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Internat. Rev. Hydrobiol. 93 2008 4–5 550–564 DOI: 10.1002/iroh.200711022 LARS GAMFELDT *, 1, 2 and HELMUT HILLEBRAND 1 1 Institute for Botany, University of Cologne, Gyrhofstraße 15, D-50931 Köln, Germany; e-mail: [email protected], [email protected] 2 Marine Biological Association of the United Kingdom, The Laboratory, Citadel Hill, Plymouth PL1 2PB, UK Review Paper Biodiversity Effects on Aquatic Ecosystem Functioning – Maturation of a New Paradigm key words: paradigm shift, functions, freshwater, marine Abstract Starting with the publication of some influential studies in the early 1990’s, the topic of biodiversity and ecosystem functioning has emerged as a major field within ecological research. Within this framework, the diversity of genotypes, species and functional groups are considered as explanatory variables of ecosystem functions rather than response variables of factors such as productivity and disturbance. Biodiversity-ecosystem functioning research has received considerable attention, and new publications are emerging at a high pace. Both the validity of experimental approaches and the way the results may be extrapolated to natural systems have, however, been widely discussed. The width of the debate regarding whether or not biodiversity is important for ecosystem functioning have encouraged many scientists to refine both experiments and theory, as well as develop novel methods to analyse the relationship between diversity and functioning. Aquatic ecologists have contributed greatly to the evolution of ideas and concepts within the field. In this review, we discuss how the paradigm that biodiversity is an important factor for the functioning of aquatic ecosystems is currently maturing with more realistic studies embracing both new and innovative approaches. We also suggest fruitful areas for future research. Each of us is trapped in a place, a time, and a circumstance, and our attempts to use our minds to transcend those boundaries are, more often than not, ineffective. From the book “Stumbling on happiness” by DANIEL GILBERT 1. Introduction The ability of scientists in general to think outside the box, to transcend traditional boundaries, is often limited. This holds true for ecologists and their fascination in the tremendous diversity of life inhabiting our globe. With a few exceptions, people have often been so mesmerized by the causes of biological diversity, that they have forgotten to ask themselves in what ways biodiversity might itself be important. Indeed, some of the most important issues within the field of ecology has been to describe the “when, where, and how” of biological diversity. Historically, the dominant puzzling question has been which factors are driving the striking and diverse patterns of species abundances observed in nature (HUTCHINSON, 1959). Classic examples of explored factors include competition (GAUSE, 1934), disturbance (DAYTON, 1971), environmental fluctuations (HUTCHINSON, 1961), and predation (PAINE, 1966). * Corresponding author © 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim 1434-2944/08/4-510-0550 Biodiversity Effects on Ecosystem Functioning 551 The role of diversity of organisms for ecosystem processes such as biomass accumulation (e.g., CARLANDER, 1955), and stability (e.g., MACARTHUR, 1955) has attracted the attention of scientists for a long time. DARWIN (1859) discussed that “… where they [animals and plants] come into the closest competition with each other, the advantages of diversification of structure, with the accompanying differences of habit and constitution, determine that the inhabitants, which thus jostle each other most closely, shall, as a general rule, belong to what we call different genera and orders.” (brackets inserted by the authors), and that “… if a plot of ground be sown with several distinct genera of grasses, a greater number of plants and a greater weight of dry herbage can (thus) be raised” (DARWIN, 1859). GAUSE (1934) used experiments to show that species with similar requirements tend not to co-exist because one species will be better at using the available resources than the other one. Following this principle, HARPER (1967) concluded that natural assemblages of mixed species should occupy different niches, and that complex ecosystems would be more efficient than simple ones in using environmental resources. In a book about the causes and consequences of species loss (EHRLICH and EHRLICH, 1982), the authors discussed that disappearing species may be much like losing the rivets of an aeroplane, because species have unique roles. When too many rivets are lost, the plane (or the ecosystem) will stop functioning. Numerous researchers have over the years addressed the relationship between single species or whole assemblages on ecosystem processes. Most have not, however, considered the role of biological diversity per se (RAFFAELLI et al., 2005). Mainly motivated by global change and rapid expansion of the human population (PIMM et al., 1995; VITOUSEK et al., 1997), the last two decades have seen an increasing interest in the consequences of biodiversity loss. A book summarising data and theory that could lend support to the hypothesis that biodiversity may be important to ecosystem functioning was published in 1993 (SCHULZE and MOONEY, 1993). Since then, ecologists have broadened their horizon to explicitly consider the functional consequences of biodiversity by viewing diversity as the independent variable. They have, if you will, transcended traditional boundaries to also include the functional consequences of altered levels of biodiversity. When, where, and how is biodiversity driving processes in nature? 2. The New Paradigm Most early work focused on the effects of biodiversity loss within one trophic level, and the bulk of these studies concerned terrestrial plants (TILMAN et al., 1996; HOOPER and VITOUSEK, 1997; HECTOR et al., 1999; MULDER et al., 2001). The conclusion from these experiments was that species and functional richness affect biomass accumulation, used as a proxy for productivity, and stability of the examined communities. Other studies inferred the same using multitrophic protist microcosms (MCGRADY-STEED et al., 1997; NAEEM and LI, 1997; PETCHEY et al., 1999). Another set of experiments considered the role of biodiversity in mediating invasion by alien species. By doing so, they revisited the classic hypothesis that more species-rich communities should also be more resistant to invasion (ELTON, 1958). STACHOWICZ et al. (1999) found that more diverse assemblages of sessile marine invertebrates were less invaded than were poorer counterparts. The authors contributed this to more complete use of settling space (a limiting resource in this system) over time in diverse assemblages. On the other hand, although diverse assembled plant communities appeared to be resistant to invasion in a riparian system, naturally diverse communities were the ones with the highest rate of invaders (LEVINE, 2000). It was concluded that other factors co-varying with diversity change in nature (e.g., propagule supply) may override the effects of diversity. Even though the biodiversity-ecosystem functioning experiments have provided new insights into community and ecosystem ecology, the new approach has not come without © 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.revhydro.com 552 L. GAMFELDT and H. HILLEBRAND friction. Many have questioned the validity of the biodiversity ecosystem function experiments (e.g., HUSTON and MCBRIDE, 2002), and have claimed that although biodiversity can be shown experimentally to be important for various processes (SPEHN et al., 2005), such patterns remain elusive in the real world (GRIME, 1997; GRACE et al., 2007). The tremendous debate concerning issues of biodiversity effects on functioning (KAISER, 2000), the attention the early studies received, and the number of studies emerging at an increasing pace suggests that biodiversity and ecosystem functioning is a rapidly evolving paradigm (Fig. 1) (NAEEM, 2002; but see PAINE, 2002). The paradigm of diversity and functioning challenges the idea (or central tenet if you will) that diversity is merely a function of extrinsic structuring forces (Fig. 1a). In doing so, it has stirred up conflicts about many fundamental issues in ecology: the relative strength of abiotic and biotic factors, the strength of observational vs. experimental approaches, the relative importance of taxonomic vs. functional diversity, and the question whether diversity begets stability or not. It has also forced scientists to integrate two previously largely separated subfields of ecology: community (studying the biotic components of ecosystems and their interactions) and ecosystem (taking a more holistic approach including material and energy flow) ecology. Obviously, biodiversity is neither a passive result of other structuring agents nor the sole vector explaining ecosystem processes. As is often the case, the truth lies somewhere in between with feedback loops between abiotic and biotic factors (Fig. 1b). All the same, if the paradigm of biodiversity and functioning receives a) Ecosystem properties and processes Diversity Regional species pool b) Metacommunity dynamics 3 Coexistence mechanisms 6 Multifunctionality Evenness 1 More ecosystem types Ecosystem properties and processes Species richness Landscape structure Trophic structure Habitat diversity Multitrophic assemblages 4 5 2 Genotypic, species, and functional diversity More functions Figure 1. Diagram representing the causality between ecosystem processes and biodiversity in the classical paradigm (a), and the paradigm shift of including species richness as a driver (as well as a consequence) of ecosystem properties and processes (b). Numbered points (1–6) refer to aspects representing the maturation of this paradigm shift towards higher generality and realism. Each point is further explained in the text. © 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.revhydro.com 553 Biodiversity Effects on Ecosystem Functioning enough support from scientists as well as from experiments and observational studies, it has the potential to evolve into a new central tenet in ecology. A common criticism of biodiversity-ecosystem functioning experiments has been that most of them have used artificially assembled communities with random species loss. The results from such studies, it has been argued, are unlikely to provide us with any useful information about the functioning of real ecosystems (HUSTON and MCBRIDE, 2002). Such claims seem to get good support from observations that many of the world’s aquatic ecosystems that are dominated by one or a few species (e.g., salt marshes, seagrass beds, kelp forests) still experience high standing biomass and/or high productivity. They are also supported by recent meta-analyses that find that individual species are as important for ecosystem processes as are mixtures of species (CARDINALE et al., 2006; GRACE et al., 2007). Another metaanalysis, however, shows that positive effects of biodiversity such as niche partitioning and positive interactions seem to dominate over sampling effects, i.e., the effects of individual species or combinations (CARDINALE et al., 2007), and that these effects grow stronger with time. Furthermore, analyses of data on marine biodiversity on both regional and global scales reveal that large-scale patterns of changes in biodiversity and ecosystem services correspond to those of experiments on a small temporal and spatial scale (WORM et al., 2006; DANOVARO et al., 2008). Even though conclusions from these broad-scale studies are based on correlations from which it is difficult to infer causation, it seems that biodiversity has a role to play on both local and regional scales in the marine realm. We focus on aquatic ecosystems in this review as the effects of biodiversity on ecosystem functioning may differ between aquatic and terrestrial environments (COVICH et al., 2004; GILLER et al., 2004). Recent reviews have highlighted important differences in trophic structure (SHURIN et al., 2006) including the importance of herbivory (CEBRIAN and LARTIGUE, 2004) and trophic cascades and top-down control (SHURIN et al., 2002). The response of plant diversity to consumer presence and fertilization also differs between aquatic and terrestrial systems (HILLEBRAND et al., 2007). Even diversity itself differs between the realms, as marine systems may harbour less species, but much higher phylogenetic diversity than terrestrial systems (GILLER et al., 2004). Though diversity-functioning research lagged behind in aquatic systems in the beginning, considerable progress has been made over the last few years. The effects of plant diversity may be common and similar to those of terrestrial systems (DUARTE, 2000; BRUNO et al., 2005), and generally weaker than the effects of consumers (DUFFY, 2002). Enough evidence currently exists to suggest that the new paradigm is maturing with more realistic scenarios and experiments providing mechanistic understanding about the relationship between diversity and the functioning of aquatic ecosystems. 3. Maturation of the Biodiversity-Functioning Paradigm 3.1. Expanding the Scales of Inference: More Systems and Functions As in many emerging paradigms, the scope of studies on the effects of biodiversity on ecosystem functioning has been narrow, very much focused on few aspects of biodiversity and few ecosystem functions. Most studies used either species richness or a measure of functional richness (NAEEM et al., 1994; HOOPER and VITOUSEK, 1997; TILMAN et al., 1997) and the response variables were mostly either productivity and nutrient uptake (NAEEM et al., 1994; TILMAN et al., 1996) or a measure of stability (TILMAN and DOWNING, 1994; MCGRADYSTEED et al., 1997). This scope has become more general and more comprising through the evolution of the paradigm (Fig. 1b). Other aspects of biodiversity, e.g., evenness, have been incorporated (DANGLES and MALMQVIST, 2004; MULDER et al., 2004; HILLEBRAND et al., 2008), and studies have started to investigate how the diversity within species (genotypic variation) affects ecosystem processes (HUGHES and STACHOWICZ, 2004; GAMFELDT et al., © 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.revhydro.com 554 L. GAMFELDT and H. HILLEBRAND 2005a; REUSCH et al., 2005; GAMFELDT and KÄLLSTRÖM, 2007) (Fig. 1b, point 1). Most of all, the definition of ecosystem function have started to include other important ecosystem processes, acknowledging the complexity of ecosystem functioning (Fig. 1b, point 2). In addition to the one-trophic level approach of many early studies, aquatic ecologists proposed the importance of looking at multitrophic assemblages (DUFFY, 2002; GILLER et al., 2004) and analyzed diversity effects on consumption rates (transfer between trophic levels) (JONSSON and MALMQVIST, 2000; CARDINALE et al., 2002; DUFFY et al., 2003). Other ecosystem processes, which are now considered, comprise predation (BYRNES et al., 2006), rates of bioturbation (EMMERSON et al., 2001), or – in terrestrial systems-pollination (FONTAINE et al., 2006). For aquatic ecosystems, GILLER et al., (2004) listed more than 20 processes of relevance in ecosystems, which potentially are affected by diversity loss. 3.2. Spatial Scale, Coexistence and Metacommunities Most experiments and models founding the biodiversity-ecosystem functioning paradigm considered only local scales and local interactions of communities. However, there are good reasons to include a broader approach including the regional species pool and dispersal processes (GILLER et al., 2004). First, regional diversity is supposed to affect local diversity by altering the number of species potentially colonizing a certain site (RICKLEFS 1987; SHURIN et al., 2000; HILLEBRAND and BLENCKNER, 2002). Second, the loss of species in a local habitat may be transient and the ability of a system to persist may be more dependent on the rate of (re-)colonization rather than the number of species present. This includes also the notion that the community composition may be constrained by dispersal limitation (SHURIN et al., 2000; CACERES and SOLUK, 2002). Third, experiments with local manipulations of species richness, as they have been conducted in the biodiversity-ecosystem functioning framework, did not provide a stable coexistence mechanism for the high number of species involved. Thus, these experiments would have lost many, if not most, of the species over time. For these reasons, GILLER et al. (2004) proposed to conduct experiments manipulating both local and regional richness as well as the dispersal rate in order to add more realism to biodiversity-functioning research. The importance of regional processes has been embedded in the concept of metacommunities (LEIBOLD et al., 2004; HOLYOAK et al., 2005), which represent local communities connected via dispersal, e.g., lakes and ponds in a landscape or rock pools along the coast. Metacommunity dynamics can lead to stable coexistence by a variety of mechanisms, among these patch dynamics (coexistence along a colonization-competition trade-off), source-sink dynamics (coexistence by high reproduction in one habitat and mass effect of propagules in other locales of the metacommunity) or species sorting across a heterogeneous metacommunity (LEIBOLD et al., 2004). The role of dispersal and propagule supply is becoming an important part of biodiversityfunctioning research (FRANCE and DUFFY, 2006; NAESLUND and NORBERG, 2006; MATTHIESSEN et al., 2007; Fig. 1b, point 3). Two models have analyzed how metacommunity dynamics may change our understanding of the effects biodiversity loss. LOREAU et al. (2003) showed that dispersal rates controlled the number of coexisting species in a metacommunity, which then transfered into higher biomass production. Thus, they found higher richness in intermediate dispersal scenarios and a positive correlation between diversity and productivity. Also a second model concluded that an extension of scale may alter the outcome of biodiversityecosystem functioning studies and suggested to acknowledge the importance of the regional scale (CARDINALE et al., 2004). Whereas a number of studies have analyzed the importance of metacommunity dynamics for species coexistence (COTTENIE et al., 2003; KNEITEL and MILLER, 2003; CADOTTE and FUKAMI, 2005), few actually transferred this to biodiversity-ecosystem functioning research. © 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.revhydro.com 555 Biodiversity Effects on Ecosystem Functioning For rock pool systems, communities arising from a regional species pool had higher abundances of zooplankton compared to communities stemming from local pools, with cascading effects on the primary producer level (NAESLUND and NORBERG, 2006). MATTHIESSEN and HILLEBRAND (2006) showed that local diversity is maximized at intermediate dispersal rate, which leads to higher resource use efficiency and primary production. Although their experimental setup with marine microcosms differed from the model concepts of metacommunities, it seemed that providing a stable coexistence mechanism actually strengthened the main hypothesis of the new paradigm that ecosystem process rates increase with increasing diversity. Regional aspects may affect biodiversity-ecosystem functioning relationships not only by dispersal, but also by the structure of the landscape and the connectivity between different habitats (Fig. 1b, point 4). GILLER et al. (2004) suggested that certain processes may be enhanced rather by the diversity of habitats involved than by species richness. This may be especially true for aquatic ecosystems with strong exchange of organisms, material and energy across boundaries (benthic-pelagic coupling, riparian-zone – river). 3.3. Roles of Consumers and Interactions among Trophic Levels Aquatic research has played an important role in elucidating the potential importance of consumer diversity for ecosystem functioning. While documented extinctions of primary producers are rare, they appear to be more common for consumer species (DUFFY, 2003). For example, overexploitation of fish and shellfish in the oceans has resulted in dramatic shifts in many ocean ecosystems (DAYTON et al., 1998; PAULY et al., 1998; JACKSON et al., 2001; BELLWOOD et al., 2004), often with significant effects on vital ecosystem processes (BELLWOOD et al., 2003; WORM et al., 2006; MYERS et al., 2007). Accumulating evidence from aquatic studies suggest that the diversity of consumer species is indeed an important feature of aquatic ecosystems (Fig. 1b, point 5). A series of experiments on fresh-water insects showed that the richness and evenness of these insect larvae can determine process rates of leaf consumption (e.g., JONSSON and MALMQVIST, 2000; JONSSON et al., 2001; DANGLES and MALMQVIST, 2004; JONSSON, 2006), and that these rates may depend on a few dominant species and per capita responses by non-manipulated species (RUESINK and SRIVASTAVA, 2001). For another set of aquatic insect larvae, resource consumption of filter feeders increased more than expected based on single species performances due to interspecific facilitation (CARDINALE et al., 2002). WOJDAK and MITTELBACH (2007) took the innovative approach to include habitat use of pond grazers as the predictor of ecosystem properties. The explicit inclusion of habitat preference of organisms is likely to contribute significantly to our knowledge about effects of diversity (SCHMITZ, 2007). Results from seagrass systems indicate that a richness of marine herbivore mesograzers is important for a range of processes such as epiphyte overgrowth and secondary production (DUFFY et al., 2003). Similar results have been obtained for rocky shore communities where herbivore and predator complementarity in feeding strategies appears to be important for community structure (LOTZE and WORM, 2000; RÅBERG and KAUTSKY, 2007; GRIFFIN et al., 2008). Food chain length can have profound effects on the importance of grazer diversity. For example, in a seagrass system, grazer richness had effects on system processes only with a predator present (DUFFY et al., 2005). Indeed, three manipulations of predator richness suggest that losing predator diversity from marine systems can have cascading effects on the algal communities (BYRNES et al., 2006; BRUNO and O’CONNOR, 2005; ELLIS et al., 2007). The results of these experiments parallel those of observational studies on functional diversity and predator extinctions. In coastal marine systems, functional diversity may be strongly correlated with species diversity, suggesting that functional redundancy is low (MICHELI and HALPERN, 2005). This indicates that coastal habitats may be sensitive to © 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.revhydro.com 556 L. GAMFELDT and H. HILLEBRAND species loss. Concordantly, kelp forests of the Californian coast were found to be strongly top-down controlled rather than bottom-up (HALPERN et al., 2006, but see FOSTER et al., 2006). It should be noted that simultaneous diversity changes at multiple trophic levels can have interactive effects that may not be predicted by studying each trophic level in isolation (GAMFELDT et al., 2005b). Using the framework of DUFFY et al. (2007), both horizontal (e.g., among species diversity) and vertical (e.g., number of trophic levels) components must be considered simultaneously for a comprehensive picture of the effects biodiversity loss. In addition to rates of consumption, consumers indirectly affect ecosystem processes by altering nutrient cycling. Since species differ in habitat use and their requirements for nutrients, species loss can affect nutrient recycling as shown for both tropical fresh-water fish communities (MCINTYRE et al., 2007) and assemblages of sediment bioturbating invertebrates (EMMERSON et al., 2001; CALIMAN et al., 2007). 3.4. Multiple and Parallel Functions Within the extensive biodiversity-ecosystem functioning research over the past decade, most studies have used a single-function perspective, i.e., addressed the consequence of species loss on single process rates or properties. The use of single response variables as proxies for ecosystem functioning may ignore other important ecosystem processes (ROSENFELD, 2002; DUFFY et al., 2003). Equating single functions with overall functioning can be highly misleading, especially if research ultimately aims to provide knowledge and advice to management and conservation. Rather, individual functions and processes may be better viewed as components of overall ecosystem functioning writ large (Fig. 1b, point 6). The effect of multiple ecosystem functions on joint ecosystem functioning has been explicitly examined quantitatively in four studies. There are indeed more studies that have measured multiple functions (e.g., TILMAN et al., 1997; DUFFY et al., 2003; SPEHN et al., 2005) but they have not been considered jointly. Using plants and an index of relative resource use, HOOPER and VITOUSEK (1998) showed that the diversity of functional groups was important for the total depletion of soil nutrients over time, and HECTOR and BAGCHI (2007) found that plant species richness is more important for multifunctionality than for single functions. GAMFELDT (2006) and GAMFELDT et al. (2008) examined a broader range of systems and organisms that included marine seagrass and grazers, and reached similar conclusions: as more functions are considered, the importance of biodiversity for ecosystem functioning becomes more apparent. The combined results from these studies suggest that focusing on individual functions can often be highly misleading, because a high level of a single function does not equal overall ecosystem functioning. In contrast, it is possible that mixtures perform worse than the best constituting species for each individual function considered (unifunctional underyielding), but still experience multifunctional overyielding because the identity of the best monoculture species switches between functions. In many cases, only one ecosystem function is of interest and biodiversity may, or may not, be important for this function. Nonetheless, the more functions considered, the closer we get to an approximation of overall ecosystem functioning. 3.5. Complementary Approaches and Experimental Design Due to logistical constraints, experimental work has been limited in both space and time, and this limits the possibility to extrapolate results to the real world. There is thus a need to complement experiments and theoretical models with observational data on larger scales (NAEEM, 2006), as in WORM et al. (2006). New approaches must also embrace the fact that species loss is not random in response to stressors and exploitation. SOLAN et al. (2004) © 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.revhydro.com Biodiversity Effects on Ecosystem Functioning 557 modelled a set of extinction scenarios for soft-bottom invertebrates and showed that the magnitude of reduction in bioturbation depended on the covariation between species functional traits and their risk of extinction. Similar results were found for alternative extinction scenarios of fresh-water fishes and their effects on nutrient recycling (MCINTYRE et al., 2007). Combining field observations with experiments and modelling, BRACKEN et al. (2008) showed that realistic scenarios of species loss in a macroalgal community resulted in declines in nutrient uptake whereas random loss had no effect. Modelling the effects of biodiversity loss using existing data from observational studies together with scenarios of global change will be an important part of biodiversity-functioning research. In terms of experimental design, there is an increased awareness within the field that the design of most experiments to date have been such that it is hard to separate effects of diversity and density of species. This is not surprising news to people well acquainted with the literature on experimental design and the replacement series (JOLLIFFE, 2000). The majority of studies have used a replacement design to examine the role of biodiversity. This means that they have used the same total density (dtot) of organisms across diversity treatments so that n number of species within species mixtures have a density of dtot/n. This obviously confounds density with diversity, and it is hard to tell if observed effects are due to increased diversity or decreased intraspecific competition. An alternative approach, the additive design (e.g., CARDINALE et al., 2003), keeps the density of each species (dmono) constant so that species mixtures have a density of dmono · n. For a thorough understanding of the ways in which both density and diversity affect functions, a combination of substitutive and additive designs may be necessary (FIRBANK and WATKINSON, 1985; WEIS et al., 2007). Furthermore, the number of possible unique species (or functional group) combinations, and thus also the similarity among replicates, varies with diversity. This produces inherent variance reduction effects, and the possibility to interpret uncontrolled error effects as true diversity effects (HUSTON and MCBRIDE, 2002). BENEDETTI-CECCHI (2006) has proposed a design that disentangles the effects of density and diversity. Furthermore, in order to detect effects of consumer diversity on consumer production it is important that experiments are run long enough for competition among species to occur, so that realised densities are reached. Actual starting densities should then be less important for the outcome. The role of functional diversity is also an important aspect of biodiversity that deserves more attention (PETCHEY and GASTON, 2006). 4. Future Outlook Current evidence from experimental studies generally shows that ecosystem functioning can often be sustained by only a few species (CARDINALE et al., 2006). In many cases, a small subset of species may be sufficient, and the loss of a large fraction of biodiversity from systems may often have little effect on the magnitude of ecosystem functions (SOLAN et al., 2004). Furthermore, the effects of biodiversity may interact with, or be much weaker than, other factors such as disturbance, flow conditions and nutrient availability (CARDINALE et al., 2000; BILES et al., 2003; DZIALOWSKI and SMITH, 2008). Manipulating both biodiversity and other factors such as propagule and resource supply in a factorial design to evaluate the relative importance of these factors will be an important feature of future experiments. Nonetheless, as the paradigm that biodiversity affects aquatic ecosystem performance matures, we are receiving information from more realistic studies. Experiments are run over larger scales spanning longer time periods, include more complex food webs, and are complemented by a wider set of approaches. These include the study of relationships between species abundances and ecosystem processes over large scales in nature (NAEEM, 2006), the exploration of the consequences of non-random species extinctions (e.g., JONSSON et al., 2002; SOLAN et al., 2004; MCINTYRE et al., 2007), and meta-analyses of published experi© 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.revhydro.com 558 L. GAMFELDT and H. HILLEBRAND ments to gain new insights about the role of different mechanisms (CARDINALE et al., 2007). Additionally, by re-examining data from studies that were not initially designed to address the effect of biodiversity on functions (as in EMMERSON and HUXHAM, 2002), we can learn from the vast literature that already exist on e.g., the roles of different species in succession (PEARSON and ROSENBERG, 1978), nutrient cycling (SNELGROVE et al., 1997), and tolerance to changing water temperatures (SOUTHWARD et al., 1995). For a thorough understanding about how and when biodiversity can influence ecosystem functioning, however, future experimental studies must be designed to evaluate specific mechanisms, something that microcosm systems are well suited to do (CADOTTE et al., 2005). Another set of informative experiments involves removing one or a few target species from otherwise intact communities, and track the effects on ecosystem functioning (SCHIEL, 2006). Removal experiments are unfortunately still not widely used in biodiversity-ecosystem functioning studies. Even though the term biodiversity refers to the diversity at any hierarchical level (genes, species, ecosystems), studies addressing the importance of diversity at the higher levels of organization are still missing. This is unfortunate, since human activity and global change are likely to have large effects at the landscape level. Modelling approaches (e.g., SOLAN et al., 2004) should be used on existing data on whole ecosystems to evaluate the importance of landscape heterogeneity to overall ecosystem functioning. One pressing question relates to the role of consumer diversity in ecosystems. The exact roles of consumers in general, and predators in particular, are hard to predict as a consequence of indirect interactions, omnivory and cascading effects (DUFFY et al., 2007). Furthermore, the primary focus of consumer diversity has been on the effects of lower trophic levels, whereas there is good reason to expect that it will also affect the production of the consumers themselves (IVES et al., 2005; GAMFELDT et al., 2005b; WORM et al., 2006). It is more difficult to examine consumer production because experiments have to run long enough for consumers to grow and reproduce. A fruitful approach would be to remove consumer species in different combinations from natural systems and study if the lost biomass is replaced by other complementary species. The role of consumer diversity at multiple trophic levels and its influence on secondary production is an interesting area for future research. In species-poor systems, or environments dominated by one or a few species, it would be especially interesting to explore the relative importance and interaction of intra- and interspecific diversity. For example, the genetic diversity of habitat-forming seagrass may govern resistance to disturbance and disease, whereas the total diversity of associated primary producers and consumers is important for processes such as productivity and nutrient recycling. To the best of our knowledge, there is to date no study that has addressed the interaction between intra- and interspecific diversity and the effect on ecosystem functions. Another challenge is to map and characterise the range of relevant functions of notable importance in aquatic systems. Which functions are considered most important, how are functions correlated across organisms and ecosystems, and how are multiple functions (overall functioning) affected by species loss (GAMFELDT et al., 2008)? This is important if we are to be able to provide society and conservation management with guidance of why and when biodiversity may matter for the functioning of natural ecosystems (SRIVASTAVA and VELLEND, 2005; STACHOWICZ et al., 2007). Only by combining efforts on different scales and with different methods can we start to fully understand the consequences of biodiversity loss in real ecosystems. 5. Acknowledgements LG was supported by a postdoc grant (2006-1173) from the Swedish research council Formas, and we would like to thank RITA ADRIAN for inviting us to write this paper. 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