Actual state of European wetlands and their possible future in the context of global climate change
The present area of European wetlands is only a fraction of their area before the start of large-scale human colonization of Europe. Many European wetlands have been exploited and managed for various purposes. Large wetland areas have been drained and reclaimed mainly for agriculture and establishment of human settlements. These threats to European wetlands persist. The main responses of European wetlands to ongoing climate change will vary according to wetland type and geographical location. Sea level rise will probably be the decisive factor affecting coastal wetlands, especially along the Atlantic coast. In the boreal part of Europe, increased temperatures will probably lead to increased annual evapotranspiration and lower organic matter accumulation in soil. The role of vast boreal wetlands as carbon sinks may thus be suppressed. In central and western Europe, the risk of floods may support the political will for ecosystem-unfriendly flood defence measures, which may threaten the hydrology of existing wetlands. Southern Europe will probably suffer most from water shortage, which may strengthen the competition for water resources between agriculture, industry and settlements on the one hand and nature conservancy, including wetland conservation, on the other.
KeywordsWetlands Carbon sequestration Hydrology Biodiversity Climate stabilization Ecosystem services
The area of the European continent is about 107 km2, including the European part of Russia. Excluding Russia and marine areas, the area is about 6,710,000 km2. The climate of Europe is characterized by marked climatic gradients (from the cold climate in the polar regions in the north to the dry and warm Mediterranean climate in the south, and from the oceanic climate in the west to the continental climate in the east). Europe is a densely populated continent (67 inhabitants per km2) but with a high heterogeneity in the distribution of the human population. More densely inhabited areas are located in central-west Europe, especially in urban areas where 400 inhabitants per km2 are frequent. In all the other zones of Europe, large areas remain with a rather low human population density (e.g., less than 10 inhabitants per km2 in parts of Finland, Spain, Greece or Poland). Much of the area of Europe has been settled at least since the beginning of the Middle Ages (i.e., for about 1,500 years).
Estimated wetland coverage in Europe as identified by the European inventory dataset (according to Nivet and Frazier 2004)
Number of countries
Total land area of study region (km2) (excluding marine areas; including Asian part of Russia and Azerbaijan) (km2)
Total area of wetlands identified (km2)
Percentage of land area (excluding marine areas) covered by these wetlands
Land area of region (km2), excluding Russia and marine areas
Total area of wetlands identified in this study, excluding Russian wetlands (km2)
Percentage of land area, excluding Russia and marine areas, covered by these wetlands
Nivet and Frazier (2004) have reported on the most recent inventory of European wetlands. The area of coastal wetlands is currently estimated at about 46,000 km2, which is about 2% of the total area of European wetlands including Russia. With their total area of 2,486,000 km2, inland wetlands comprise the largest proportion of the total wetland area. The reported area of human-made wetlands is about 20,000 km2 or 1% of the total wetland area. However, there still seems to be large uncertainty in these estimates because not all national inventories are complete, different national inventories use different definitions of wetland types and, last but not least, some of the areas reported as wetlands apparently also comprise former wetlands that have been drained.
Despite its relatively small geographic size, Europe has a very long coastline, approximately 326,000 km (Pruett and Cimino 2000). The European coastline comprises the main marine regions of the northeast Atlantic, part of the Arctic, the Baltic Sea, the North Sea, the Mediterranean Sea and the Black Sea. Much of the European coastline consists of a chain of extensive estuaries, lagoons and intertidal bays interspersed through stretches of rocky shore and sandy beaches (Adam 1990). These areas support various wetland types. Airoldi and Beck (2007) distinguish macroalgal beds, seagrass meadows, biogenic reefs, sedimentary habitats (mudflats, sandflats and subtidal soft bottoms) and emergent coastal wetlands including salt marshes.
In addition to climate, the occurrence of various wetland types is determined mainly by local geomorphological features and tidal range. The tidal range is up to several meters along the Atlantic coast (including the North Sea). Along the North Adriatic Sea, the tide range can be 1 m, both diurnal and semidiurnal (depending on the moon phase), while it is much narrower to negligible (usually several centimeters) along the remaining European coasts. In strongly tidal areas of the Atlantic and the North Sea, the seaward zone of gently sloping shores is occupied by soft-sediment habitats emerging at low tide. Muddy habitats usually occur in sheltered areas, such as sea lochs, enclosed bays and estuaries, whereas sandflats and coarser sediments tend to occur in more exposed situations on the open coast. Salt marshes are developed on suitable more elevated sites all along the European coast, often bordering estuaries. A wide tidal range is responsible for the occurrence of tides even within predominantly freshwater wetlands fringing the estuaries of rivers or streams flowing into the Atlantic or North Sea. This is the case even for the flat parts of the Mediterranean coast with an average maximum tide of 0.20 m. Long wave based phenomena caused by barometric changes (seiches) can be reflected on the coasts as 2 m wide changes of sea level in a few hours (Ranwell 1972).
Palustrine wetlands (peatlands)
The first group of natural inland wetlands (peatlands) occur in habitats characterized by the presence of organic soils, waterlogged or saturated with water, with fairly narrow annual water-table fluctuations (Rydin and Jeglum 2006). Definitions vary to some extent among countries, but peat thickness usually needs to be at least 30 cm for a site to be classified as peatland. Mires represent a subset of peatlands. Mires are living ecosystems, where peat is being formed and accumulated. In addition, peatlands also comprise drained sites e.g., in agricultural use, where a peat layer is still present (e.g., Joosten and Clarke 2002). According to Joosten and Clarke, the original peatland area in Europe (excluding Russia) was about 374,500 km2, that is around 6% of total land area; more than 50% of the original area has ceased to accumulate peat due to human exploitation, and almost 20% has ceased to exist as peatlands. Lappalainen (1996), on the other hand, has estimated that peatlands cover about 960,000 km2, that is about 20% (!) of the land area of Europe.
According to the shore exposure to wave action, one can distinguish either accumulation or erosion littoral habitats. The former ones occur in sheltered situations where ample accumulation of detritus-derived autochthonous sediments rich in organic matter takes place. The latter habitats occur in wind- and wave-exposed situations where most of the plant litter and detritus is washed away into the adjacent water body, and the underlying mineral layer consists of sand, gravel or withered bedrock. One can add sedimentary wetland habitats with the deposition of allochthonous sediments at the mouths of running waters entering standing water bodies. The granulometric composition of these sediments depends on the inflow velocity at each site while their chemical composition reflects that of soils in the catchment areas of the inflowing streams or rivers. For more details on the formation of lake sediments see, e.g., Bloesch (2004).
Marginal wetlands, in which the terrestrial ecophase prevails for most of an average year (Květ et al. 2002), often reach long distances from the shoreline in areas that have not been artificially drained. These marginal lacustrine wetlands are very similar to palustrine wetlands. At eu- to mesotrophic sites, they are fen-like and colonized either by shrubby or forest vegetation dominated by hygrophytic woody plants (typically Salix or Alnus), or by both natural and human-made wet grassland dominated by hygrophytic grasses and sedges. At oligotrophic sites, the character of the marginal lacustrine wetlands resembles that of transition mires or even bogs, and the water stored and flowing out of them is more or less dystrophic.
The functional interaction between lacustrine wetlands and the adjacent water body or land depends, naturally, on the width of the littoral zone which, in turn, is determined by the shore slope. Only rather wide littoral belts, like that of lake Neusiedlersee/Fertö in Austria/Hungary (Löffler 1974; Löffler and Gunatilaka 1994), possess structural and functional features of ecosystems, showing a high degree of independence of their adjacent biomes. Narrower littoral wetlands strongly interact with adjacent land and water. Nevertheless, the predominance of the detritus-bases food web is characteristic of all lacustrine littoral wetlands dependent mainly on the primary production by macrophytes, while the grazing-predatory food chain predominates in the food web in open-water (pelagial) habitats dependent mainly on the primary production by phytoplankton (Straškraba 1963, 1968; Straškraba et al. 1967; Gopal et al. 1993; Hillbricht-Ilkowska and Pieczyńska 1993).
Inland salt marshes and saline lakes (Fig. 2e) occur predominantly in south-western and south-eastern Europe (e.g., Spain, Hungary, Balkan countries) on sites where summer evaporation is intense and brings about capillary rise of soil water rich in salts (sulphates and/or chlorides) from the subsoil. This provides a unique type of wetland ecosystem for European ecodiversity, which is more abundant in other continents (Comín and Alonso 1988; Comín and Williams 1993). Elsewhere in Europe, small inland salt marshes can be found around mineral springs.
Diverse and highly dynamic systems of habitats are associated with riverine wetlands, i.e., those fed with running water—from springs and small streams through preserved segments of floodplains to both freshwater and brackish habitats of large river deltas (e.g., Purseglove 1988; Junk and Welcomme 1990; Prach et al. 1996; Middleton 2002; Haslam 2008). The hydrological régime is decisive for the structure and functioning of riverine wetlands (Duever 1990; Mitsch and Gosselink 2000). It varies according to the climatic zone and geomorphological features of the respective river and stream headwater as well as remaining catchment areas. At high altitudes, it is only on mountain plateaus or gentle slopes that smaller or larger floodplains develop around springs and along slowly flowing and often winding streams. In many of them, smooth transitions can be observed to peat-forming wetland systems, i.e., mires. In steep mountains, on the other hand, there is often hardly any place for the formation of a floodplain of an appreciable size along swiftly running streams or rivers, often squeezed into narrow gorges or ravines.
In the foothills, where the water flow slows, relatively large floodplains can be formed, which are differentiated into a shifting and meandering river or stream bed, leaving behind partly or fully cut-off backwaters, oxbow lakes and high-water whirlpools, whose natural land-filling can be reset by disturbances during high floods (Fig. 3b); they reset these habitats to or near to their initial stages. In spite of a common strong water flow regulation by dams and canals, high spring floods occur in floodplains of both small and large rivers fed with water from melting snow in spring. Examples of such rivers are the Danube, Rhine, Rhone or Ebro. For the lower reaches and delta of the Danube, for example, the high-water period can extend into the summer months when the water from melting snow combines with water from heavy June or July precipitations in the Danube catchment area, especially in the Alps and Carpathians. Such heavy floods occurred, e.g., in June and July 1966 and August 2002.
Generally less dynamic (with notable exceptions such as the floods in England in 1997) is the hydrological régime of rivers and streams with completely or prevailingly lowland catchment areas with little or no snow accumulating during winter. Here, it is the actual precipitation in the catchment area that controls the water table and flow velocity. As a result, fluctuations of these hydrological parametres are less regular here than in watercourses fed with water from abundant snow in the mountains. Nevertheless, the resulting diversification of the floodplains of predominantly lowland watercourses is similar to that of the previous type of floodplains. Both sharp boundaries between land and water, and smooth ecotones between them are abundant in natural floodplains (Naiman and Décamps 1990). Specific for floodplains of watercourses in the Mediterranean parts of Europe is a regular alternation of relatively high-water periods in the rainy winter and low-water periods in the dry summer (Britton and Crivelli 1993).
Highly dynamic and exposed to frequent disturbances is the herbaceous and shrub vegetation fringing river and stream banks (Prach et al. 1996, 2003). Sites exposed to more or less frequent disturbances by flood water are colonized by a mosaic of temporary stages of a hydrarch succession of wetland herbaceous vegetation, from submerged and floating-leaved hydrophytes (Fig. 2f) through helophytes (e.g., Phalaris arundinacea or Glyceria maxima) to marsh plants such as sedges (Carex spp.) (Fig. 2c). Calmer floodplain sites are occupied, as a rule, by forest vegetation (Penka et al. 1985, 1991). In eutrophic habitats, they are dominated by softwood trees (e.g., Salix, Populus) at low elevations above the normal water table, while hardwood trees (e.g., Fraxinus, Ulmus, Quercus) dominate at higher elevations (Fig. 2d). In oligotrophic habitats, the dominant softwood trees tend to be Alnus and Salix. Over large areas of all European floodplains, alluvial forest has been forced to give way to plantations of fast-growing trees (e.g., introduced cultivars of Populus or Eucalyptus), or to more or less intensely managed alluvial grassland. Local drainage has even enabled crop cultivation at places. Large areas of floodplain forests are still preserved in various European floodplains (e.g., the Rhine in Alsace, the Danube near Vienna, the Morava/March and Dyje/Thaya rivers in southern Moravia, western Slovakia and Lower Austria, the Drava and Sava rivers in Slavonia). Diverse algal vegetation as well as species-rich assemblages of fish and amphibians occur in the still preserved alluvial backwaters, oxbow lakes and pools (e.g., Pechar et al. 1996; Prach et al. 2003).
Unfortunately, only few floodplains or their segments have preserved their natural or semi-natural structure and dynamics, as a result of large-scale straightening and channelization not only of larger rivers, but also of small watercourses all over Europe during the last 200 years (Purseglove 1988; Haslam 2008). As to large European rivers, segments of natural or near-natural floodplains remain, e.g., along the Danube, Rhine, Elbe and Loire and some of their tributaries. On some sites, attempts have been made to restore natural floodplain dynamics, e.g., on the Rhine in Alsace (France), on the Morava river in the Czech Republic, Austria and Slovakia or on the Elbe river in Germany.
The deltas and estuaries of large European rivers represent highly complex systems of habitats characteristic of floodplains, also with sand bars, tidal mudflats and lagoons with water that shows a gradient of salinity depending on the ratio between freshwater and saltwater inputs to each particular zone or site of the delta or estuary at particular phases of the hydrological régime of the respective river and of the tidal régime of the respective sea (e.g., Rodewald-Rudescu 1974). This variation in environmental conditions is reflected in a high biodiversity in deltas and estuaries, unless they have been heavily modified by water engineering. The largest and relatively well-preserved European deltas are those of the Danube and Volga, but valuable wetlands are also found in the Rhone and Ebro deltas or in the estuaries of the Rhone, Elbe, Oder or Loire.
Human-made wetlands comprise diverse types of human-made biotopes created for various purposes. For instance, drainage or irrigation ditches may constitute the last remnants of formerly large wetland areas. Paddy rice fields can be found in southern Europe where most of them occupy former natural wetlands. Buffer zones involving natural wetlands, wetlands created for capturing agricultural runoff, and constructed wetlands designed for wastewater treatment, have received increasing attention since the 1980s from both the technological and scientific points of view and presently occur in most European countries (Vymazal et al. 1998). Artificial lakes of all sizes have been created for various reasons in river floodplains. Provided they are in a good ecological state, their littoral zones have the potential to host littoral and submerged vegetation that is very similar to that of natural lakes.
In terms of area, shallow lakes created for fish rearing, or fishponds, probably represent the largest proportion of artificial wetlands in Europe. They have been constructed since the Middle Ages in countries of Central Europe as well as France, Serbia and Ukraine. In the Czech Republic, which does not have large natural lakes, fishponds represent about 50% (or 560 km2) of the country’s total wetland area. Although the fishponds were constructed mostly for fish rearing in the course of history (Šusta 1898), they have successively become harmonious parts of the surrounding landscapes and have evolved into ecosystems in many respects similar to natural shallow lakes (Dykyjová and Květ, 1978; Kořínek et al. 1987; Kubů et al. 1994; Pechar et al. 2002). In addition to fish production, they have provided numerous additional ecosystem goods and services such as flood control, water retention, modification of local climate and enhancement of biodiversity (Hejný et al. 2002; Pechar et al. 2002). These benefits were the main reasons for declaring the well-preserved fishpond-rich landscape of the Třeboň Basin (Czech Republic) a biosphere reserve by UNESCO (Květ et al. 2002).
Other inland wetlands
Temporary freshwater pools, ponds and marshes (both natural and human-made) are also abundant and represent, together with rice fields, a well recognized type of habitat and ecosystem (European Commission 1999) as they contain a diverse and distinguished flora and fauna. Particularly valuable plant species, often with a short life cycle, occur on emerged bottoms, shores or banks of both standing and running waters. When inundated, their propagules can survive long periods in a dormant state (see, e.g., Hejný 1960, 1969; Hejný et al. 1998, for weeds of rice fields and temporary vegetation of emerged fishpond bottoms and shores in Central Europe). In areas with a continental climate and salt-rich subsoils, the temporary wetlands exhibit a slight to medium salinity (e.g., Löffler 1982 for the so-called “Lacken” in the Seewinkel near Lake Neusiedlersee).
Wetland research and university teaching of wetlands ecology
Wetland research has a relatively long tradition in Europe. It has developed simultaneously at several scientific centres of marine and coastal ecology, limnology, telmatology (i.e., peatland science) and aquatic botany or zoology, especially since the times of the IBP (International Biological Programme, 1965–1974, see Westlake et al. 1998). Research centres where wetlands are studied can be found in most European countries. The level of wetland research (both fundamental and applied) is generally high at these centres, although the emphasis on various aspects of wetland ecology varies among them. European authors have either written, edited or significantly contributed to several textbooks or handbooks devoted to wetland ecology and management (e.g., Moore and Bellamy 1974; Gore 1983a, b; Moore 1984; Whigham et al. 1993; Paavilainen and Päivänen 1995; Fustec and Lefeuvre 2000; Vymazal 1995; Vymazal et al. 1998, 2006; Westlake et al. 1998; Charman 2002; Jeglum and Hooijer 2006; Haslam 2008). University education in wetland ecology is carried out within the curricula of a number of European universities, albeit not always in courses so entitled. Quite often, for example, the courses of limnology deal also with wetlands.
The Society of Wetland Scientists has relatively recently (in 2004) established its European chapter whose annual meetings (since 2006) aim at becoming a representative forum of European wetland scientists. The Wetlands Working Group of INTECOL has a broad base of collaborating wetland scientists who gather at international conferences on wetlands every 4 years. Three out of eight of these conferences held so far took place at the European wetland research centres at Třeboň (Czechoslovakia), Rennes (France) and Utrecht (The Netherlands).
Wetlands have been exploited and/or traditionally managed for various purposes since the beginning of the human settlement in the area (Haslam et al. 1998; Löffler 1990 and another 10 chapters (17–26) in Patten 1990, 1994). Many of the traditional uses such as fishery, harvesting of reed, mowing of wet grasslands, hunting and floodplain forestry have locally been preserved till today. These uses are considered sustainable provided their extent and technology comply with the carrying capacity of the ecosystems (Verhoeven et al. 2006). The same applies to one of the recent wetland uses by modern society, i.e., soft tourism. The intensity of research also has to be adjusted to the ecological sensitivity and resilience of each wetland studied.
Along with the above-mentioned (potentially sustainable) uses, various types of unsustainable wetland uses occur (Williams 1990). They cover peat and sand or gravel extraction and drainage for agricultural or forestry use. Although these uses have occurred through the history of human settlement in Europe, both their extent and impact have dramatically increased since the middle of the nineteenth century.
In many densely inhabited regions, most nutrient-rich waterlogged sites with mineral soil as well as fens were drained for agriculture quite early (alluvial sites along the River Po, Italy are among the earliest documented). All uses involving drainage lead to decreased beta-diversity in the flora and fauna (e.g., Laine et al. 1995; Vasander et al. 1997). Agricultural use also has lead to the loss of the carbon sink function—subsidence and a gradual loss of soil organic matter. Agricultural use involving both drainage and nutrient enrichment affects the sites with organic soils differently from those with mineral soils. While sites with mineral soils rapidly respond by changes in plant diversity, on sites with organic soils the peat or humus mineralisation enhances the CO2 efflux from the soil. Forestry use is somewhat less aggressive, since a plant cover with new C inputs into the soil is maintained for most of the time. The effect of forestry use on the C sink function varies with wetland type and climate, often leading to C loss from the soil, but in some relatively poor peatland types in Fennoscandia, a C sink may be maintained. Forested drained sites are somewhat easier to restore than the more disturbed agricultural sites.
Extraction of peat is always linked with a lowering of the water table on respective sites. Thus, apart from direct loss of the peat, another type of loss comes into question, namely that due to mineralization of the peat, which leads to further subsidence of the extracted mire surface. The speed of this subsidence varies from place to place according to both climatic and soil factors. It may also be substantial in peatlands used for intensive agriculture. The most striking examples of peat subsidence originate from regions where peatlands were drained in the late Middle Ages or soon afterwards (the Netherlands, East Anglia). In the Netherlands, subsidence mainly through oxidation and compaction occurred at a rate of about 0.3 m per 100 years. Consequently, the main embanked rivers were soon flowing some 1.5–2 m above the general level of the peat, a difference that has now increased to 3.5–4 m (Williams 1990).
Both riverine and lacustrine wetlands occurring in Europe tend to be affected by large-scale eutrophication (e.g., Phillips 2005) occurring under the impact of agricultural and forestry management of their catchments, effluents from human settlements, feedlots and industrial plants, and atmospheric deposition, especially nitrogen compounds. Most European wetland restoration projects aim at mitigating the effects of eutrophication of various types of wetlands—from floodplains and shallow lakes to wet grassland—on their ecosystem structure and functioning. Most successful are such projects that succeed in reducing the input of organic pollutants and mineral nutrients from whole catchments, also including the most important point sources of these substances or intentional fertilization of wetland habitats (Jörgensen and Löffler 1990; Eiseltová 1994; Eiseltová and Biggs 1995; O’Sullivan and Reynolds 2004, 2005; Verhoeven et al. 2006).
Other threats include land filling, building of navigation canals, accelerated water discharge caused by straightening of watercourses, permanent inundation by reservoirs, fragmentation of residual wetland biotopes, pollution. In addition, an unproved assumption that all wetlands are important sources of greenhouse gases (especially CH4), and therefore speed up climate change, may be misused as an argument for further drainage of wetlands.
Wetland conservation and restoration
In spite of many destructive uses of wetlands in Europe, important activities exist, whose aim is to protect or even restore wetlands. At the international level, the European Union has signed international conventions aimed at nature protection, including the Ramsar Convention on the Conservation of Wetlands (http://www.ramsar.org), the Bonn Convention on Migratory Species (http://www.cms.int), and the Rio Convention on Biological Diversity (http://www.biodiv.org/convention/default.shtml). To date, the Ramsar Convention is the primary basis for the conservation of the most valuable wetlands in Europe. There are 47 contracting parties to the Ramsar Convention in Europe (of the world’s total of 160), which have designated 898 European wetlands of international importance. These 898 sites represent about 50% of the total number of all Ramsar sites worldwide. However, these sites occupy only 14% of the area of all Ramsar sites of the world. This fact reflects the fragmentation of the still existing wetlands but, at the same time, also a fairly strong public awareness of wetlands values. The Montreux record, listing Ramsar sites exposed to actual or potential unfavourable changes in the past, present or future times, comprises 23 European Ramsar sites.
Other administrative and legislative tools have strengthened or enlarged the impact and extent of wetlands conservation and wise use at the EU level and in its particular member countries. At the European level, the Bern Convention (http://www.coe.int/T/E/Cultural_Co-operation/Environment/Nature_and_biological_diversity/Nature_protection/) has led to the development of policy and action in nature conservation in Europe. It lists protected species and requires its parties to prevent the disappearance of endangered natural habitats including wetlands. Within EU legislation, the Birds Directive (79/409/EEC) and the Habitats Directive (92/43/EEC) have been promoted to rectify or reduce damage to European natural habitats and associated species. The Birds Directive is aimed at the protection of endangered bird species through designation of areas where these species are given special protection. Following the same principle, the Habitats Directive is aimed at the conservation of wild fauna and flora on the European territory on the basis of protection of their natural habitats. Following the criteria set out in the directives, each Member State must draw up a list of sites hosting the wild species of fauna and flora and put in place a special management plan to protect them, combining long-term preservation with economic and social activities, as part of a sustainable development strategy. Special protection areas for birds (SPAs) and special areas of conservation (SACs) are designated according to the Birds Directive and the Habitats Directive, respectively, and approved by the European Union to become part of a European Ecological Network called Natura 2000. By December 2008, 24,831 sites belonging to 27 European State Members covering 859,411 km2 are included in this Network (http://ec.europa.eu/environment/nature/natura2000), which represents 17% of the whole European territory. Wetlands are particularly important in the Natura 2000 network.
Indirect protection to a variety of habitats also comes from EU Directives that regulate water quality, especially the Water Framework Directive (2000/60/EC). Additional policies concern coastal and marine areas (Airoldi and Beck 2007).
The European Directive (2007/60/EC) on the assessment and management of flood risks has recently been established to reduce adverse consequences associated with floods on human health, the environment, cultural heritage and economic activity. For this, the European member countries should establish flood risk management plans based on flood hazard maps and flood risk maps at the scale of the river basin by 22 December 2015. This European law clearly states that a preliminary flood risk should be assessed by December 2013 considering potential impacts of climate change, or use already existing management plans on the occurrence of floods and the role of floodplains with respect to this risk, both for inland (river and lake floodplains) and coastal areas. While the objective of preventing and buffering damages to human health and economic activity may be encouraged, an adequate integrated conservation of floodplains with their role as valuable natural ecosystems is not assured (Comín et al. 2008).
Of the Europe-based organisations taking care of the scientific basis for wetlands management, conservation and restoration, one should mention Wetlands International (http://www.wetlands.org) as a global science-based non-profit organisation dedicated solely to wetland conservation and sustainable development. Wetlands International, whose office is in Ede, The Netherlands, closely cooperates with the Ramsar Secretariat at Gland, Switzerland.
The basic principles of wetland conservation, restoration and creation are described, e.g., by Bobbink et al. (2006). There are numerous examples of successful conservation and restoration measures in European wetlands of all types (e.g., Gilman 1994; Janda and Ševčík 2002; Bragg et al. 2003; Farrel and Doyle 2003; Vasander et al. 2003). They are based largely (though not solely, see Sliva and Pfadenhauer 1999; Gorham and Rochefort 2003) on the successful conservation or restoration of the respective wetland’s hydrological regime. New wetlands have been spontaneously developed or created, e.g., in the littoral zones of artificial lakes (e.g., Rajchard et al. 2008) or in association with reclaimed Dutch polders, such as Wolderwijd en Nuldernauw adjacent to South Flevoland polder (Anonymous 2003). The European Comission promoted the long-term programme Life-Nature that included many cases of wetland restoration during the last 15 years (D.G. Environment-EC 2007). This programme was responsible for the restoration of coastal and inland wetlands all around Europe and elsewhere, as it involved also neighbouring countries. It is still operating, active and stimulating the cooperation of managers, stakeholders, scientists and landowners. International and national legislation primarily aimed at improving the quality of surface waters (e.g., the Convention on the Protection of the Rhine (http://www.iksr.org), Convention on the International Commission for the Protection of the Elbe (http://www.ikse-mkol.org), the Danube River Protection Convention (http://www.icpdr.org), as well as various river and floodplain restoration projects have indirectly contributed to the conservation of existing wetlands.
As 60–90% of the European wetland area disappeared during the last century (Mitsch and Gosselink 2000), there is an important deficit of wetlands with respect to earlier times and also to potential wetlands recovery. A practical limit to the official approach to wetland protection in Europe is prioritising between wetland protection and restoration on the one hand and agriculture and tourism exploitation on the other. Frequently, immediate economic profit prevails over interests of nature conservation and restoration. The recent incorporation of further countries into the European Union could be an opportunity to integrate these wetland activities into the socio-economic development of these countries.
Sea level rise
It is generally accepted that global climate change will bring about a rise in water level in all seas. IPCC models estimate the global average rise at about 3–4 mm per year. The highest sea level rise is expected in the Arctic region, thus affecting also the northern coast of Europe (Scandinavian countries and Russia) (Meehl and Stocker 2007). The local sea level rise will be further modified by vertical land movement. Taking vertical land movement into account gives slightly larger sea level rise projections relative to land in the more southern parts of the UK where land is subsiding, and somewhat lower increases in relative sea level for the north. We have, for example, derived projected relative sea level increases for 1990–2095 of approximately 21–68 cm for London and 7–54 cm for Edinburgh (5th–95th percentile for the medium emissions scenario) (Lowe et al. 2009). The sea level rise will also bring about increased frequency and amplitude of extreme sea level events. This increase is determined by the atmospheric storm intensity and movement, and coastal geometry. Within Europe, increases in extreme sea level events are to be expected along the continental North Sea coast (Christensen and Hewitson 2007), thus affecting costal areas of all countries from Denmark in the North to northern France in the South.
Temperature and precipitation
According to most regional climate change models, annual mean temperatures are likely to increase more in Europe than is the global mean increment. In addition, spatial and temporal differences in the intensity of the warming will be substantial. In winter, the largest warming is likely to take place in northern Europe. In summer, on the other hand, the maximum temperatures are likely to increase most in southern Europe (Christensen and Hewitson 2007).
Annual precipitation is expected to increase in northern Europe but decrease in most of southern Europe. The seasonal patterns may, however, be more important than average annual sums of precipitation. In northern Europe, the increased annual precipitation will be caused mainly by increased precipitation in the winter months. Nevertheless, water input from increased precipitation will be offset by the effects of higher temperatures. Because of higher winter temperatures, the snowy season is likely to be shorter and the snow depth will probably decrease over much of northern Europe. Also, the increased evapotranspiration owing to higher summer temperatures is expected to override the increased summer precipitation. Consequently, summer drought will probably be the most important stressful effect of the changing climate on inland wetlands. Its risk will penetrate further northwards in comparison with the present-day situation. The frequency and intensity of summer droughts is most likely to increase from the north to the south.
More frequent occurrence of extreme meteorological conditions (temperature, precipitation, air humidity) is envisaged and may be more important than the overall trends. Depending on the local and regional climate character, the resulting meteorological events may include strong winds, heavy rains possibly followed by floods, a greater frequency of extremely high temperatures for a given region, and longer periods without precipitation. Also, events of low frequency but intense ones (e.g., droughts) are important phenomena related with climate teleconnections (e.g., Atlantic and tropical air pressure oscillations such as El Niño Southern Oscillation) that regulate the dynamics of many inland wetlands all around Europe (Rodo et al. 1997; Rodo 2003).
Anticipated effects of climate change on wetlands
One of the first assessments of possible climate change effects on wetlands can be found in Boer and de Groot (1990). Our considerations in the further text are largely in agreement with their assumptions given on pages 41–46. The authors regard the sea-level rise, changing air and water temperatures, and evaporation to precipitation ratios as the main driving forces affecting wetlands.
Sea level rise
Among various impacts of the ongoing climate change, the sea level rise will probably be the most important factor affecting coastal wetlands (mainly mudflats and salt marshes) because of the strong dependence of these habitats on water-level fluctuations and tidal régimes. It has been suggested that the projected sea-level rise could cause the loss of up to half of the present European coastal wetlands, with some of the largest losses expected to occur around the Mediterranean and Baltic Seas (Airoldi and Beck 2007, and references therein). With a higher sea level, salt water will penetrate deeper into estuaries, converting a part of brackish aquatic and wetland ecosystems into saline ones. At the same time, some freshwater wetlands connected with the sea will become brackish.
In a natural coastal zonation, the sea level rise would just cause a landward shift of all wetland zones. This is, however, unlikely to happen in the densely populated coastal areas of Europe because most of the suitable upland areas are already used by people for various purposes. The coastal wetlands may therefore be sandwiched and squeezed between the shifting boundary of the shoreline on the seaward side and the fixed boundary given by the current land use on the landward side (Doody 2004). In some areas along the Atlantic coast (mainly in The Netherlands), the advance of sea water will be resisted by building new or strengthening existing barriers. In these areas, the hydrology of the remaining wetlands would be fully controlled by the associated technical measures and the space left for wetlands will again depend on priorities of land use. Provided the coastal wetlands are given sufficient priority, their anticipated loss can theoretically be minimized or compensated for by political and socio-economic tools such as wise and timely land-use planning and consequent management measures. This will possibly happen in some large protected areas such as the Wadden Zee (Fig. 1), which has received continued attention by both nature conservationists and national and local administrations (Hofstede 2003). Outside the strictly protected areas, we must fear that the area of coastal wetlands will be forced to shrink.
This will be the case if a defensive short-term strategy (reactive strategy according to the Millenium Ecosystem Assessment, Finlayson et al. 2005) is followed to avoid sea water intrusion into the coastal zone. Such an approach will be sustainable neither economically nor ecologically (with respect to the preservation of a healthy coastal zone). As an alternative, an adaptive long-term strategy (proactive strategy according to Finlayson et al. 2005) can be adopted that should include re-allocation of land uses and re-definition of services provided by the coastal zone ecosystem. A socio-economically acceptable compromise would probably be a mixed strategy, establishing defensive structural measures where important social assets are established and let coast dynamics determine future distributions of ecosystems and land uses. In the socio-economic context, one should consider that coastal wetlands provide a high value (Martínez et al. 2007) to coastal zone inhabitants, including protection against storms and other impacts of sea level rise induced by climate change.
It is commonly accepted that the anticipated increase in temperature will considerably affect both coastal and inland wetlands. The temperature increase will directly affect biological processes such as photosynthesis, respiration and transpiration. It will affect the biological processes also indirectly through changed physico-chemical properties of ecosystem components, such as changed solubility of various substances in water. In addition, the increased evaporation to precipitation ratio is expected to lead to lowered water levels and increased probability of drought.
The described climate development is generally unfriendly to inland wetlands, with increased summer dryness being the key factor. It will translate into wider water level fluctuations (both seasonal and irregular) and a generally greater water shortage in most wetland types over much of Europe. Maintenance or restoration of a hydrological régime ensuring the continued existence of any wetland will gradually become more and more difficult as it will require water supply from larger catchments or infiltration areas.
Boer and de Groot (1990) argue that the temperature rise and increased evaporation to precipitation ratio could have a profound impact on inland wetlands because of internal eutrophication, salinization, dessication and invasion of thermophilous species. They conclude that the isolation of individual wetlands can increase because of the fragmentation of biocorridors as a result of water shortage.
Riverine wetlands, including those in estuaries and river deltas, may be reduced in area, especially in the South European and inland East European regions, as a result of decreased water discharge in rivers and streams in the growing season. The same is true for wetlands associated with lakes and other standing waters, where the water shortage will be associated with accelerated land-filling and a consequent establishment of terrestrial species of plants and animals. This development may eventually lead to a shift towards terrestrial ecosystems. The reduced water volume will also result in higher concentrations of dissolved nutrients and suspended solids in both running and standing waters. Additionally, higher temperatures will promote the mineralisation of soil organic matter resulting in an increased availability of nutrients in wetland soils.
In peatlands (both bogs and fens), the anticipated water shortage in summer will lead to lowered water levels and, thus, oxic conditions in deeper surface layers, but also increasing dryness of the surface peat. Oxic conditions allow for an increased rate of decomposition of organic matter contained in peat or peaty soil. But if the water level falls deep enough, dryness may impede decomposition in the topmost layers (Lieffers 1988; Laiho et al. 2003). Many anticipated effects will depend on the range of water level fluctuations. If dry and wet years alternate, increasing the instability of water levels, the systems will enter a stage of “constant disturbance”, with a limited number of plant species tolerating both extremely wet and extremely dry conditions, forming distinct community compositions (Laitinen et al. 2008). In such cases, C sequestration can cease, and extensive C loss from soil may take place, since the “best” C accumulators disappear from the plant community, and decomposition during dry periods may compensate and even exceed any accumulation during wet years.
An essentially similar situation has been found during dry years in contemporary mire ecosystems (e.g., Schreader et al. 1998; Alm et al. 1999; Moore et al. 2002). Leaching of dissolved organic carbon and nutrients may be accelerated as in other wetland types. More or less permanently lowered water levels, on the other hand, will lead to a “forest succession” with increasing abundance of shrubs and trees (e.g., Laiho et al. 2003), except for bog sites so poor in nutrients or so cold that increased tree growth is not feasible (Vasander 1982). Succession will continue until a new equilibrium between the vegetation composition and the new water level régime has been achieved, which may take several decades. This development will lead to changes in runoff patterns, and may eventually lead to decreased leaching of DOC and some elements such as K, whereas the leaching of other elements, such as Ca and Mg, may increase. The drier systems will become more acid. They will lose most of their specialized wetland vegetation, which will be replaced by common forest species. This change will lead to lowered beta-diversity (e.g., Laine et al. 1995; Vasander et al. 1997). In most cases, C loss from soil can still take place, even though its rate may slow down as the decomposition potential of the exposed peat decreases (Jaatinen et al. 2008). On the most productive bog sites with increased tree growth and, consequently, litter inputs, but still relatively poor substrate quality, the C sink function may continue, at least at higher latitudes. While C losses are likely to increase in temperate and southern boreal regions, C sequestration may increase in subarctic regions. The palsa mires (with local permafrost formations), specific for cold regions, may disappear, and permafrost melt in the northernmost Scandinavia and northern Russia, especially, may lead to partly unpredictable changes.
As a result of temperature changes and the associated changes in water availability, the latitudinal zonation of different peatland types may change considerably. However, predictions differ according to the presumed driving forces. Crawford (2008) stresses that the greater increase temperatures in winter than in summer will lead to an expansion of oceanic climate into northeastern Europe and Siberia. This in turn may support a southward expansion of Sphagnum-dominated mires in spite of a northward expansion of boreal forest, as it has been commonly assumed.
Prolonged dry periods that have been observed in southwestern Europe since the second half of the twentieth century can change the spatial distribution of wetland habitats, particularly inland wetlands. An example is Lake Gallocanta in Aragon, NE Spain, a playa lake in a closed endorheic basin. It serves as a climatic sensor with its water level mostly fluctuating in accordance with its climate regulated water balance. More frequent and prolonged dry periods have been observed in Lake Gallocanta in accordance with global climate change (Rodo et al. 1997). More frequent and prolonged dry periods will turn this temporary wetland, an area of high biodiversity at the European scale, into a dry salt pan (Comín et al. 1991).
Fire hazards will increase especially in summer-dry Mediterranean wetlands as well as peatlands with a dried-out surface vegetation and peat layer.
Further interacting effects
Land use and land cover may significantly affect climate at regional and local scales. Recent modelling studies also show that in some instances these effects can extend beyond areas where land cover changes occur, through the teleconnection processes. (Christensen and Hewitson 2007). For the fate of wetlands in a changing climate, various interacting effects may become more decisive than the anticipated increased impact of the atmospheric greenhouse effect. Drainage of wetlands is a more potent driver of local climate change than the changed greenhouse gas balance. The reduced transformation of incoming solar radiation into latent heat of evaporation leads to increased overheating of dry surfaces. This holds not only for “reclaimed” wetland areas, but for drained and urbanised areas in general (Denman and Brasseur 2007). An extreme situation is represented by urbanised areas that create urban heat islands associated with considerable warming (Arnfield 2003). The rapid urbanization of the European landscapes (Antrop 2004) cannot leave the mostly fragmented European wetlands unaffected.
Socioeconomic trends resulting from the public perception of climate change may significantly interact with the direct impacts of the changing climate. The socio-economic priorities are likely to differ between regions exposed to different main impacts. Preservation of carbon storage is a key issue for the northern part of Europe, where large areas of peatlands occur. Extreme meteorological events and their consequences such as downpour rains followed by floods are likely to be perceived the most in Central and Western Europe. They may promote public requirements for technological (hard) flood control measures resulting in faster water discharge, which would threaten the hydrology of existing wetlands. Continental and south European wetlands will probably suffer most from water shortage. Consequently, competition for water between agriculture and urban land use on the one hand, and environmental protection on the other hand, may substantially reduce the water supply to wetlands.
Need for a change in the perception of wetland values
In recent years, the scientific community has contributed to the formation of environmental policies by synthesising scientific knowledge in the form of background materials addressed to decision-makers. Apart from scientific knowledge, these documents incorporate the elements of strategic considerations (scenarios, strategies) that facilitate the assessment of the limits and/or alternatives of future development (e.g., Finlayson et al. 2005). The principles of nature and ecosystem conservation, which are generally accepted as social priorities, have, under European conditions, the chance to raise funds for maintaining the desirable state or restoration of valuable sites including wetlands. The financial means invested in this way have already brought visible results (http://ec.europa.eu/environment/life).
Although there has been a considerable improvement in the human attitude toward wetlands over the last decades (especially in areas where most wetlands had previously been lost), rapid climate change and anticipated impacts call for further change in the human perception of wetlands. The impact of climate change on biodiversity has long been of widespread concern. In addition, however, it is worth considering that there is a feedback relationship between wetland ecosystems (the same as any living systems) and their environment, including climate. This becomes particularly important for large wetland areas such as boreal peatlands and deltas of large rives. This feedback relationship encompasses not only the greenhouse gas balance, which is in the focus of attention today, but also the climate stabilization through the air-conditioning effect of evapotranspiration. In addition, the specific features of wetlands, such as their hydrology, predispose them for playing an important role in large landscape complexes, where their impact considerably surpasses their physical boundaries. This statement applies mainly to such hydrological functions of wetlands as water retention on the one hand and flood mitigation on the other.
It must be taken into account that wetland function can be performed if the ecosystem is well-established and that long-term water retention and flood mitigation in floodplains requires a dynamic floodplain (Comín et al. 2008). Otherwise, the negative impacts of artificially created infrastructures for water retention can override the water retention function and eliminate it in the long term.
Biodiversity support is commonly listed as one important wetland value (Mitsch and Gosselink 2000; Gopal et al. 2000, 2001). Under biodiversity, we understand not only a variety of species, but also that of their life forms, habitats and niches occupied by them. In this respect, European wetlands are highly diverse ranging from acidic and nutrient poor bogs on the one hand to highly fertile and productive wetlands in estuaries, salt marshes and freshwater littoral or riparian wetlands on the other. Some wetlands (such as many mires and springs) are island ecosystems sensu MacArthur and Wilson (1967). They serve as refuges of rare and relic species and their relative isolation in the landscape may promote microevolution of specialized phenotypes. Ecotonal wetlands (such as littoral reed belts or riparian wetlands) host species both from adjacent larger-scale ecosystems and species that are confined only to the ecotone itself (Naiman and Décamps 1990). Biodiversity has also a time dimension associated with water level fluctuations (Hejný 1957; Hejný et al. 1998). When the water table sinks to or below the ground surface, specialized plant species often appear, which can survive long-term flooding in the bank of dormant propagules. Examples are the communities of emerged shores or bottoms of lakes, pools and ponds (e.g., Hejný and Husák 1978; Rejmánek and Velásquez 1978; Hroudová 1981; Šumberová et al. 2005, 2006). Their preservation in European landscapes is enabled by harmonising water level fluctuations with the species requirements during their life cycle.1
Rajchard et al. (2008) have suggested that littoral zones of quarries and sandpits formed by surface mining provide oligotrophic habitats for wetland species that are endangered and disappearing from the surrounding eutrophicated wetlands in intensely managed areas. In reality, the fulfilment of this potential depends on other factors such as shore morphology and intensity of recreational use.
To date, beta diversity of wetlands seems to be broadly accepted by the European population. Since it is reflected by both national and EU legislations, it frequently provides the most powerful argument for protecting a particular wetland site. This argument will become even stronger in the near future as a result of the implementation of the EU Habitat Directive (92/43/EEC). Because of the administrative feasibility of this approach, biodiversity is used in advocating protection of wetland sites whose other values (see the text below) are obvious but are not protected by legislative tools. Such substitute arguments for wetlands conservation are often only unwillingly accepted by the predominantly technocratically oriented decision-makers. Development of complementary legislative and administrative tools, based on the assessment of all wetland functions in the landscape, is a pre-requisite of establishing a more balanced basis for sustainable decision-making concerning wetlands.
Greenhouse gas balance
The greenhouse gas balance is currently in the centre of attention of both the scientific community and the general public because it is considered to be one of the main causes of global climate change (Janssens et al. 2005). In contrast with terrestrial ecosystems, wetlands emit methane as an important component of their greenhouse gas budget (Segers 1998; LeMer and Roger 2001). The greenhouse gas balance of a wetland is the outcome of the rate of net CO2 uptake (CO2 sequestration) on the one hand and the rates of CH4 and N2O efflux (greenhouse gas emissions) on the other hand. This outcome, expressed as radiative forcing, may be either positive or negative depending on the rates of the processes involved. The dynamics of greenhouse gas exchange is largely determined by specific site conditions including hydrological conditions, soil type, vegetation, and management and meteorological and climatic conditions. Depending on meteorological conditions, wetlands (the same as other ecosystems) may act as CO2 sinks in some periods and as sources in others. The emissions of CH4 and N2O from wetlands are similarly variable in time.
Compared to other terrestrial ecosystems of Europe, especially forests (Janssens et al. 2005), less information is available on the greenhouse gas balance of wetlands. Among wetland types, Phragmites-dominated wetlands are understood relatively well (Brix et al. 2001). Case studies have been published for boreal sedge fens (Aurela et al. 2004, 2007), temperate wet grasslands (Hendriks et al. 2007; Dušek et al. 2009) and constructed wetlands for wastewater treatment (Picek et al. 2007). Special attention has been paid to the determinants of methane dynamics (Kaki et al. 2001; Kankaala et al. 2003, 2004; Rinne et al. 2007).
There is insufficient information needed to provide simple guidelines for management aimed at achieving a positive balance of greenhouse gases in the existing variety of wetland types. Yet, the current knowledge provides a basis for some important generalizations. Of all natural wetland types, peatlands are by far the most important ecosystems affecting the global balance of greenhouse gases. Peatlands globally represent a highly important store of carbon, sink for carbon dioxide, and a significant source (from the point of view of its importance for the greenhouse effect) of atmospheric methane. In general, nitrous oxide (N2O) emissions are small in natural peatlands (Joosten and Clarke 2002). In addition to live peatlands (mires), littoral wetlands with abundant plant cover, such as reed (Phragmites australis) dominated marshes in Central and North Europe, can be important sinks for carbon (Brix et al. 2001). Floodplains can play an important role by accumulating organic matter and carbon if floods are maintained and river-floodplain connectivity lets the plant communities (especially riparian woodlands) develop at an integrated ecohydrological rhythm (Cabezas et al. 2009).
Two types of impact considerably affect the greenhouse gas balance of wetlands: changed hydrology and nutrient enrichment. More frequent summer droughts increase the frequency of situations under which wetlands, especially peatlands, act as sources of CO2. At the same time, the CH4 emissions decrease. There is also evidence that peatlands “reclaimed” for agricultural use are releasing significant amounts of nitrous oxide (N2O) because they have become enriched with mineral nutrients including nitrogen. Long-term nutrient enrichment of wetlands with organic soils can also promote CO2 efflux (Zemanová et al. 2008). Eutrophication of permanent wetlands associated with standing waters can promote anaerobic decomposition processes including methane production.
Generally, it is important to consider that wetlands have been both taking up and releasing greenhouse gases continuously since their formation, and thus their influence on the atmosphere must be modelled over time. When this is considered, the sequestration of CO2 in peat outweighs CH4 emissions. In terms of greenhouse gas management, the maintenance of large carbon stores in undisturbed peatlands should be a priority, as recently pointed out by Joosten and Clarke (2002).
The air-conditioning effect of the water cycle has been recognised as one positive function of vegetated areas sufficiently supplied with water such as most wetlands (Mitsch and Gosselink 2000). Its importance has been demonstrated by a negative practical experience: drainage associated with changed land use has been recognised as one cause of climate change at local and regional scales (Christensen and Hewitson 2007; Denman and Brasseur 2007). Yet, no information exists as yet about its importance for associated changes in energy fluxes at the global scale.
Evapotranspiration of water from vegetated surfaces not only increases air humidity, but also cools surfaces from which water vapour evaporates by the amount of energy needed for vaporisation (latent heat). The evaporated water vapour then condenses in cool air or on cool surfaces, which thereby receive the energy of the released latent heat. In this manner, the evaporation and condensation processes have a double air-conditioning function—plant stands are cooled in the evapotranspiration process while heated are places where the water vapour precipitates. Evapotranspiration possesses a huge capacity to equalize temperature differences in time and space. This air-conditioning effect is associated with enormous energy fluxes.2 Kravčík et al. (2008) explain the effect of drainage on temperature extremes and the role of water and wetlands for mitigation of climate change.
The predicted more frequent summer droughts are likely to bring about a general decrease of evapotranspiration from most vegetated areas: the heat balance will be shifted towards an increased sensible heat flux (thus to a higher Bowen ratio) more often than now. The value of wetlands, as “oases” in dry landscapes, will therefore increase, provided they remain saturated with water. This is particularly important in river and stream floodplains, where the water supply to wetlands can partly be subject to human control.
The concept of ecosystem services may change the perspective, by which wetlands are perceived. By services, we mean different benefits or goods, that wetlands (or any other ecosystems) provide to human welfare. In the Millenium Ecosystem Assessment (Finlayson et al. 2005), ecosystem services are defined as “the benefits people obtain from ecosystems”. Flood and drought mitigation, water purification, carbon sequestration, biodiversity refuge, production of commodities (fish, reed, wood, etc.), mitigation of storm effects, coastal erosion, and recreation are examples of services provided by wetlands (Mitsch and Gosselink 2000; Costanza et al. 1989; Jeník et al. 2002). These services can be expressed also in monetary values (Turner et al. 2008), although the introduction of economic tools would rise the level of complexity of these proposed ecological-economic systems. Financial evaluation of the services has at least three consequences: (1) wetland values are more understandable to technically-oriented decision makers and the general public; (2) the values are additive, i.e., we can approach the overall value of a wetland and compare it with other types of land uses—which should alter the decision-making process in land use towards wetland promotion; (3) the valuation of particular services will help to understand their importance from the perspective of contemporary human welfare.
If the overall value of ecosystem services per unit area is compared among different world biomes, wetlands are quite valuable, especially river estuaries (first position) and floodplains (second position together with seagrasses, Costanza et al. 1997). However, examples of complete calculation of wetland values are scarce. Successful practical solutions, based on the acknowledgement and application of this concept to the life of modern human society, remain rather a challenge for the future (Ruhl et al. 2007) than being successfully applied in contemporary decision-making processes. The reason is the difficulty to connect a broad multidisciplinary ecological approach, encompassing the quantification of different processes (fluxes of water, carbon, nutrients, etc.) related to ecosystem structure (biodiversity, land use), with the economic sphere (marketable and non-marketable services, values and prices, discount rates), that quite often results in an instinctive rejection of environmental issues.
Quantification of selected ecosystem services for a natural segment of the Lužnice river floodplain (South Bohemia, Czech Republic)
Value (USD ha−1 year−1)
Total retention volume 4.7 million m3, per ha 10,250 m3
Average point value 38, monetary value of one point per ha 8,000 USD
Sequestration of 2,029 t of C per ha, i.e. 7.54 t of CO2, market price of emission limit 20 USD per t
Catches: 3.4 t of fish in the area, 7.2 kg per ha, average price 5 USD per kg
183 ha of cut medows with production of 2 t of hay per ha, price 100 USD per t
Growth rate 5 m3 ha−1 year−1; 33 USD per m3 and 61 ha of the floodplain forest
Current status of wetlands:
There is a strong deficit of wetland number and area compared to wetlands existing in the early twentieth century. Consequently, the ecological restoration of wetlands is still a major activity to be performed to obtain the benefits of wetlands functions all around Europe.
Response of wetlands to anticipated climate change:
Sea level rise will be the main factor affecting coastal wetlands. It is also an opportunity for developing adaptive coastal management and recovering lost and degraded wetlands.
Wider water-level fluctuations will occur in most inland wetlands.
Human impact on wetlands (especially drainage) can strongly interact with or even prevail over the effects of climate change.
Preservation of carbon storage should be the priority in the northern part of Europe where large areas of peatlands occur.
Extreme meteorological events and their consequences such as downpour rains followed by floods are likely to be perceived most sensitively in Central and Western Europe. They may promote public requirements for technological (hard) flood control measures resulting in faster water discharge from respective catchments, which would threaten the hydrology of existing wetlands.
Continental and south European wetlands will probably suffer most from water shortage. Consequently, competition for water between agriculture and urban land use on the one hand, and environmental protection on the other hand, may substantially reduce the water supply to wetlands.
Need for a changed attitude toward wetlands:
The anticipated climate change imposes a threat to the current condition of European wetlands in addition to the historically existing and recently strongly increasing human impact.
At the same time, our facing the climate change encompasses an opportunity for developing adaptive wetland management and recovering lost and degraded wetlands.
Socio-economic appreciation of wetlands will be enhanced if the scientific community is able to develop a broadly accepted system of economic evaluation of wetland ecosystem services such as carbon sequestration, climate stabilisation, or flood mitigation.
This is true, e.g., for Coleanthus subtilis (Tratt.) Seidl, a tiny (3–11 cm tall) annual grass that occurs on the bare or almost bare soils of emerged lake or pond bottoms and shores after its dormant caryopses have survived a long-term flooding of the biotope. The life cycle of the shoots of this grass lasts only 4–6 weeks, and the plants usually flower and fruit in June and July. The reproduction and hence also survival of C. subtilis at each particular site of its occurrence is ensured by a periodical drawdown of the water table at that time of year. One plant can produce over 1,000 ripe caryopses. In Europe, its geographical range of occurrence is narrow, covering only Central Europe and within it especially the basin of Třeboň and adjacent areas in the Czech Republic and Lower Austria. C. subtilis has thus become one of the 434 plant species protected within the EU “Natura 2000” framework (Habitat Directive 92/43, Annex 2). It is also listed in the Red List of threatened plants of the Czech Republic and in the IUCN list of threatened plants. The central area of its European occurrence lacks natural lakes, but is rich in artificial fishponds where the water table can be set at a certain level at any time. For securing permanent occurrence of C. subtilis within this area, agreements have therefore to be made with fishpond owners as to the occasional maintenance of a low water table in the respective fishponds at the optimum of this species’ seasonal development. Such an arrangement can result in a certain loss of the fish crop in a fishpond whose water area is temporarily diminished by the drawdown (summer drainage is not any more a regular part of the fishpond management in Central Europe), and provisions have to be made for a financial compensation of this loss. Obtaining of reliable data on the occurrence of C. subtilis on a certain territory thus requires a period as long as several years. It is advantageous that the protection of C. subtilis at any site brings with itself the protection of the whole rather rare plant community colonizing the emerged pond bottom or shore. For a thorough treatment of the biology and ecology of C. subtilis see, e.g., Hejný (1969).
Let us consider an herbaceous wetland that evaporates 6 l of water per 1 m2 during a summer day. The solar energy consumed in evapotranspiration of 6 l is equal to 4.2 kW (latent heat of 1 l of water = 0.7 kW, or 2.45 MJ). This amount of energy represents an average 24-h flux of about 180 W m−2. Expressed per an area of 5 km2, the latent heat flux equals 900 MW, which is equivalent to the power output of a large electric power station.
Work on this paper was supported by the projects NPV 2B06023 and MSM 6007665801 of the Ministry of Education, Youth and Sports of the Czech Republic, 526/09/1545 of the Grant Agency of the Czech Republic and QH 82078 of the Czech National Agency of Agricultural Research. We warmly thank Hana Šantrůčková for helpful comments on the manuscript, Štěpán Husák for providing the information on Coleanthus subtilis, Václav Nedbal for techical help with the compilation of the map of European wetlands (Fig. 1), Jakub Brom for providing photographs in Fig. 5, and Ondřej Novák for technical help with the preparation of the manuscript.
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