Climate change and freshwater biodiversity - University of Idaho

BIOLOGICAL REVIEWS

Cambridge Philosophical Society

39

Biol. Rev. (2009), 84, pp. 39–54. doi:10.1111/j.1469-185X.2008.00060.x

Climate change and freshwater biodiversity: detected patterns, future trends and adaptations in northern regions Jani Heino1,2*, Raimo Virkkala3 and Heikki Toivonen3 1

Finnish Environment Institute, Research Programme for Biodiversity, P.O. Box 413, FI-90014 University of Oulu, Finland Department of Biology, University of Oulu, P.O. Box 3000, FI-90014 University of Oulu, Finland 3 Finnish Environment Institute, Research Programme for Biodiversity, P.O. Box 140, FI-00251 Helsinki, Finland 2

(Received 23 May 2008; revised 7 October 2008; accepted 9 October 2008)

ABSTRACT Current rates of climate change are unprecedented, and biological responses to these changes have also been rapid at the levels of ecosystems, communities, and species. Most research on climate change effects on biodiversity has concentrated on the terrestrial realm, and considerable changes in terrestrial biodiversity and species’ distributions have already been detected in response to climate change. The studies that have considered organisms in the freshwater realm have also shown that freshwater biodiversity is highly vulnerable to climate change, with extinction rates and extirpations of freshwater species matching or exceeding those suggested for better-known terrestrial taxa. There is some evidence that freshwater species have exhibited range shifts in response to climate change in the last millennia, centuries, and decades. However, the effects are typically speciesspecific, with cold-water organisms being generally negatively affected and warm-water organisms positively affected. However, detected range shifts are based on findings from a relatively low number of taxonomic groups, samples from few freshwater ecosystems, and few regions. The lack of a wider knowledge hinders predictions of the responses of much of freshwater biodiversity to climate change and other major anthropogenic stressors. Due to the lack of detailed distributional information for most freshwater taxonomic groups and the absence of distribution-climate models, future studies should aim at furthering our knowledge about these aspects of the ecology of freshwater organisms. Such information is not only important with regard to the basic ecological issue of predicting the responses of freshwater species to climate variables, but also when assessing the applied issue of the capacity of protected areas to accommodate future changes in the distributions of freshwater species. This is a huge challenge, because most current protected areas have not been delineated based on the requirements of freshwater organisms. Thus, the requirements of freshwater organisms should be taken into account in the future delineation of protected areas and in the estimation of the degree to which protected areas accommodate freshwater biodiversity in the changing climate and associated environmental changes. Key words: climate change adaptation, conservation, range shifts, species distribution, thermal regimes. CONTENTS I. Introduction ...................................................................................................................................... (1) Projected patterns of climate change and freshwater ecosystems ............................................. (2) Status of freshwater biodiversity ................................................................................................. (3) Climate change, the environmental filters perspective and freshwater biodiversity ................. (4) Thermal guilds of freshwater organisms .................................................................................... II. Observed and predicted changes in species’ distributions ...............................................................

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* Address for correspondence: E-mail: [email protected] Biological Reviews 84 (2009) 39–54 Ó 2008 The Authors Journal compilation Ó 2008 Cambridge Philosophical Society

Jani Heino, Raimo Virkkala and Heikki Toivonen

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III.

IV.

V.

VI. VII. VIII.

(1) How to assess the effects of climate change on species’ ranges? .............................................. (2) Observed and predicted shifts in species’ distributions at large geographic scales .................. (3) Specific surveys of streams and lakes across latitudinal gradients ............................................ (4) Synthesis ...................................................................................................................................... Interactions between climate change and multiple anthropogenic stressors .................................. (1) Climate change and acidification ............................................................................................... (2) Climate change and eutrophication ........................................................................................... (3) Climate change and land cover alterations ............................................................................... (4) Climate change and exotic species ............................................................................................. Climate change adaptation, conservation and freshwater biodiversity ........................................... (1) Networks of protected areas ....................................................................................................... (2) Protection of large and environmentally heterogeneous areas ................................................. (3) Dispersal corridors ...................................................................................................................... (4) Habitat restoration and management ........................................................................................ (5) Management of the matrix between protected areas ................................................................ (6) Reintroduction of native species ................................................................................................. Key topics for future research .......................................................................................................... (1) Acquisition of comprehensive distributional data ...................................................................... (2) Long-term monitoring networks for examining temporal changes in biodiversity .................. (3) Assessment of species and habitats most threatened by and responsive to climate change .... (4) Assessment of the present protected area network .................................................................... (5) Estimation of dispersal corridors for freshwater species ............................................................ Conclusions ....................................................................................................................................... Acknowledgements ............................................................................................................................ References .........................................................................................................................................

I. INTRODUCTION Global change has been shown and predicted to have major effects on biodiversity at global, regional, and local scales. Although global change constitutes a number of different forms of anthropogenic impacts (Kappelle, Margret & Baas, 1999; Sala et al., 2000), including land use alterations, nitrogen deposition, and invasions of exotic species, much recent interest has been directed at climate change (Root et al., 2003; Parmesan, 2006). Although the Earth has experienced considerable climate changes in the past, the rate and magnitude of the recent and projected future changes are unprecedented (IPCC, 2001b). Furthermore, the effect of future climate change on biodiversity has been predicted to be unprecedented as well, with 15 – 37% of terrestrial species possibly becoming extinct due to climate change alone in the next 50 years (Thomas et al., 2004), and a similarly dark future has been suggested for freshwater species in the next few decades (Xenopoulos et al., 2005). Thus, it is not surprising that research on the effects of climate change on terrestrial, marine, and freshwater organisms and ecosystems has increased very rapidly in the last two decades. Despite this increase of research on the topic, we still lack a comprehensive understanding of climate change and predictive capability of its effects on biodiversity in various organism groups and ecosystems. Aquatic ecosystems are as vulnerable to global change as terrestrial and marine ecosystems. Sala et al. (2000) considered lentic (i.e. lakes and ponds) and lotic (i.e. streams and rivers) ecosystems to be most sensitive to land use change, exotic species, and climate change in a global-scale

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assessment. However, these drivers of change may vary among regions and latitudes, with aquatic ecosystems at high latitudes being more strongly threatened by climate change than by other drivers. Given that most reviews on the effect of climate change on lentic and lotic systems have concentrated on regions south from boreal regions (Magnuson et al., 1997; Moore et al., 1997; Poff, Brinson & Day, 2002; Mooij et al., 2005; Ryan & Ryan, 2006) or stressed mainly ecosystem processes (Carpenter et al., 1992; Rouse et al., 1997; Schindler, 1997; ACIA, 2005; Allan, Palmer & Poff, 2005, Wrona et al., 2006), the aim of the present article was to complement the existing information by reviewing current knowledge of climate change effects on species’ distribution patterns in boreal regions. Boreal was considered broadly in this context, including in addition to the boreal coniferous zone, also arctic and northern temperate regions north of approximately 50oN. Contrary to most other reviews on climate change effects on freshwater ecosystems, we will concentrate on describing the distributional changes of species at large scales, and combine this information with other major anthropogenic stressors on aquatic biodiversity (i.e. acidification, eutrophication, land cover changes, exotic species; Fig. 1), climate change adaptations, and conservation of freshwater biodiversity. We conclude this review by summarising the areas of research that are necessary for furthering our knowledge of the influences of climate change on freshwater biodiversity. We will not concentrate on shortterm climatic fluctuations (e.g. North Atlantic Oscillation and El Ninõ Southern Oscillation), as these have been given considerable previous attention (e.g. Blenckner & Hillebrand, 2002; Durance & Ormerod, 2007).

Biological Reviews 84 (2009) 39–54 Ó 2008 The Authors Journal compilation Ó 2008 Cambridge Philosophical Society

Climate change and freshwater biodiversity

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Increased levels of greenhouse gases in the atmosphere

Humans

Climate

L A and Eu cidi -use Al tro fica ch ien ph tio an sp ica n ges ec tio ies n

Biodiversity at regional and local scales

Changes

Changes

Ecosystem and habitat characteristics Viability of populations and metapopulations Species composition in local communities Species richness in local communities Species abundances in local communities Geographical distributions

Reduction of the availability and production of natural resources (e.g. game fish) Altered and diminished conservation value

es s im re g tu r e ra al pe ic m og te ol d dr se hy ea cr red In lte A

Ecosystems effects of altered community characteristics (e.g. increases in CO2 concentrations)

Increased global mean temperature Altered patterns of precipitation

Biodiversity loss

Fig. 1. Schematic diagram showing the relationships between climate change and other major anthropogenic influences and their effects on biodiversity. The two major factors resulting directly from climate change and the four major anthropogenic factors have both individual and interactive effects on biodiversity in freshwater ecosystems. Adapted from Kappelle et al. (1999).

(1) Projected patterns of climate change and freshwater ecosystems Climatic changes have been recurring in the history of the Earth, but the present rates of climate warming have not occurred previously in the last 1000 years (IPCC, 2001b). Global surface temperatures increased by 0.6oC during the 20th Century, and it has been predicted that mean global surface temperatures will rise by 1.4 to 5.8oC in the next 100 years or so, mainly depending on the amount of carbon dioxide emission from anthropogenic sources (IPCC, 2001b). Increases in carbon dioxide and other greenhouse gases have been claimed to be responsible for the recent increases in global temperatures, and if the anthropogenic emissions of these gases are not limited, then global temperatures may rise even more. Furthermore, although rise in mean global temperature is suggested to be considerable, most land areas are warming more rapidly than the global average, and there also will be clear regional

differences (IPCC, 2002). In this regard, high-latitude regions may be expected to experience even more dramatic warming, with up to 10oC increases in temperature projected for high-latitude regions (Flato & Boer, 2001; ACIA, 2005). In addition to direct effects on water temperature, changes in growing season and ice cover are likely to have profound influences on freshwater organisms (e.g. Magnuson et al., 2000). For example, increased length of growing season and decreased period of ice cover may have both direct (e.g. increased macrophyte production) and indirect effects (e.g. elimination of predation-sensitive fish with the invasion of low-oxygen sensitive predators) on freshwater organisms (e.g. Tonn, 1990). In addition to climate warming, climate change effects are related to precipitation patterns. Although it has been projected that there will be considerable among-region differences in precipitation patterns, with both regional increases and decreases, globally averaged annual precipitation has been predicted to increase during the 21st Century (IPCC, 2002). In high-latitude regions, precipitation is likely to increase in both summer and winter, with increased year-to-year variation in precipitation (IPCC, 2002). Altered rainfall patterns in terms of more unpredictable heavy rains are also likely to increase the probability of flash floods in boreal streams (IPCC, 2001b). However, more general reductions in summer flow may also occur following earlier snowmelt and increased evapotranspiration at higher temperatures (Poff et al., 2002). Given that precipitation mainly determines the hydrological regimes of freshwater ecosystems, any considerable change in the amount and timing of precipitation is likely to have direct and indirect effects on the characteristics of freshwater ecosystems, as well as on the organisms inhabiting these ecosystems (Poff, 1992). Climate change effects can be envisaged to differ between broadly defined ecosystem types (Schindler, 1997; Poff et al., 2002). For example, small streams may be affected more strongly than larger rivers due to the close relationships between air temperature and water temperature of small streams. Small streams are also more vulnerable to low flows and flash floods stemming from anticipated changes in precipitation. Similarly, small ponds may suffer more from thermal stress than larger lakes. Higher water temperature also threatens cold-water species in shallow lakes, which do not possess deep summer refugia for these species. Small ponds may develop towards an intermittent type with longlasting drought conditions in summer, and the shallow littoral zones of larger lakes may also suffer from extended periods of low water levels.

(2) Status of freshwater biodiversity Global change has already had effects on biodiversity in the terrestrial, marine, and freshwater realms, and predictions promise a dim future for biodiversity in natural ecosystems (Sala et al., 2000; Thomas et al., 2004). Declines in the biodiversity of freshwater ecosystems have been suggested to be especially rapid, being comparable to, or even exceeding, those estimated for the cradles of biodiversity, the tropical


42 rainforests (Ricciardi & Rasmussen, 1999; Xenopoulos et al., 2005). This rate of decline is due to the fact that freshwater ecosystems support disproportionate levels of biodiversity compared to their spatial coverage. Despite the fact that fresh waters constitute only 0.01% of the world’s water and cover only 0.8% of the Earth’s surface area, they support at least 100,000 species, amounting to about 6% of the estimated 1.8 million described species (Dudgeon et al., 2006; Balian et al., 2008). Further increases in the known levels of freshwater biodiversity are expected when ground water systems (Gilbert & Deharveng, 2002), and especially micro-organisms and invertebrates in tropical surface waters, are studied more intensively (Dudgeon et al., 2006; Balian et al., 2008). Freshwater ecosystems are well represented in the high latitudes of Eurasia and North America. Although abundant and diverse in northern regions, freshwater ecosystems in these regions do not generally harbour the levels of biodiversity found in more southerly regions, and especially fish species numbers are low at both local and regional scales in boreal regions (Matthews, 1998; Reist et al., 2006). Yet, boreal lentic and lotic ecosystems harbour both taxonomically and functionally diverse biota, although differences even among boreal regions may be evident. The regional numbers of freshwater species typically decrease with increasing latitude (Rosenzweig, 1995; Gaston & Blackburn, 2000), and such latitudinal decreases are evident even within the boreal region alone (Heino, 2001; Heino & Toivonen, 2008). Due to these relationships of freshwater biodiversity to latitude, responses of various taxonomic groups to climate change by latitudinal range shifts are likely to occur, because species’ distributions are primarily determined by temperature at large scales. These responses are likely to be seen not only in increased regional species numbers, but they also have various effects on community structure, food web dynamics, and ecosystem characteristics at the local scale (Schindler, 1997; Poff et al., 2002; ACIA, 2005; Wrona et al., 2006). The declines in freshwater biodiversity have been associated with multiple anthropogenic drivers, with climate-induced changes in water temperature and hydrological regimes typically ranking amongst the most influential in global-scale analyses (Lake et al., 2000; Xenopoulos et al., 2005). Although climate change is a global phenomenon, its effects on biodiversity are manifested at different scales of biological organisation (i.e. genes, populations, species, communities, and ecosystems) and spatial scales (e.g. habitat, local, regional, and continental). In this review, we will concentrate on species’ distributions at large scales (i.e. across regions and ecosystems).

(3) Climate change, the environmental filters perspective and freshwater biodiversity Predicting the effects of climate change on biodiversity requires consideration of multiple spatial and temporal scales (Fig. 2). Tonn (1990) devised a framework for understanding the effects of climate change on lake fish

Jani Heino, Raimo Virkkala and Heikki Toivonen communities, although this model is very helpful in general in providing a heuristic approach to understanding the relationships between climate, environment, and freshwater communities across multiple scales. In this model, environmental variables prevailing at different scales, ranging from global to habitat, are understood as filters selecting species from the species pool at larger scales to coexist in local communities (Fig. 2). Given the considerations of various scales, it has become clear that local communities are not solely a product of local environmental filters, but they also bear imprints of filters at larger scales (e.g. Gaston & Blackburn, 2000). In the context of climate change, it is thus important to consider large-scale processes, although not in isolation to more localised processes (e.g. Poff, 1997). This is because climatic effects may vary with regard to variation among regions and among ecosystems (e.g. Wrona et al., 2006). Such observations have also been made in the terrestrial realm, with bird species’ distributions being found to be hierarchically structured (e.g. Luoto, Virkkala & Heikkinen, 2007). Thus, climatic variables are large-scale determinants, followed by land cover, and habitat composition at finer resolutions. This observation suggests that climate change effects on species’ distributions are dependent on local-scale variables.

(4) Thermal guilds of freshwater organisms Freshwater species can be divided into warm-, cool-, and cold-water types with regard to preferred temperature conditions. Originally these divisions were suggested for temperate fishes, which were divided into warm-water (preferred temperature during summer centred upon 27-31oC), cool-water (21-25oC), and cold-water (11-15oC) species (Magnuson, Crowder & Medwick, 1979). Furthermore, an additional division can be made that comprises ‘‘very cold-water species’’ that live in low water temperatures (summer temperature less than 10oC) at high latitudes (Reist et al., 2006). In boreal regions, water temperatures are not likely to exceed the upper physiological limits of warm-water species, which may thus show increased success in boreal fresh waters following climate warming. By contrast, cold-water species inclined to suitable habitats in northernmost areas may not be able to find suitable thermal conditions to escape novel, stressful thermal conditions (Wrona et al., 2006). Many high-latitude species occurring at their lower physiological thermal limits are also highly sensitive to climate changes (Danks, 1992). Thus, the effects of climate warming on species’ distributions mirror the type of organisms as related to their physiological and life-history characteristics. However, given the strong responses of individual freshwater organisms to temperature, anticipated climate warming is likely to have considerable effects on the geographical distributions of freshwater organisms. These effects are likely to be species- and ecosystem-specific (Reist et al., 2006), making it difficult to predict the responses of all species and whole assemblages to climate change (Tonn, 1990).


Climate change and freshwater biodiversity

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Continental species pool

Regional species pool

Watershed

Ecosystem

Macrohabitat

Decreasing spatial scale of environmental filters

Feed-back effects from smaller scales to species pools

History and Climate

Microhabitat

ASSEMBLAGE

Fig. 2. A schematic model of environmental filters affecting regional and local assemblages. The continental species pool is determined by speciation and extinction processes at very large spatial and temporal scales. Filters at the largest scale in this scheme are history (e.g. speciation, extinction, dispersal) and climate (e.g. temperature, precipitation, energy) which determine the structure of regional species pools. Within the limits of regional species pools, there are filters at four major levels that eventually determine local assemblages. These are watershed (e.g. vegetation, hydrological regimes), ecosystem (e.g. temperature, water chemistry), macrohabitat (e.g. depositional versus erosional habitats, macrophyte cover) and microhabitat (e.g. macrophyte structural complexity, substratum particle size) filters. These filters determine the diversity and composition of assemblages through species traits. Only species with suitable traits are able to overcome the challenges presented by the filters at each scale. For original ideas, see Tonn (1990) and Poff (1997).

II. OBSERVED AND PREDICTED CHANGES IN SPECIES’ DISTRIBUTIONS (1) How to assess the effects of climate change on species’ ranges? There are two main avenues to examine the effects of climate change on species’ distributions. First, one can examine the responses of species’ distributions to presentday climatic conditions, and use correlations from such ‘‘bioclimatic envelopes’’ to predict shifts in distributions

with regard to projected future climate scenarios (Berry et al., 2002; Heikkinen et al., 2006). Second, one can examine changes in the range sizes and limits based on historical distribution records and present-day distributions. Both these approaches have their strengths and limitations. Bioclimatic envelope models have been criticised, because they assume that climate is the primary factor determining species’ distributions, and that range shifts will occur very rapidly with climate change (Woodward & Berling, 1997; Hampe, 2004). This criticism may be particularly appropriate for many groups of freshwater organisms that inhabit isolated systems in a matrix of inhospitable terrestrial environments and require suitable corridors for dispersal to track changing climate (Poff et al., 2002). Although it has been suggested that bioclimatic envelope models should be supplemented by information on land cover characteristics, biotic interactions, and dispersal (Heikkinen et al., 2006), it may still be more difficult to predict the future distributions of many groups of freshwater organisms than of terrestrial organisms. There may also be considerable among-group differences in the degree to which freshwater species’ responses to projected climate change can be predicted. These range from low predictability for those organisms relying on purely aquatic means of dispersal (e.g. fish) to higher for those capable of crossing terrestrial habitats between isolated water bodies either by active adult flying (e.g. aquatic insects) or by mainly passive means (e.g. diatoms). Thus, the latter two types of organisms are likely to show the strongest potential in bioclimatic envelope modelling of freshwater species. However, due to the lack of comprehensive surveys over large areas and, consequently, detailed distribution maps for most freshwater groups compared to those for a number of terrestrial groups, it may be difficult to derive reliable species-climate models and use this information in accurate bioclimatic modelling of freshwater species. However, although comprehensive presence-absence data (e.g. analysed using general additive models) is desirable for studying species’ responses to climate change, even less extensive presence-only data (e.g. analysed using genetic algorithms for rule-set prediction) may be utilised in bioclimatic modelling when more sophisticated data are not available (for examples, see Iguchi et al., 2004; Herborg et al., 2007). The absence of comprehensive distribution records for most freshwater taxa across large regions also hinders comparisons of historical and present-day data. Although some freshwater groups have been surveyed more fully in recent decades, there is typically a lack of historical data on species’ distributions at the same sets of sites or regions. Thus, basing predictions of future shifts in species’ distributions to comparisons of historical data is limited in most regions of the world. An exception in this regard is Britain, where many freshwater groups have been surveyed for a long time, enabling at least suggestions about the likely responses of species to projected climate change (Hickling et al., 2006). However, the degree to which the freshwater species studied in Britain respond similarly to climate change in other regions is unclear, given the differences in the anthropogenic alteration of landscapes there and in more northerly regions.


44 (2) Observed and predicted shifts in species’ distributions at large geographic scales Many species have been and are expanding their ranges to higher latitudes and altitudes in response to climate change (Parmesan & Yohe, 2003; Hickling et al., 2006). However, most of the evidence for these changes comes from the observed shifts in the distributions of a few well-studied terrestrial organism groups, including vascular plants, birds, and butterflies, while the evidence is more limited for a great majority of organism groups. There is also only limited information on the responses of freshwater species’ distributions to past and anticipated climate changes, although the observed and potential effects of climate change on the characteristics of freshwater ecosystems have been widely considered (Schindler, 1997; Poff et al., 2002; ACIA, 2005; Wrona et al., 2006). Recently, however, there has been increasing interest in examining changes in the distributions of freshwater species in relation to climate change (Hickling et al., 2006). Among freshwater species, the interest has mainly been on fish, and only scattered information exists on the effects of climate change on other freshwater groups. However, Hickling et al. (2006) studied several terrestrial and freshwater taxa in Britain. Their main finding was that the shifts in the ranges of species in the freshwater taxa and poorly studied terrestrial taxa were as evident as, or even exceeded, those previously observed for vascular plants, birds, and butterflies in Britain over the last few decades. Amongst the freshwater taxa, they studied dragonflies and damselflies, aquatic bugs, and fishes. Species in these taxa showed, on average, as strong distributional shifts northwards as the terrestrial groups, but there were some differences among the freshwater groups. Dragonflies and damselflies showed the greatest northward shift, followed by aquatic bugs and fish. The degree to which these findings are related to the dispersal capability of the taxa can only be speculated. It is possible, however, that being the strongest dispersers amongst the three taxa, dragonflies and damselflies were able to track changed climatic conditions more easily than fish that rely on suitable dispersal corridors for successful dispersal to new regions and water bodies. The findings for dragonflies and damselflies were strongly similar in a closer comparison of the British species: most species showed increased range sizes accompanied by northward shifts of range limits (Hickling et al., 2005). Despite the general trend of increasing range sizes, species classified as ‘‘northern’’ showed either retractions of their southern range limits northwards (Somatochora arctica, Coenagrion hastulatum) or declines in range size (Aeshna caerulea, Leucorrhinia dubia). However, Thomas, Franco & Hill (2006) noted that the failure to find range reductions in many northern species may be due to a failure to survey the distributions of species at sufficiently fine resolutions, and thus it is likely that climate-driven range retractions may be more common than has been generally found. Hickling et al. (2005) also suggested that species in many other aquatic insect groups might show similar distributional shifts to dragonflies and damselflies in response to climate change,

Jani Heino, Raimo Virkkala and Heikki Toivonen but there is a dearth of data outside Britain in documenting such shifts for typically poorly recorded aquatic insect taxa. Chu, Mandrak & Minns (2005) modelled the distributions of a number of Canadian fish species in relation to temperature and connected fish distributions to climatic conditions, including growing degree days, mean annual air temperature, and total annual precipitation. They found that the ranges of most species were determined, at least in part, by present-day regional climatic conditions across Canada. Thus, given these relationships, they predicted that there will be considerable range shifts in response to changed climatic conditions, although the directions of these shifts would differ among species. In general, coldwater species were predicted to become extirpated from the southern parts of their present ranges, while cool- and warm-water species will expand northwards (Chu et al., 2005). Especially species that are presently restricted to southern Canadian regions were predicted to respond strongly to climate change and show northward range shifts (Chu et al., 2005). Similar predictions of increasing ranges of cool- and warm-water fish species have been suggested in Finland, while cold-water species may show range restrictions along with shifts northwards (Lehtonen, 1996). Indeed, the distributions of warm-water fish species have been predicted to shift 500 km northwards with a doubling of CO2 in the atmosphere (Eaton & Scheller, 1996). However, range expansions can only be manifested if fish are able to disperse to new regions where conditions have become favourable. This requires suitable dispersal routes for fish in terms of connecting watercourses, or if such range shifts are facilitated by deliberate introductions. Concentrating on a single warm-water species that currently occurs at its northern range limits in southern Canada, Jackson & Mandrak (2002) found that smallmouth bass (Micropterus dolomieu) distribution was limited by temperature-related variables. They predicted that this species would expand to new regions, increasing the number of occupied lakes, with projected changes in temperature. The invasion of new lakes would also have considerable repercussions on fish species sensitive to this efficient predator, with estimates that more than 25,000 populations of a few cyprinid species would be decimated following the expansion of smallmouth bass in Ontario alone (Jackson & Mandrak, 2002). Similarly negative effects might be expected to occur with the expansion of this warm-water species across the whole of Canada by 2100 (Sharma et al., 2007). Thus, as an indirect effect of smallmouth bass expansion, other fish species also were expected to show strongly altered distributions at regional and local scales. Indirect studies on the distribution patterns of a number of other freshwater groups in northern regions have also suggested a role for climate in controlling species distributions at geographical scales. Heino (2001) found that the composition of regional freshwater biotas was rather closely related to climatic variables in Northern Europe, although there were some notable exceptions in the responses of taxonomic groups to climatic conditions. While macrophytes, dragonflies, beetles, and fish as groups showed clearer responses to regional climatic conditions, with


Climate change and freshwater biodiversity decreasing species richness towards north and higher altitudes, stoneflies showed an opposite pattern of species richness. Being mainly inclined towards cold- and coolwater stream environments at high latitudes and mountainous areas, these deviating responses of stoneflies were not unexpected. Thus, one important lesson from these analyses emerged again: different taxonomic groups may not show similar responses to climate change, which is likely related to the proportions of cold-, cool-, and warm-water species in the taxonomic group. Although the other taxonomic groups also included some seemingly cold-water species with ranges inclined towards high latitudes, they did not change the general pattern for the groups as a whole (Heino, 2001). For instance, although there are some northern species, most macrophyte species generally have ranges inclined towards southern parts of boreal regions, likely being a consequence of an indirect response to climate (Heino & Toivonen, 2008). While maximum summer temperature per se may not necessarily determine the success of macrophyte species, the increase in the growing degree days and ice-free period with higher mean annual temperatures may have major effects on the geographical distributions of macrophytes in boreal regions. In an analysis of freshwater animal species across Europe, Hof, Bra¨ndle & Brandl (2008) found that regional species richness varied to some extent with latitude. An additional important finding was that patterns differed between lentic and lotic species, with the former showing a non-significant and the latter a significant negative richness-latitude relationship. The authors hypothesised that differences among species of the two habitat types reflect their propensity for dispersal. Thus, because lentic species live in a less persistent environment than lotic species, they have evolved more efficient dispersal capacities. These efficient dispersal strategies, in turn, facilitated the colonisation of denuded areas by lentic species after the last ice age (Hof et al., 2008). If these findings are also applicable to the responses of species to climate change, then one would expect that lentic species show more rapid colonisation of northern regions than lotic species. (3) Specific surveys of streams and lakes across latitudinal gradients While the above studies considered the climate-biota relationships based on broad-scale data across grids or regions, there is also some information on the distributions of freshwater organisms based on surveys of streams and lakes along large geographical gradients. Surveys of streams in northern regions have suggested that biotic assemblages show at least some variability related to latitude and, consequently, to covarying temperature conditions. Evidence for this comes primarily from studies of macroinvertebrates in northern North America (ACIA, 2005) and northern Europe (Sandin & Johnson, 2000; Heino et al., 2002). Although studies of latitudinal variability based on local stream samples may be confounded by the effects of local-scale factors instead of large-scale climatic gradients, these findings suggest that the distribution and abundance of many stream macroinvertebrate species may be determined

45 by water temperature, and they should thus be sensitive to anticipated climate change. Indeed, in addition to a suite of environmental variables at various other scales, spatial and regional factors also appeared to be important in determining the structure of stream macroinvertebrate assemblages in northern Europe (Heino et al., 2003; Sandin & Johnson, 2004). There is only circumstantial evidence about the relationships between climate and geographical distribution patterns of lake organisms. For lake zooplankton, Patalas (1990) found that the maximum species richness peaked in regions where mean open-water temperature was approximately 15oC, whereas species richness declined with both increases and decreases in temperature across a geographical gradient between 45oN and 55oN. This observation was attributed to the overlap in the distributions of southern warm-water and northern cold-water species that were able to occupy the same regions and lakes at this temperature. Regions with the highest regional and local diversity of zooplankton may move northwards with shifts in the distributions of species in response to climate change (Schindler, 1997). These responses to climate forcing may be rapid, as suggested by a multi-year survey of lake zooplankton just south from boreal regions (Stemberger et al., 1996). If these predictions have a more general applicability, then at least the southern parts of boreal regions can be expected to receive several new warm-water species, while cool-water species may not be negatively affected by moderate changes in water temperature. Thus, the northward expansion of zooplankton species would result in gains of zooplankton biodiversity in boreal regions. Direct and indirect effects of climate through physiological limitation and increase in the length of growing season may also be responsible for the observed present-day relationships between temperature and diatom distributions across lakes in northern regions (Weckstro¨m & Korhola, 2001). Thus, it is not surprising that diatoms should also respond strongly to projected climate change, and such responses by this group have already been observed in highlatitude lakes (Sorvari, Korhola & Thompson, 2002). Surveys of lakes along geographical gradients have also shown that temperature is amongst the most important factors in accounting for variability in the distributions and assemblage structure of macroinvertebrates. These observations are not limited to any single high-latitude region, but similar findings for midges have emerged from both Eurasia and North America (Nyman, Korhola & Brooks, 2005; Barley et al., 2006). Thus, climate change is likely to modify strongly the distributions of these organisms across highlatitude lakes (Walker et al., 1991; Smol et al., 2005). In the same vein, differences in lake macroinvertebrate assemblages among ecoregions in Sweden suggest some climatic control on these organisms (Johnson, 2000), with possible shifts in assemblage structure following climate change. (4) Synthesis The few studies reviewed above give only a coarse picture of the relationships between climate and the distribution patterns of freshwater species. The lack of detailed


46 bioclimatic models of species’ ranges in most freshwater groups across broad geographical regions severely hinders the prediction of species’ responses to projected climate change. Given that a 4oC rise in water temperature (i.e. well within the predicted limits for the near future) broadly corresponds to a 680 km latitudinal shift in thermal regimes (Sweeney et al., 1992), changes in the ranges of freshwater species are very likely to occur (Poff et al., 2002). Although changes in the thermal regimes of lakes and streams are probably the most important factors modifying the distributions of freshwater organisms, other direct and indirect effects may also be active. In lentic ecosystems, for example, hydrological changes also include the shortening of ice cover duration and the prolongation of summer stratification, which may affect the distributions of species within and among lakes (Magnuson, Meisner & Hill, 1990; DeStasio et al., 1996; Schindler, 1997). In lotic ecosystems, more intense flash floods and extended periods of low flow may also contribute to changes in biodiversity (IPCC, 2001b; Poff et al., 2002). Understanding the present-day determinants and future shifts in species’ ranges and diversity gradients is extremely important, because geographical distributions ultimately determine the composition of regional species pools from which local communities are assembled (Tonn, 1990). Regional species pools in boreal regions are likely to become more diverse with the influx of new members of mainly warm-water species, although regional species pools in the southern limits of boreal regions may also lose some cold-water species. However, given that boreal regions were covered by ice until approximately 10,000 years ago, the post-ice age colonisation may not yet have been accomplished by all potential species. Furthermore, it is possible that most species that have already colonised boreal regions are generalists as related to temperature tolerance and other environmental demands, and thus more specialist species are likely to colonise freshwater ecosystems in northern regions in the near future.

III. INTERACTIONS BETWEEN CLIMATE CHANGE AND MULTIPLE ANTHROPOGENIC STRESSORS Because the interactive effects of climate change and other anthropogenic stressors on freshwater ecosystems have been considered recently (Schindler, 2001; Wrona et al., 2006), the following will concentrate on the potential effects of these interactions on biodiversity. Particular emphasis is directed at acidification, eutrophication, landuse change, and exotic species, yet acknowledging that many other anthropogenic and natural changes in ecosystem conditions may affect biodiversity. Our aim was not to provide an exhaustive review, but rather short overviews of these topics. The reader should thus consult previous comprehensive articles on these topics (acidification: Schindler, 1997; eutrophication: Flanagan et al., 2003; land use changes: Allan, 2004; exotic species: Rahel & Olden, 2008).

Jani Heino, Raimo Virkkala and Heikki Toivonen (1) Climate change and acidification Climate change is likely to have complex interactions with acidification. For freshwater ecosystems, climate warming may accelerate the acidification of streams and affect negatively the recovery process of acidified lakes (Schindler, 1997). However, contrasting observations of increasing alkalinity of lakes with climate change have also been made (Schindler et al., 1996). Thus, it is difficult to predict the overall consequences of climate change for the acidity of freshwater ecosystems, because regional differences in the atmospheric deposition of acidifying substances and acid pulses through runoff from the catchments may affect the degree of acidification or increased alkalinity of fresh waters (Schindler, 1997). Given that increasing anthropogenic acidity generally leads to a reduction or overall impoverishment of freshwater biodiversity (Giller & Malmqvist, 1998), while decreasing acidity typically has the opposite effect, the influences of climate change through acidity on biodiversity are difficult to predict. The likely scenario is that freshwater ecosystems subject to acidification show exacerbation of negative effects on biodiversity. However, even within a region, climate warming may have contrasting effects on temporal changes in the biodiversity of acidic and neutral streams (Durance & Ormerod, 2007). (2) Climate change and eutrophication Climate change may be associated with both decreased and increased levels of nutrients entering freshwater ecosystems. For example, it has been found that climate warming led to a concomitant decline in the phosphorus concentration of lake water, which resulted in changes in phytoplankton communities in a Canadian region (Schindler, 1997). Chlorophyll concentrations declined, but the biomass and diversity of phytoplankton increased to some degree (Schindler et al., 1990). Opposite effects of increased nutrient levels with increasing precipitation and runoff should also be expected, and if increases are sufficient enough, they may result in reduced phytoplankton diversity in lakes. Similar responses may be found in other organism groups as well. However, it is important to keep in mind that the effects of increased levels of nutrients on biodiversity may depend on the natural state of an ecosystem. Because most lentic and lotic ecosystems in the northern parts of boreal regions are naturally oligotrophic, increases in nutrient levels may, at least at the beginning of the eutrophication process, lead to increased levels of biodiversity. By contrast, in the southern parts of boreal regions, many freshwater ecosystems are eutrophic due to either natural or anthropogenic causes, and increased nutrient levels in such systems may often lead to reductions in biodiversity. In the Arctic, climate warming may thaw upper layers of permafrost, leading to increased levels of phosphorus in streams and lakes (Hobbie et al., 1999). Streams typically respond to this nutrient increase by a higher production of diatoms, although grazers may control the effects of increased nutrients on algae. Increases at lower trophic levels are also likely to lead to changes at higher trophic


Climate change and freshwater biodiversity levels, resulting in increased fish biomass (Hobbie et al., 1999). If nutrients are available in excess, mosses may become dominant primary producers and use most of the available nutrients. Changes in such key organisms that provide structural habitat for other organisms may also have various effects, and mosses in particular have been attributed this role in streams (Stream Bryophyte Group, 1999). Increased moss abundance could, for example, lead to increased abundance and diversity of benthic macroinvertebrates. Thus, whether changes in the community structure of arctic streams result from the direct effects of warming and eutrophication or indirect effects is difficult to judge. However, warmer and more nutrient rich waters are likely to support novel communities in arctic streams in the future. Arctic lakes have also been predicted to show increased algal production and biomass with both warming climate and increased nutrient inputs. Based on the relationships between latitude, phosphorus, and algal biomass in lakes, Flanagan et al. (2003) predicted that climate change will lead to increased productivity in arctic lakes if these systems are subjected to the projections of both increasing temperatures and nutrient inputs of climate change models. Increases in the productivity and biomass at the base trophic level may have considerable bottom-up effects throughout the food web. For example, fish are generally limited by very low levels of productivity typical of arctic lakes, and increased algal productivity might allow these systems to become colonised by predatory fishes (Flanagan et al., 2003). If fish are absent, then the system may remain in a state where increased algal production is controlled by grazing zooplankton (Wrona et al., 2006). In the southern parts of boreal regions, streams and lakes vary widely in nutrient levels, ranging from oligotrophic to eutrophic. The degree to which increased nutrient inputs in association with climate change are portrayed in biodiversity is thus likely related to the starting conditions of an ecosystem. Oligotrophic lakes may experience increases in the diversity of various organism groups, while eutrophic lakes may show reduced levels of biodiversity if excess nutrients change conditions to hypereutrophic. Indirect support for these predictions comes from studies of fish communities, where it has been found that fish community structure typically changes along a gradient from oligotrophic to eutrophic lakes (Tammi et al., 1999; Jeppesen et al., 2000). Oligotrophic lakes are typically dominated by fish species from families other than cyprinids, whereas cyprinids typically are dominant in eutrophic conditions. While both highly oligotrophic and highly eutrophic lakes may support low fish diversity, mesotrophic lakes typically harbour greater levels of diversity. Similarly to fish communities, surveys of lakes for macrophytes have shown discernible shifts in community structure and species richness from oligotrophic to eutrophic lakes, with increasing diversity towards slightly eutrophic conditions (Rørslett, 1991; Toivonen & Huttunen, 1995). However, some boreal lakes may even develop towards hypereutrophic conditions typical of more southerly regions, which is likely to lead to drastic reductions of macrophyte diversity in lakes (Sand-Jensen et al., 2000). Increases in nutrient levels in association with

47 climate warming have been predicted by climate change models (Flanagan et al., 2003), and climate warming may also render some boreal regions suitable for land uses that increase nutrient inputs into fresh waters. (3) Climate change and land cover alterations Climate change may affect freshwater ecosystems through changes in the land cover of catchments and the characteristics of riparian zones. These changes may be both natural consequences of shifts in terrestrial vegetation and anthropogenic alterations of land cover. Climate change has been suggested to lead to changes in terrestrial vegetation, and even shifts in dominant vegetation have been predicted following climate change in the near future (Saetersdal, Birks & Peglar, 1998). In the boreal zone, forests presently dominated by coniferous trees may show altered dominance by the presently occurring species (Kelloma¨ki et al., 2001, 2005), and may even become replaced by broad-leaved deciduous trees currently characterising forests in the more southerly regions (Sykes & Prentice, 1995). Such changes in dominant forest trees may have considerable effects on the community structure and ecosystem functioning in freshwater ecosystems. In particular, headwater streams and upper littoral zones of lakes may respond to shifts from the dominance by coniferous trees to broad-leaved deciduous trees in the riparian zone (Allan et al., 2005). This is because these systems are to a considerable degree driven by allochthonous inputs from surrounding terrestrial vegetation. It has been shown for both headwater streams (Allan, 1995) and littoral zones of lakes (France, 1995) that leaf inputs from riparian vegetation form a major source of food for invertebrate consumers in these systems. Thus, given that the needles of coniferous trees are inferior food sources to leaves of deciduous trees to shredding freshwater invertebrates (Friberg & Jacobsen, 1994), a shift in dominance patterns in riparian vegetation will likely lead to increases in at least the shredder functional group of freshwater macroinvertebrates (Wallace & Webster, 1996; Covich, Palmer & Crowl, 1999). Further changes in invertebrate communities could also be expected. A conservative prediction is that at least the relative abundances of macroinvertebrate species will be different in novel conditions, but potential effects may even be reflected in clear shifts from the species composition and species richness typical of coniferous forest streams and lakes to those typical of broad-leaved deciduous forests. Similarly pronounced changes may be observed in the Arctic, where the encroachment of forest vegetation to mountain birch woodlands and barren tundra (Krankina et al., 1997; Chapin et al., 2005) is likely to shift the community structure and ecosystem functioning in headwater streams and lake littoral zones. Climate change may increase some types of land uses when these become possible in northern regions. For example, climate change may facilitate expanding agricultural areas further north, changing the dominant trees grown in managed forests, and altering patterns of runoff from the catchment to freshwater ecosystems. All these exemplary changes in land use would obviously have direct


48 and indirect influences on biodiversity in freshwater ecosystems, because changes in the land use at the catchment level are portrayed in the physical and chemical characteristics of these ecosystems (Allan, 2004). For example, if agricultural land uses increase in importance and area extent in boreal regions, the likely consequence is increased loads of nutrients to freshwater ecosystems, with effects on water quality, habitat characteristics, and biodiversity (Richards, Johnson & Host, 1996, Sponseller, Benfield & Valett, 2001). Biodiversity in freshwater ecosystems generally declines with increased levels of agriculture in the catchment and increased amounts of nutrient runoff (Allan, 2004), although most evidence comes from studies conducted across heavily altered catchments in temperate regions. However, the situation may not be the same in less altered catchments of boreal regions (see section III. 2 above). The conversion of forest vegetation of riparian zones to agricultural land would also exacerbate the effects of climate warming on headwater stream ecosystems. For example, algae may be strongly affected by increased light and temperature after the removal of riparian vegetation, with increased algal biomass and altered community composition (Allan, 2004). Similarly, the community structure of macroinvertebrates may change, with increased abundances of grazers and decreased abundances of shredders, following changes in the production base from allochthonous to autochthonous (Delong & Brusven, 1998). Such changes could occur even in the absence of increased nutrient inputs, although both the alteration of riparian zone and increased runoff of nutrients from the catchment are effects of increased agricultural land use. Furthermore, increased temperatures due to the removal of riparian vegetation are likely to change the performance of cold- and warm-water organisms. These effects could, however, be mitigated by land-use planning to retain mostly intact riparian zones and by planting trees in the riparian areas currently lacking them. (4) Climate change and exotic species Climate change is likely to interact with the invasions of exotic species by (i) increasing the invasibility of ecosystems, (ii) through effects of altered climatic conditions on native species, and (iii) increasing the invasive potential of exotic species (Thuiller, Richardson & Midgley, 2007). Increases in temperature may be especially relevant in boreal regions through increasing the probability of the invasion and establishment of exotic species that typically originate from more southerly regions. The northern range limits of such species are typically determined by minimum winter temperatures. Thus, following climate change, many boreal freshwater ecosystems may become suitable for the breeding populations of various exotic species, with often dramatic influences on native species, biotic communities, and ecosystem processes (Rahel & Olden, 2008). These influences might stem from predation, competition, and spread of parasites and diseases to which species native to boreal freshwater ecosystems are not adapted (Wrona et al., 2006). The negative impacts of exotic species are dire

Jani Heino, Raimo Virkkala and Heikki Toivonen especially if they are (i) directed at keystone species, (ii) lead to general reductions in biodiversity, or (iii) change trophic relationships in a recipient ecosystem. The invasion process and impacts of aquatic exotic species have been considered extensively in other contexts (Rahel, 2002; Ward & Ricciardi, 2007), as well as in association with anticipated climate change (Rahel & Olden, 2008), so we will not pay much attention to this topic here. In summary, the number of exotic species is likely to increase in boreal freshwater ecosystems following climate change. However, it is difficult to provide any general predictions about what the effects of these introductions will be on biodiversity, because these effects are often context-dependent (Ward & Ricciardi, 2007), and because no studies on the interactive effects of climate change and exotic species on freshwater biodiversity have been conducted in boreal regions.

IV. CLIMATE CHANGE ADAPTATION, CONSERVATION AND FRESHWATER BIODIVERSITY Because climate change is likely to have major effects on the distributions of freshwater organisms, adaptation to such changes and measures of change are necessary. These adaptations must be understood broadly, and they include not only evolutionary adaptations of organisms, but also human-assisted means to accommodate organisms to changed conditions. Recently, mainly addressing terrestrial ecosystems, adaptation to climate change was divided into two types (IPCC, 2001a; Carter & Kankaanpaä¨, 2003). First, autonomous adaptation is reactive, and it is related to the responses that take place after climate change impacts have been realised. This is the type of adaptation typically shown by ecosystems and organisms, as has also been suggested by the responses of organisms to past climate changes by shifts in distributions rather than by showing genetic adaptations to changed conditions (Poÿry & Toivonen, 2005). For freshwater organisms, the greatest challenges of climate change adaptation will be related to the degree to which certain aquatic habitats are diminished, the abilities of species to disperse to higher latitudes or altitudes, and possibilities to overcome the effects of increased isolation of aquatic habitats due to human activities (Poff et al., 2002). Second, planned adaptation relates to the conservation of biodiversity in the face of climate change. This entails establishment of protected areas, habitat management, planning of dispersal corridors for various types of species (Hannah et al., 2002), and minimising anthropogenic effects such as pollution, habitat destruction, and exotic species introductions on aquatic ecosystems (Poff et al., 2002). Planned adaptation is not only reactive, but also pro-active. The term pro-active is related to the fact that adaptations to climate change may be accomplished prior to the projected impacts. This is what is the most effective means of humans in countering the possible impacts of climate change. The whole topic of climate change adaptation constitutes a wide variety of subjects, but in the following we will concentrate on six


Climate change and freshwater biodiversity adaptation measures discussed previously (CBD, 2003: Poÿry & Toivonen, 2005) and examine these in the context of the conservation of freshwater biodiversity in northern regions. More specific, recent reviews of the importance of freshwater protected areas for biodiversity conservation can be found elsewhere (Abell, Allan & Lehner, 2007; Suski & Cooke, 2007). (1) Networks of protected areas Extensive networks of protected areas provide the most efficient way of conserving biodiversity in the face of climate change. A number of studies have examined the capacity of currently protected areas to maintain biomes and species (Hannah et al., 2002; Scott, Malcolm & Lemieux, 2002), and some studies have also aimed to develop methods that could aid in selecting protected areas with regard to the current and projected future changes of species’ ranges (Arau´jo et al., 2004, Pyke & Fischer, 2005). In general, considerable changes in biome composition and species pools are expected to occur in the currently protected areas (Poÿry & Toivonen, 2005). Thus, unless the protected areas network encompasses a wide variety of regional climatic conditions, the representation of all species in the protected areas network is not possible. There are further problems with the current protected areas network with regard to the conservation of freshwater biodiversity. Given that currently protected areas are typically delineated based on the representation of terrestrial ecosystems and a low number of taxonomic groups (e.g. vascular plants and terrestrial vertebrates), it is unclear if freshwater biodiversity is adequately protected in current protected areas network, and if future shifts in freshwater species’ ranges could be accommodated by these areas. To be efficient for freshwater organisms, protected areas should naturally be based on the characteristics of freshwater ecosystems and the requirements of freshwater organisms (Saunders, Meeuwig & Vincent, 2002; Toivonen, Leikola & Kallio, 2004; Abell et al., 2007). For example, taking a catchment perspective instead of strict conservation of terrestrial areas that can be easily bounded and protected would be more desirable for conserving freshwater biodiversity (Dudgeon et al., 2006). This is not to say that the protected areas should only be delineated based on freshwater biodiversity alone, but rather an integrated approach for both freshwater and terrestrial ecosystems is likely to be the most fruitful avenue for conserving wholesale biodiversity in reserve networks (Abell, 2002). (2) Protection of large and environmentally heterogeneous areas This is certainly a desirable approach for conserving freshwater biodiversity, as it is for terrestrial biodiversity. Because large and heterogeneous areas are more likely to incorporate a wider array of different types of lentic and lotic ecosystems than smaller and more homogenous areas, this approach should lead to preservation of much of regional freshwater biodiversity. Large protected areas

49 should also accommodate larger parts of whole catchments that are of vital importance for the functioning of freshwater ecosystems and harbouring diverse ecological communities (Pringle, 2001; Dudgeon et al., 2006). More heterogeneous protected areas, for example, in terms of mountainous and lowland areas would also provide possibilities for freshwater organisms to track suitable temperature conditions following climate change. However, establishment of large and heterogeneous areas comprising mostly minimally affected ecosystems may be problematic even in boreal regions (that are less affected by anthropogenic land use than more southerly regions) due to socio-economic factors and conflicting demands for land development. (3) Dispersal corridors Dispersal corridors are vital for species to track changes in climatic conditions. This is especially relevant for freshwater organisms that rely on rivers and streams for successful dispersal among water bodies. If suitable dispersal corridors are absent, or if dispersal is prevented by man-made constructs such as dams (Allan & Flecker, 1993; Malmqvist & Rundle, 2002), then the responses of freshwater species to climate change may not be realised. In the worst case, if dispersal is prevented, species may be extirpated regionally if they are not able to escape novel, unsuitable thermal and hydrological conditions. For example, species may be unable to find river systems connected to northern fresh waters, although there are regional differences in this regard (compare the directions of major river systems in North America, Siberia, and Europe). These problems of dispersal are possibly more severe for fish than for aquatic insects with a winged terrestrial adult stage and for various small organisms surviving passive overland dispersal. However, even for such dispersive species, the distances between suitable water bodies are likely to be important with regard to their chances of tracking climate change. In the natural settings of formerly glaciated northern regions, lakes are typically numerous and generally highly interconnected by streams, facilitating the movements of species in response to climate change after these northern regions have been reached by freshwater species (Poff et al., 2002). (4) Habitat restoration and management Habitat restoration and management are important means to maintain close-to-natural ecosystems and viable populations of species. Management of protected areas (Suffling & Scott, 2002) can be static (i.e. continued protection of current components of biodiversity based on current protected areas and defined aims), passive (i.e. allowing ecological responses to climate change without human intervention), and adaptive (i.e. actions aiming at the maximisation of abilities of current communities and species to adapt to climate change). With regard to freshwater biodiversity in protected areas, some hybrid of these approaches would likely to be most efficient, because not all means of adaptive management are ethically or scientifically grounded. A further problem is which species


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Jani Heino, Raimo Virkkala and Heikki Toivonen

to concentrate on, because limited funds do not allow all species to be managed (Suffling & Scott, 2002). However, as examples of adaptive management, the maintenance of the habitats and viable populations of species most valued by the society and the eradication of exotic species benefiting from climate change should be goals for the management of protected areas.

groups in Britain (Hickling et al., 2006). Preliminary analyses of bioclimatic envelopes based on coarse range maps could be conducted across large geographic scales in Europe and North America. However, more detailed distributional records would make it possible to model the relationships between climate and species’ distributions more accurately and predict future shifts in distributions more fully.

(5) Management of the matrix between protected areas

(2) Long-term monitoring networks for examining temporal changes in biodiversity

Such management is very important for freshwater biodiversity, and it constitutes the regulation of harmful anthropogenic effects on freshwater ecosystems in the matrix. This would entail planning land use so that harmful influences from built-up areas, agriculture, and forestry do not degrade the state of freshwater ecosystems. In the worst case, freshwater ecosystems in the protected areas may remain as oases of suitable environments for freshwater species, while those in the matrix areas between protected areas may be unsuitable for many species to inhabit or use for dispersal between protected areas in the face of climate change. Thus, because a great majority of freshwater ecosystems are located outside protected areas and affect those in protected areas, the matrix and its adaptive management is likely to be as important as the protection of new areas for biodiversity conservation.

Given the limited number of monitoring programs addressing year-to-year variability at multiple freshwater sites, new programs should be launched to facilitate examining directly the effects of climate change and other major anthropogenic stressors on biodiversity (e.g. Daufresne et al., 2004; Durance & Ormerod, 2007). Such programs should also help to determine how reference conditions for freshwater bioassessment are altered by climate change. Studies conducted in association with long-term ecological research sites (LTER) would be highly suitable for this purpose (see http://www.lter-europe.ceh.ac.uk/).

(6) Reintroduction of native species Although this approach is likely to be unsuitable for most freshwater species due to economic constraints and practical difficulties in breeding and moving living organisms, it is highly desirable for species that are prevented from tracking climate change due to man-made dispersal barriers, such as dams, or limited dispersal ability. Furthermore, it may be desirable to introduce such native species that play key roles in ecosystems (Hunter, 2007). Amongst such species are predatory fish species that are also favoured for sport fishing, and species that may be considered as ecosystem engineers that participate in processes important to a suite of other species (Jones, Lawton & Shachak, 1994).

V. KEY TOPICS FOR FUTURE RESEARCH There are a number of key topics that should be valuable in aiding us to predict the effects of climate change on freshwater biodiversity. The following is largely based on the themes previously suggested for biodiversity in general (Poÿry & Toivonen, 2005). (1) Acquisition of comprehensive distributional data Such data are largely only available for fish in a number of regions and for a number of other freshwater taxonomic

(3) Assessment of species and habitats most threatened by and responsive to climate change Such species in northern regions may have distributions restricted to the Arctic, as well as more southerly distributed species that may become established in northern regions. Among the most vulnerable habitats are small headwater streams across the whole boreal region, as they may be severely affected by both changed thermal conditions and hydrological regimes (Poff, 1992). Furthermore, as has been shown by the paleolimnological evidence, ponds and lakes in the Arctic may also be severely affected by climate change (Smol et al., 2005). (4) Assessment of the present protected area network This would entail modelling of the current distributions of freshwater species in the protected areas, and predicting shifts in their distribution following climate change. This information would also be valuable when planning the locations of new protected areas. (5) Estimation of dispersal corridors for freshwater species Because different types of freshwater organisms show differing dispersal capacity, dispersal corridors should be examined from a number of different perspectives. To facilitate the previous goal, it would also be important to increase our knowledge about the dispersal abilities of species. For many groups of freshwater organisms, dispersal abilities of species are poorly known (Bohonak & Jenkins, 2003), yet such information would be highly valuable when assessing the potential of species to track climate change.


Climate change and freshwater biodiversity VI. CONCLUSIONS (1) Freshwater ecosystems and their biodiversity are highly vulnerable to climate change. There is some evidence that freshwater species have exhibited range shifts in responses to climate change in the last millennia, centuries, and decades. (2) In general, the numbers of freshwater species are very likely to increase in boreal regions following future climate change. This increase is mainly attributable to the northward range shifts of warm-water species, while some coldwater species may go extinct in the southern parts of boreal regions. (3) Detected range shifts are based on findings from a relatively low number of freshwater organism groups and few regions. The lack of a wider knowledge hinders predictions with regard to general responses of freshwater biodiversity to climate change and other associated anthropogenic stressors. Thus, due to the lack of detailed distributional information for most freshwater organism groups and the absence of distribution-climate models, future studies should aim to further our knowledge about these aspects of the ecology of freshwater organisms. (4) Such information is not only important with regard to predicting the responses of freshwater species to both directional climate change and short-term climatic oscillations, but also when assessing the capacity of protected areas to accommodate future changes in the distributions of freshwater species. (5) This is a huge challenge, because most currently protected areas have not been delineated based on the requirements of freshwater organisms. Thus, the requirements of freshwater organisms should be taken directly into account in the future delineation of protected areas and in estimating the degree to which protected areas support biodiversity in the changing future climate. (6) Given that most freshwater ecosystems are, however, located outside protected areas, management of the matrix between reserves should also be an important part of adaptation to climate change.

VII. ACKNOWLEDGEMENTS We thank Risto K. Heikkinen, Kai Korsu, Steve Ormerod, anonymous referees, and the editors for comments on earlier drafts of this article. J. H. was supported by a grant from the Maj and Tor Nessling Foundation. VIII. REFERENCES ABELL, R. A. (2002). Conservation biology for the biodiversity crisis: a freshwater follow-up. Conservation Biology 16, 1435–1437. ABELL, R. A., ALLAN, J. D. & LEHNER, B. (2007). Unlocking the potential of protected areas for freshwaters. Biological Conservation 134, 48–63.

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