeem, 2002). Sampling effects are involved in community assembly, and thus in determining the number of phenotypic traits present in the community. Subsequently, this phenotypic diversity influences ecosystem processes through mechanisms that can be viewed as a continuum ranging from the selection of species with particular traits to complementarity among species with different traits (Loreau et al., 2001).  
  
[[Image: Figure 4a.jpg|thumb|left|150px|Figure 4a: Mesocosm design to hold invertebrate communities and algal resources. Blue grids represent fine mesh, allowing tidally associated environmental fluctutations in the mesocosms, [23]]]
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Mathematical modelling has also been used recently, to investigate the relationships between biodiversity and ecosystem stability. For example, McCann et al. (1998) have shown that weak to intermediate interaction strengths within food webs are important in promoting community persistence and stability (Humbert and Dorigo, 2005).
  
[[Image: Figure 4c.jpg|thumb|right|350px|Figure 4b: Mesocosms containing invertebrate communities and algal resources pushed 10 cm into the mud, [23].]]
 
  
Most biodiversity-ecosystem function studies, use only a subset of the natural community to construct gradients of diversity, often selectively choosing those species that are either highly abundant or that may give the greatest response [18]. A move away from random extinction scenarios has relied primarily on modelling approaches to predict the consequences of diversity loss [26, 27]. The derived scenarios use extinction drivers that are explicitly associated with trait-based extinction probabilities and thereby provide a more direct way of assessing the possibilities of ecosystem responses to biodiversity loss. Such modelling approaches have the potential for exploring the consequences of biodiversity loss at greater spatial and temporal scales by making use of large data sets which are already available for many marine and terrestrial areas [19].
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== Theories and Hypothesis ==
  
== Phase 3 – Incorporating environmental variation ==
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In a recent review, Naeem et al. (2002) proposed three hypotheses to account for biodiversity-ecosystem functioning:
Despite the fact that there has been a steady increase in complexity of experimental and theoretical investigations of the biodiversity-ecosystem function relationship, it is still not clear whether the same patterns observed in experimental systems are just as strong and clear in natural systems. Studies based on field observations (Figure 4), have given some insight into the biodiversity-ecosystem function relationship in natural systems [28-30], however such studies are generally correlative in nature, lack direct experimental control and do not allow for replication. In addition, the inability to control confounding factors restricts the determination of cause-effect relationships between biodiversity loss and ecosystem function [28-30]. Despite efforts to overcome such problems (e.g. selecting sites of similar abiotic conditions, 28; or collecting data on additional environmental variables, 29), results are still open to alternative interpretations and have subsequently not shown any consistent results.
 
  
This third phase (Figure 1) is currently in its infancy, however it heralds a point in biodiversity-ecosystem function research history where the discipline has matured and a full suite of evidence (i.e. theory, methodology, laboratory and field experiments, or field observations) exists, providing insight into the likely ecosystem consequences of biodiversity loss. The biodiversity-ecosystem function relationship has been found to vary depending on the relative contribution of dominant and minor species [9], environmental context [31, 32], density dependence and species interactions [21, 23, 33], but few studies have explicitly incorporated those structuring abiotic (environmental heterogeneity) and biotic (movement, dispersal) features that are key to species co-existence and vital for the maintenance of species diversity [34]. In addition, biodiversity effects on ecosystem properties are significantly weaker under less-well controlled conditions [15], suggesting that the effect of biodiversity on ecosystem properties may be masked by abiotic factors in natural systems [35]. At present there is insufficient empirical evidence to determine the modifying effects of environmental factors, such as nutrient concentration, heterogeneity or elevated CO2 on biodiversity and community dynamics and, subsequently, ecosystem properties [15]. Even fewer attempts have been made, however, to establish and distinguish the relative importance of biodiversity and environmental factors in modifying ecosystem properties. Thus, a main challenge for the biodiversity-ecosystem function community is to demonstrate whether the observed importance of biodiversity in controlled experimental systems also persists in natural systems.
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*The first hypothesis is that species are primarily redundant, which means that one species can partially replace another. Many species have the same function, and the loss of one species can therefore be offset by some other species.  
 +
*The second hypothesis is that species are essentially singular, and make unique contributions to ecosystem functioning. The loss or gain of species (generally referred to as Keystone or Key species) therefore has a measurable impact on ecosystem functioning.
 +
*The third hypothesis is that species impacts are context dependent such that the impact of the loss or gain of a species on ecosystem functioning is idiosyncratic and unpredictable.  
 +
What happens, will depend on the local conditions under which the species extinction or addition occurs (Humbert and Dorigo, 2005)
  
== See also ==
 
Further detailed reviews on the relationship between biodiversity and ecosystem function include:
 
  
Balvanera, P., Pfisterer, A.B., Buchmann, N., He, J.S., Nakashizuka, T., Raffaelli, D. & Schmid, B. (2006). Quantifying the evidence for biodiversity effects on ecosystem functioning and services. ''Ecology Letters'' 9: 1146-1156.
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== How do we measure Ecosystem Functioning? ==
 +
 
 +
 
 +
Describing or measuring ecosystem functioning is difficult, as it encompasses a number of phenomena (Hooper et al., 2005). The overall functioning of an ecosystem is complex and involves many factors relating to the chemical, physical and biological components of the system. The way in which differences between species affect diversity-function relationships can be very complex (Lawton et al., 1998; Ricotta, 2005).  
 +
 
 +
Functional diversity (FD), i.e. the diversity and range of functional traits possessed by the biota of an ecosystem (Wright et al., 2006) or else defined by Tilman (2001) as ‘‘those components of biodiversity that influence how an ecosystem operates or functions’’ (Ricotta, 2005), is likely to be the component of biodiversity most relevant to the functioning of the ecosystems (Hooper et al., 2002; 2005; Heemsbergen et al., 2004), even though, there is no clear relationship demonstrated between species diversity and ecosystem functioning (Somerfield et al., 2008).
  
Cardinale, B.J., Srivastava, D.S., Duffy, J.E., Wright, J.P., Downing, A.L., Sankaran, M., & Jouseau, C. (2006). Effects of biodiversity on the functioning of trophic groups and ecosystems. ''Nature'' 443: 989-992.
+
Whereas traditional diversity indices focus on species richness (Jiguet et al. 2005), rarity (Schmera 2003) or the uncertainty of predicting species identity from abundance data (Magurran 1988), functional diversity formulae are used to measure ‘‘those components of biodiversity that influence how an ecosystem operates or functions’’ (Tilman et al. 1997; Schmera, Erös  and Podani, in press). Functional Diversity relates the number, type and distribution of functions performed by organisms within an ecosystem (Diaz & Cabido, 2001). It incorporates interactions between organisms and their environment into a concept that can portray ecosystem level structure in marine environments (Bremner et al., 2003) and conjectures to be useful in predicting the consequences of changes in species richness and composition, or biodiversity in general, on ecosystem properties (Somerfield et al., 2008).
  
Covich, A.P., Austen, M.C., Barlöcher, F., Chauvet, E., Cardinale, B.J., Biles, C.L., Inchausti, P., Dangles, O., Solan, M., Gessner, M.O., Statzner, B. & Moss, B. (2004). The role of Biodiversity in the functioning of freshwater and marine benthic ecosystems. ''BioScience'' 54: 767-775.  
+
Biodiversity can influence ecosystem functioning through changes in the amount of resource use complementary among species. Functional diversity is a measure of biodiversity that aims to quantify resource use complementarity and thereby explain and predict ecosystem functioning (Petchey, Hector and Gaston, 2004). Many studies have focused on calculating Functional Diversity, in order to measure the functioning of an Ecosystem. Methods and indices have been applied and tested on a long series of data concerning abiotic and biotic measures of fresh and sea water.  
  
Hooper, D.U., Chapin, F.S. III, Ewel, J.J., Hector, A., Inchausti, P., Lavorel, S., Lawton, J.H., Lodge, D.M., Loreau, M., Naeem, S., Schmid, B., Setala, H., Symstad, A.J., Vandermeer, J. & Wardle, D.A. (2005). Effects of biodiversity on ecosystem functioning: a consensus of current knowledge. ''Ecological Monographs'' 75: 3 - 35.
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Recent methods for calculating functional diversity of a community, include Functional Attribute Diversity (FAD), as used in a study of Australian rangelands by Walker et al. (1999), and Functional Diversity (FD) proposed more recently by Petchey & Gaston (2002) which is computed as the total branch length of the functional dendrogram that results from clustering the species in trait space (Ricotta, 2005). Trait variance, measured as the width of a trait distribution, has been proposed by Norberg (2004). Beyond the simple measurement of diversity, Mason et al. (2005) proposed also estimating functional richness, functional evenness and functional divergence, to enable descriptions of niche use and competitive interactions in communities. In order to take a functional approach and to use these new measures, however, we must have descriptors of the functional groups present in a community.
  
Hughes, J.B. & Petchey, O.L. (2001). Merging perspectives on biodiversity and ecosystem functioning. ''Trends in Ecology and Evolution'' 16: 222-223.
+
Most recently, based on the methodology proposed by the formers, Somerfield et al. (2008) defined average functional distinctness (X+, from χαρακτηριστικό, meaning a trait) simply as the average resemblance among species in a sample.  
 +
Incidentally, the same logic may be applied to Δ+ (Clarke and Warwick, 1998). Once branch lengths are defined between taxonomic levels, a matrix of resemblances (Euclidean distances) between species becomes implicit, and the index is the average resemblance between species. In the same study, the authors concluded that the type of information we get from the functional level is complementary to the information we take from the taxonomic level.  
  
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: 804-808.
 
  
Solan, M., Godbold, J.A., Symstad, A., Flynn, D.F.B. & Bunker, D. (2009). Biodiversity-ecosystem function research and biodiversity futures: early bird catches the worm or a day late and a dollar short? In: Biodiversity and human impacts. Ecological and societal implications. Naeem, S., Bunker, D.E., Hector, A., Loreau, M. & Perrings, C. (Eds.). Oxford University Press.
+
== How do we calculate Ecosystem Functioning in practice? ==
  
Stachowicz, J.J., Bruno, J.F. & Duffy, J.E. (2007). Understanding the effects of marine biodiversity on communities and ecosystems. ''Annual Review of Ecology, Evolution and Systematics'' 38: 739-766.
+
The categorization of species into [[functional groups]] can be done by simply assigning each species found in the assemblage to a given a priori defined functional group (Hector et al., 1999), or by standard multivariate clustering methods (Gitay & Noble, 1997; Deckers, Verheyen, Hermy, & Muys, 2004; Roscher et al., 2004)- see also [[Biological Trait Analysis]] (BTA).  
 +
To cluster species into functional groups, first, a set of [[functional traits]] thought to be of significance for ecosystem functioning is measured for each species obtaining an S x τ matrix of τ functional traits measured on S species (Petchey & Gaston, 2002). Next, the trait matrix is converted into a distance matrix Δ the elements dij of which embody the functional distances between the ith and the jth species such that dii=0 and dij =dji for any i≠j. Finally, the distance matrix is clustered with standard multivariate methods to separate species from different functional groups (Ricotta, 2005).  
 +
Generally, regardless of the proposed index, in most cases the information available for computing the FD of a given species assemblage is the set of pair wise species functional distances dij of a Δ matrix (Ricotta, 2005).
  
The influence of the lugworm (''Arenicola marina'') on biodiversity and ecosystem functioning in an intertidal mudflat [http://www.marbef.org/outreach/newsletter.php]<p>
 
  
==References==
+
== References ==
 
<references/>
 
<references/>
  
<br>
+
 
 +
== See also ==
 +
 
 
{{author
 
{{author
|AuthorID=19133
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|AuthorID=
|AuthorFullName=Godbold, Jasmin
+
|AuthorFullName=Vassiliki, Markantonatou |AuthorName=Markantonatou}}
|AuthorName=Jasmin}}
 

Revision as of 17:39, 9 March 2009

Introduction

In recent years, the recognition that species may play important roles in ecosystems and the rapidly emerging interest in the biodiversity conservation have prompted ecologists to ask new questions on the relationships between `diversity' and `ecosystem function' (for example, Walker, 1992[1]; Schultze and Mooney, 1993; Jones and Lawton, 1995; Johnson et al., 1996).


Why it is important?

One reason for the interest in the functional role of biodiversity (rather than structural) in ecosystems is that society might be more likely to take action to preserve biodiversity if it could be shown that there was some direct economic gain by doing it (Bengtsson, 1998). Over the last fifteen years, an increasing number of studies have focused on biodiversity. This is principally because the world’s flora and fauna are disappearing at rates greater than during historical mass extinction events (Chapin et al, 2001). As recently suggested by Thomas et al. (2004), there is an 18 to 35% risk of species-level extinction resulting from climate changes by the year 2050. Moreover, other processes, for example, agricultural expansion in response to an increasing demand for food, have a negative impact on biodiversity as a result of habitat destruction (Tilman et al., 2001; Humbert and Dorigo, 2005).

Biodiversity and Ecosystem function are central to both community and ecosystems ecology and need to be understood to predict, for example, how communities and ecosystems respond to environmental change (Bengtsson, 1998) and on understanding how declining diversity influences ecosystem services on which humans depend (Duffy, 2003).


Research on Ecosystem Functioning

Research on Biodiversity - Ecosystem Functioning (the BEF agenda) has stimulated a new and highly productive intercourse between population, community, ecosystem, and conservation ecology (Kinzig et al. 2002; Loreau et al. 2002; Duffy, 2003). Most experimental evidence for biodiversity effects on ecosystem functioning has come from terrestrial ecosystems, particularly grasslands (Naeem et al. 1994, Tilmann et al. 1997a, Hector et al. 1999, Schmid et al. 2001; Giller et al., 2004). These studies have shown that changing biodiversity in natural ecosystems is likely to have much more complicated impacts on ecosystem functioning than predicted from changes in plant diversity alone (Duffy, 2003). For example in trophic levels of plant communities, as diversity is lost from a system, impacts will also depend from the loss of predators which will evoke change in the structure of all trophic levels (Hairston et al. 1960; Power 1990; Estes et al. 1998; Duffy, 2003).

The mosaic of habitat patches in aquatic systems often is more spatially compact than in terrestrial environments, presenting more tractable experimental systems at the landscape scale (Schindler and Scheuerell 2002). Because each aquatic ecosystem is composed of multiple habitat types, assessing the effects of biodiversity changes on the functioning of aquatic ecosystems requires experimental designs that allow a scaling up from individual homogenous patches to large scale, often highly heterogeneous areas (Giller et al. 2004).

The most influential empirical research on biodiversity-ecosystem functioning linkages has been the series of experiments manipulating diversity in grasslands (reviewed by Tilman et al. 2002) and in aquatic microbial microcosms (reviewed by Petchey et al. 2002). Typically these have tested how ecosystem-wide biomass accumulation or metabolic rates change along gradients of species richness achieved by randomly assembling experimental communities from a pool of species. The grassland experiments have manipulated plant species richness, and sometimes also

functional group richness. These studies have demonstrated significant positive correlations between species richness and plant biomass. Loreau et al. (2002) provide a global overview of concepts and debates concerning the relationships between biodiversity and ecosystem functioning (Humbert and Dorigo, 2005).

It has been clearly established that ecosystem functioning depends both on biotic factors and/or processes (such as the diversity and functions of the species, and interactions between species) and abiotic factors (such as climate or geology). However, what relative contribution these factors make is still a central question in the debate about diversity and ecosystem functioning (Huston and McBride, 2002; Humbert and Dorigo, 2005).

Species deletion stability can also be linked easily to removal experiments that address the consequences of species loss for ecosystem functioning (Thébault, et al. 2007). With a few exceptions, theoretical work on the direct impact of species loss has focused on the study of secondary extinctions but has not considered associated changes in ecosystem properties (see King and Pimm 1983, Petchey et al. 2004).


Many of the studies that dealt specifically with the mechanisms involved in the relationships between biodiversity and ecosystem functioning investigated the niche complementarity mechanism, stimulating both theoretical and experimental approaches (e.g., Naeem et al., 1994; Loreau, 1998). The sampling effect, difficult to distinguish from the niche complementarity, is defined as the greater likelihood of finding species with a strong impact on ecosystem functioning in highly diversified communities (e.g., Huston, 1997; Hector et al., 1999; Wardle, 1999). These are not either-or mechanisms, but may be viewed as concomitant processes (Naeem, 2002). Sampling effects are involved in community assembly, and thus in determining the number of phenotypic traits present in the community. Subsequently, this phenotypic diversity influences ecosystem processes through mechanisms that can be viewed as a continuum ranging from the selection of species with particular traits to complementarity among species with different traits (Loreau et al., 2001).

Mathematical modelling has also been used recently, to investigate the relationships between biodiversity and ecosystem stability. For example, McCann et al. (1998) have shown that weak to intermediate interaction strengths within food webs are important in promoting community persistence and stability (Humbert and Dorigo, 2005).


Theories and Hypothesis

In a recent review, Naeem et al. (2002) proposed three hypotheses to account for biodiversity-ecosystem functioning:

  • The first hypothesis is that species are primarily redundant, which means that one species can partially replace another. Many species have the same function, and the loss of one species can therefore be offset by some other species.
  • The second hypothesis is that species are essentially singular, and make unique contributions to ecosystem functioning. The loss or gain of species (generally referred to as Keystone or Key species) therefore has a measurable impact on ecosystem functioning.
  • The third hypothesis is that species impacts are context dependent such that the impact of the loss or gain of a species on ecosystem functioning is idiosyncratic and unpredictable.

What happens, will depend on the local conditions under which the species extinction or addition occurs (Humbert and Dorigo, 2005)


How do we measure Ecosystem Functioning?

Describing or measuring ecosystem functioning is difficult, as it encompasses a number of phenomena (Hooper et al., 2005). The overall functioning of an ecosystem is complex and involves many factors relating to the chemical, physical and biological components of the system. The way in which differences between species affect diversity-function relationships can be very complex (Lawton et al., 1998; Ricotta, 2005).

Functional diversity (FD), i.e. the diversity and range of functional traits possessed by the biota of an ecosystem (Wright et al., 2006) or else defined by Tilman (2001) as ‘‘those components of biodiversity that influence how an ecosystem operates or functions’’ (Ricotta, 2005), is likely to be the component of biodiversity most relevant to the functioning of the ecosystems (Hooper et al., 2002; 2005; Heemsbergen et al., 2004), even though, there is no clear relationship demonstrated between species diversity and ecosystem functioning (Somerfield et al., 2008).

Whereas traditional diversity indices focus on species richness (Jiguet et al. 2005), rarity (Schmera 2003) or the uncertainty of predicting species identity from abundance data (Magurran 1988), functional diversity formulae are used to measure ‘‘those components of biodiversity that influence how an ecosystem operates or functions’’ (Tilman et al. 1997; Schmera, Erös and Podani, in press). Functional Diversity relates the number, type and distribution of functions performed by organisms within an ecosystem (Diaz & Cabido, 2001). It incorporates interactions between organisms and their environment into a concept that can portray ecosystem level structure in marine environments (Bremner et al., 2003) and conjectures to be useful in predicting the consequences of changes in species richness and composition, or biodiversity in general, on ecosystem properties (Somerfield et al., 2008).

Biodiversity can influence ecosystem functioning through changes in the amount of resource use complementary among species. Functional diversity is a measure of biodiversity that aims to quantify resource use complementarity and thereby explain and predict ecosystem functioning (Petchey, Hector and Gaston, 2004). Many studies have focused on calculating Functional Diversity, in order to measure the functioning of an Ecosystem. Methods and indices have been applied and tested on a long series of data concerning abiotic and biotic measures of fresh and sea water.

Recent methods for calculating functional diversity of a community, include Functional Attribute Diversity (FAD), as used in a study of Australian rangelands by Walker et al. (1999), and Functional Diversity (FD) proposed more recently by Petchey & Gaston (2002) which is computed as the total branch length of the functional dendrogram that results from clustering the species in trait space (Ricotta, 2005). Trait variance, measured as the width of a trait distribution, has been proposed by Norberg (2004). Beyond the simple measurement of diversity, Mason et al. (2005) proposed also estimating functional richness, functional evenness and functional divergence, to enable descriptions of niche use and competitive interactions in communities. In order to take a functional approach and to use these new measures, however, we must have descriptors of the functional groups present in a community.

Most recently, based on the methodology proposed by the formers, Somerfield et al. (2008) defined average functional distinctness (X+, from χαρακτηριστικό, meaning a trait) simply as the average resemblance among species in a sample. Incidentally, the same logic may be applied to Δ+ (Clarke and Warwick, 1998). Once branch lengths are defined between taxonomic levels, a matrix of resemblances (Euclidean distances) between species becomes implicit, and the index is the average resemblance between species. In the same study, the authors concluded that the type of information we get from the functional level is complementary to the information we take from the taxonomic level.


How do we calculate Ecosystem Functioning in practice?

The categorization of species into functional groups can be done by simply assigning each species found in the assemblage to a given a priori defined functional group (Hector et al., 1999), or by standard multivariate clustering methods (Gitay & Noble, 1997; Deckers, Verheyen, Hermy, & Muys, 2004; Roscher et al., 2004)- see also Biological Trait Analysis (BTA). To cluster species into functional groups, first, a set of functional traits thought to be of significance for ecosystem functioning is measured for each species obtaining an S x τ matrix of τ functional traits measured on S species (Petchey & Gaston, 2002). Next, the trait matrix is converted into a distance matrix Δ the elements dij of which embody the functional distances between the ith and the jth species such that dii=0 and dij =dji for any i≠j. Finally, the distance matrix is clustered with standard multivariate methods to separate species from different functional groups (Ricotta, 2005). Generally, regardless of the proposed index, in most cases the information available for computing the FD of a given species assemblage is the set of pair wise species functional distances dij of a Δ matrix (Ricotta, 2005).


References

  1. Walker, B.H., 1992. Biodiversity and ecological redundancy. Conserv. Biol. 6: 18-23.


See also

The main author of this article is Vassiliki, Markantonatou
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Citation: Vassiliki, Markantonatou (2009): Biodiversity and Ecosystem function. Available from http://www.coastalwiki.org/wiki/Biodiversity_and_Ecosystem_function [accessed on 28-03-2024]