It is estimated that the world lost at least 50% of its wetlands during the 20th Century (UNWWAP 2003, Davidson 2014). Some two thirds of the European wetlands have been lost in the same period (European Commission 1995), leading to a substantial decrease in the number, size, and quality of bogs, marshes, wet grasslands, and shallow lakes. Over a timespan of multiple centuries wetland loss is much higher because draining, conversion, and infilling of coastal and inland wetlands in Europe has been ongoing since at least Roman times (Russi et al. 2013). During that period wetlands were converted from uninhabitable, remote areas, with often harsh and unhealthy living conditions, into more productive, accessible, and human-friendly rural landscapes.
Exact figures of contemporary wetland loss in Europe are hard to find. For the period 1950–1985 wetland loss in six European countries was roughly estimated to lie between 55% and 67% (European Commission 2007), however without providing any supporting data or references. Based on Corine (the EEA’s “coordination of information on the environment”) land cover data, it was estimated that between 1990 and 2006 another 5% (1267 km²) of Europe’s marshes and bogs were lost (EEA 2010). On the other hand, coastal wetlands remained more or less stable and open waters had even increased by 4.4% (1581 km²) in the same period (EEA 2010), the latter probably mainly in the form of artificial water bodies such as new dam and water storage constructions (Acreman 2012). The European Habitats Directive now protects 47 different wetland habitat types and 290 species (plants, invertebrates, fish, amphibians, reptiles, mammals) linked to wetlands. In 2006 however, nearly two-thirds of the species and more than three-quarters of the habitats throughout the EU member states were in unfavorable conservation status (ETC/BD 2008). Even more worrying, for the Boreal and Atlantic regions where large areas of wetlands (used to) occur, none of the habitats was in a favorable status. The successive (2007–2012) assessment indicates a further decrease (European Commission 2015).
This ongoing loss and deterioration of European wetlands contrasts sharply with their well-known values for society, recognized decades ago (e.g., Thibodeau and Ostro 1981, Batie and Shabman 1982, Farber 1987, Costanza et al. 1989, Folke 1991, Gren et al. 1994). The TEEB-review study (“The Economics of Ecosystems and Biodiversity”) for water and wetlands (Russi et al. 2013) clearly mentions the major ecosystem services provided: flood protection, water supply, water purification, carbon sequestration, climate regulation, production of raw materials and food, tourism and recreation, aesthetic and cultural values. In 2012, United Nations Environment Programme (UNEP) called for urgent integration of the key role of wetlands into decision making, and the need for their future protection, restoration, and sustainable use as a vital component of the transition into a resource-efficient, sustainable world economy (http://www.unep.org/newscentre/Default.aspx?DocumentID=2697&ArticleID=9305&l=en). The importance of wetlands has, on various occasions, been recognized within the framework of the CBD (e.g., COP Decision X/28, UNEP/CBD/COP/DEC/X/28, 29 October 2010; COP Decision XI/23, UNEP/CBD/COP/DEC/XI/23, 5 December 2012; Message of the Executive Secretary of the Convention on Biological Diversity Braulio F. De Souza Dias of 2 February 2016, https://www.cbd.int/doc/speech/2016/sp-2016-01-28-wwd-en.pdf).
Because many benefits of wetlands are of nonmarket and public nature they are rarely represented nor defended in decision-making processes. Governments who want to develop an evidence-based policy on wetlands will rely heavily on the availability of scientific information. A first and essential step in this process is to make the consequences of different land-use scenarios as explicit as possible. In the present paper we follow a spatially explicit approach balancing past wetland losses with potential gains from restoration and associated ecosystem service benefits and translating this into restoration opportunities. This approach enables us to identify synergies and trade-offs between alternative land-use planning policies and restoration scenarios. We test this integrated approach in Flanders (northern part of Belgium), one of the most degraded wetland regions in Europe. We successively describe and discuss the following:
The region of Flanders is situated in the northern half of Belgium and covers 13,522 km². Bordering the North Sea and the Netherlands, the area is rather flat, partly reclaimed from the sea (polders), and large parts are dominated by wide river valleys and a dense network of slow-running watercourses. The highest point only reaches 156 m above sea level. It has a maritime climate with an annual precipitation of 800 mm and mild winters and summers (average of 3°C in January and 21°C in July). These conditions explain the large historical density of wetlands. Currently 45% of the region is used for intensive agriculture, heavily fertilized and drained or irrigated. Another 26% of the land is urbanized (470 inhabitants/km²) and 13% of the soils are sealed (De Meyer et al. 2011). This has resulted in a substantial and steady increase of the number of recorded floods since the 1970s and an average yearly economic damage of 50 million euro (VMM 2014a). Remaining wetlands cover only 5% of the region and suffer from eutrophication, pollution, and disturbed hydrological regimes. All 25 wetland habitat types protected by the Habitats Directive are in an unfavorable conservation status (Louette et al. 2013). Most peat soils were extracted in medieval times and nearly all of the 6000 ha of remaining peat soils are heavily fragmented and assumed to be in a degraded, mineralized state.
In this paper wetlands are defined as temporary or permanently wet, nonmarine areas where typical wetland biodiversity is (still) more or less present. Consequently, “lost wetlands” must be understood as areas that, apart from ditches and small rivers or ponds, can no longer be considered as “wet” and lack typical wetland communities, including the temporary residing of migrating waterfowl. In our case wetland loss is not to be confused with degraded, damaged, or polluted wetlands, as is sometimes discussed in other literature (Davidson 2014).
We distinguished seven wetland categories (Table 1) based on drainage class (open water, permanently or temporary wet soil, tidal marsh) and trophic state (meso-, eu- or oligotrophic). Open waters (artificial water bodies, lakes, and large ponds) were included in the mapping and calculation of historical wetland loss, but were not considered in the restoration scenarios because their restoration/creation preconditions are less stringent.
Spatial analysis considered the following main maps:
For GIS analysis all maps were transformed into grid cells of 20 x 20 m. The area of water courses was considered constant over time and excluded from the analysis to avoid large errors in the calculation of area because of this grid transformation. All currently urbanized areas were considered to be not suitable for wetland restoration. For nonurbanized areas we assumed that in the long term the environmental conditions, as they were recorded in the 1950s, can be restored with appropriate measures. Further details about maps and spatial analysis are provided in supplementary material (Appendix 1 and 2).
For the calculation of the area of potential wetlands we distinguished two management scenarios: (1) an open (not forested) landscape scenario, and (2) a closed (forested) landscape scenario. To obtain realistic scenarios, the legal protection (“standstill principle”) for existing forests and open habitat types with nature value were taken into account.
We followed two approaches. First, the ecosystem service (ES) supply potential of the restoration scenarios was estimated for a broad bundle of services (see Table 2), based on the results of the Flanders Regional Ecosystem Assessment (Jacobs et al. 2014a, 2015, Stevens et al. 2015). Additionally, the socioeconomic relevance of this change in supply is demonstrated by estimating a monetary value for a selection of five services for which reliable monetary data are available from the ECOPLAN project (http://www.ecosysteemdiensten.be/cms/, https://www.uantwerpen.be/en/rg/ecoplan/).
The ES profile of the different wetland categories was obtained by a direct overlay of the wetland maps with ES supply maps from the Flanders regional ecosystem assessment (Stevens et al. 2015), and the consequent calculation of median supply per wetland category. Prior to the overlay, maps were normalized (0-100) to allow cross-service comparison and graphing along the same unit-axis. Scenario-changes in relative provision of the entire bundle of ESs were estimated based on the surface changes of 114 land-use classes for the whole of Flanders and their averaged and normalized ES-supply per ha and per year. This derived ES supply map based on 114 land-use classes allows direct translation of land-use scenarios to ecosystem service supply impact without redefining all biophysical and socioeconomic variables in the original quantification maps. To calculate the total impact on the level of the Flanders region, first, the total supply per ES of each habitat was multiplied with the surface area of this habitat, and normalized to obtain an ES supply profile for the whole of Flanders. Second, changes in these surfaces for each scenario provided alternative profiles. Finally, the relative difference between the scenario profile and the reference (current) profile provides the impact of the scenario in terms of increases and decreases in ES supply.
Monetary estimates were performed for wood production, climate regulation (as carbon storage in soils), food production, water quality regulation, and flood risk regulation (as water quantity regulation). The quantification and valuation methods have all been developed specifically for the Flemish Region (Broekx et al. 2013a,b, VITO 2014) and adapted to spatially explicit models at high resolution (ECOPLAN-project, details in Appendix 3). This monetization only aims at demonstrating the socioeconomic relevance of the multiple benefits from wetlands. Valuation for (societal) cost-benefit analysis, would have to include many other services, cost estimates, discount rates, assumptions on constant demand per service, additional quantification of nonmarket values, etc. (Dendoncker et al. 2014, Boeraeve et al. 2015). For all five services, the two scenarios were compared to the actual land use as baseline.
To translate the forest and open landscape scenarios into the data-driven models, the following assumptions were made:
An important indicator for the ambition level of any European government to restore part of the lost biodiversity are its (legally defined) conservation objectives for the implementation of the Habitats and Birds Directives. We translated the habitat types and habitats of protected species into our seven wetland categories and compared the Flemish objectives to increase wetland habitat area with the calculated “restoration opportunities” (defined as potential area for wetland restoration in the two scenarios, reduced with already existing wetland area).
In the 1950s 244,000 ha (19% of Flanders) could still be considered wetland (Table 3). Currently only 68,000 ha (5% of Flanders) remain, implying a substantial loss of almost 75% of wetland habitats over 50–60 years’ time. Thirty-seven thousand ha (15%) has been urbanized; the rest was mainly lost to intensification of agriculture and to a lesser extent also to an increase in forest production. The proportion of wetland loss differs between categories. Moist to wet heathlands and nutrient-poor grasslands decreased by 95% (-24,000 ha), with an identical rate of loss for the forested parts on these soils (-7,000 ha). Wet floodplain grasslands and polders decreased by 75% (-95,000 ha) and floodplain forests by 55% (-8,000 ha). Historical rich fens and marshes decreased by 95% (-41,400ha) and swamp forests by 60% (-4,500 ha). Permanently wet heathlands and open bogs showed a loss of 95% (-4,000 ha), while the forested version of this habitat decreased by 50% (-500 ha). Tidal marshes showed a reduction of 80% in area (-2,400 ha), mainly for land acquisition in the neighborhood of the port of Antwerp. In general, 20,000 ha (10%) of open wetland habitats disappeared because of active or spontaneous afforestation, with those on permanently wet soils proportionally being most affected. In contrast with the dramatic numbers above, deep waters tripled and shallow waters doubled in surface area over those years. At present, 100% of the deep waters and 90% of the shallow waters are eutrophic. Their trophic state could not reliably be reconstructed for the 1950s, but it is fair to assume that many meso- and oligotrophic waters have shifted to a eutrophic or even hypertrophic state.
According to our calculations (Table 3) there is still a potential to restore 147,000 ha of wetland in Flanders (deep and shallow waters excluded). In the long term this could bring the total amount of wetland to 215,000 ha or 17% of the territory. With appropriate measures to restore the conditions of the 1950s, floodplain grasslands and forests and wet polder areas can theoretically triple in surface area to a significant 120,000 ha. Oligotrophic wetland habitats on temporary wet soils could increase 14-fold to 26,500 ha. Restoration of wetlands on permanently wet soils would lead to a 6-fold increase of open and forested wetland habitats: 36,500 ha on meso-eutrophic soils and 4500 ha on oligotrophic soils, or 72% and 88%, respectively, of the original surface area of the 1950s. There is a huge potential for the restoration of tidal marsh along the river Schelde if embankments are moved inland. With many of these embankments already in place in the 1950s, this implies a 3-fold increase in area compared to the reference period and a 15-fold increase in area compared to the current situation. The potential for restoration of shallow waters was not calculated: in principle they can be artificially created in many sites. Maps with the modeled distribution of historical, current, and potential wetland categories in Flanders are provided in Appendix 4–7.
Wetlands in Flanders provide a broad bundle of services (Fig. 1). Ecosystems on permanently wet soils provide most ecosystem services, especially forested habitats. Provision of water quality regulation, pollination, and climate regulation are the most prominent. Cultural services and flood risk regulation are also important, as are air quality regulation, sound buffer, and wood production in forested permanently wet habitats. Systems on temporary wet soils have a very similar profile, but perform poorer on water quality regulation. The meso- and eutrophic habitats include seminatural grasslands that can be combined with food production (haymaking, grazing). Tidal marshes differ from the former wetlands by a high cultural value and hunting potential, but deliver a lower supply of water and air quality regulation and of sound buffer. Shallow waters provide a remarkably high supply of flood risk regulation but lower supplies of water and air quality and sound buffer.
The different restoration scenarios have a significant impact on total ecosystem services supply (Fig. 2). Both the forested and the open landscape scenario lead to a decrease in food production (-16% to -19%), an increase in both flood risk regulation and climate regulation (5% to 10%), and a strong increase in water quality regulation (31% to 46%). The forested scenario leads to an additional increase in sound buffer and wood production (9%), while these services slightly decrease in the open landscape scenario (respectively, -9% and -2.5%). Slight decreases occur also in the supply of coastal protection, air quality regulation, and production of energy crops.
Land-use choices involve a broad range of values, including economic, social, and ecological values. Economic values, especially market values, have the advantage of direct and tangible valuation. Using various pricing techniques, a surrogate economic value can be obtained for several services. Although the economic nonsense of such a pricing is understood, there is sense in that it conveys the order of magnitude of socioeconomic importance these services might well represent, thus demonstrating the need to capture their broader values in decision making. Absolute losses and gains in wood production, climate regulation, food production, water quality regulation, and flood risk regulation under the restoration scenarios are shown in Table 4. Agricultural production losses for the forested landscape scenario amount up to €185 million per year, which is 14 % of the total agricultural production compared with the current situation. If extensive agriculture is allowed on the temporary wet zones, this impact would be reduced with €50 million. Note that the real economic balance could look very different when taking into account changes in subsidies that would follow from a decrease in agricultural surface.
Benefits for water quality regulation (total nitrate release to surface water) are comparable for both scenarios, but depend on different aspects. While nitrate leaching is reduced most in the forested scenario, denitrification decreases significantly. In the open landscape scenario, nitrate leaching decreases less dramatically (11.3% instead of 16% for forested), but denitrification is relatively more performant because the decrease is only 4.2% (instead of 12% for forested). The monetary benefit ranges from €15 to €225 million per year. The high estimate (€74/kg N-NO3) is based on shadow prices of effectively implemented policy measures for nitrate in surface water (marginal cost method). Implementing a large scale restoration scenario that decreases nitrate release up to 20% would make a range of current technical measures dispensable. For correct valuation one should be able to derive a mean value for the dispensable measures. On the other hand, most water bodies do not meet the water quality standards despite the current measures.
Carbon sequestration in soils is relatively insensitive compared to the total stock in the Flemish Region, but highest under the forested scenario. The nature conservation management and agricultural management imply harvest of aboveground biomass, which results in less input to the soil compartment. This, however, does not mean that local changes cannot be important. Especially for the permanently wet ecosystems, active peat formation could be restored. Unfortunately, the quantification methods for soil organic carbon do not incorporate carbon stocks from potential peat formation.
Water quantity regulation is an ecosystem service that is likely to become more important in the next decades. Rewetting former wetland ecosystems allows increasing water retention by 7.6% under the forested scenario and by 4.8 % under the open landscape scenario. This volume of additionally retained water compares to a river with a steady flow of 2.8 and 1.8 m³/s, respectively. Whether the retained water could all be used for consumption can be disputed, but on the other hand this would be a service that is of strategic and crucial importance in terms of climate adaptation.
Comparing the restoration opportunity (potential area for wetland restoration in the two scenarios, reduced with already existing wetland area) and the objectives for wetland expansion in the Flemish Natura 2000 policy (Table 5), a significant discrepancy between the two figures appears. Present policy foresees a total wetland expansion of 8900-13,000 ha (or 7400-10,600 ha with open waters excluded) in 2050, including 1800-3000 ha forested wetland and 2500 ha tidal marsh. All figures are much lower than what could be reached with a more ambitious policy. The ambitions for oligotrophic and meso-eutrophic wetlands on temporary wet soils and meso-eutrophic wetlands on permanently wet soils appear to be especially modest with an increase of only 1–8% of the restoration opportunity. With a projected increase of 19–26% of the potential restorable surface, ambition levels are significantly higher for tidal marsh and wetlands on oligotrophic permanently wet soils.
Reliable estimations for (sub)national wetland loss, subdivided into different wetland subtypes and for an identical time period are very rare in literature (see Davidson 2014). The combination with an accurate and spatially explicit analysis of the wetlands that can potentially still be restored makes our study rather unique. Obviously, possible errors in area estimations and their mapping largely depend on the accuracy and scale of the map layers that are available for GIS analysis (Joao 1998). In the case of Flanders we were fortunate with the availability of detailed maps on a scale of 1:25,000 describing the soil conditions in the 1950s and the recent distribution of 180 habitat types, including 40 types of wetland habitat. Such accurate maps may not be available in other regions of the world and this poses a challenge to the exact replication of our methods (see also Clare and Creed 2014). Another error source may be the use of discrete values derived from the basic map layers to define the different classes of abiotic and biotic conditions. Within the limitations of the used basic data we believe our approach provides the best possible proxy for estimating wetland loss and the present potential for wetland restoration in Flanders. However, the maps that were generated (see Appendix 4-7 in Supplementary materials) should be interpreted with caution when zooming in on the individual site level. The transformation in to grid cells of 20 x 20 m, in combination with possible errors in the used basic map layers, may inevitably generate inaccuracies when maps are scaled down.
Concerning application of this approach in other areas, one should be aware of the impact of accuracy on the final area estimates, as was also observed by Davidson (2014). Especially in areas with low data availability, applying a min-max fork estimate could provide confident and transparent estimates.
Estimations of ecosystem service (ES) supply per wetland category as well as the impact estimation of the scenarios on total ES supply should be handled with caution. Here, we want to point out three caveats for ES supply estimates, which apply for any case study engaging in ES quantification. First, the data used for this exercise are the best available data on ES supply at this moment. These indicators are often combinations of several data layers and combined models involving a number of reasonable (and checked) assumptions. Indicators and maps should therefore be interpreted alongside their confidence and used within the boundaries of their specific purpose. Although the indicators used in this study robustly support our conclusions, they cannot be applied to answer just any question, especially not questions that require much higher accuracy and confidence, e.g., accounting, trend analysis, development of payment schemes, etc. (see also Jacobs et al., in press). Second, the maps are made and reviewed for the regional scale. Zooming in to local levels will bring to light biases caused by local physical, ecological, or social conditions that are not captured by the models. This has little repercussions for regional-scale analyses, but it is clear that the local ES supply of wetland types will differ strongly from one location to another. Third, and following from this local scale, there might be ecosystem services relevant on the Flanders scale that are not important at all at some locations (because either supply or demand is lacking). In fact, there might be important services missing from this analysis when scaling down to the implementation level, while the basic valuation at this level does not include a differential societal importance of the services. Conclusions drawn on this exercise are strictly general and may not be used to guide a local planning process.
Our monetary estimations indicate that benefits derived from the regulating services (water quantity regulation, water quantity regulation and carbon storage in soils) range from 20 to €268 million/yr. The decrease in production services (agriculture and timber production) ranges from 137 to €186 million/yr. Much can be debated about the quantification and valuation methods, including the validity of the scenario. Nevertheless these estimates demonstrate that for at least three services, substantial benefits could be obtained. A more sophisticated scenario would probably allow decreasing the impact on agricultural production and timber production, while maintaining these regulating services. Moreover, including health benefits, tourism, and recreation could tip the balance to positive numbers (e.g., Broekx et al. 2013b). Also the current mean cost of €50 million/yr to compensate for economic damage due to flood hazards (VMM 2014a) needs to be taken into account.
Despite these caveats, this analysis clearly shows the overall importance of specific wetland habitats for supply of mainly regulating and cultural services on the scale of the entire region. Restoring or creating wetland habitat in Flanders can result in a strongly increased supply of several important services, and in a decrease of food production. Basic economic valuation demonstrates the high societal importance of these services. Without being conclusive, this simple valuation opens a rational debate on whether the benefits and costs involved in food production might be outweighed by the broader benefits supplied by restored areas.
Our results broadly concur with earlier valuation studies (e.g., Thibodeau and Ostro 1981, Batie and Shabman 1982, Farber 1987, Costanza et al. 1989, Folke 1991, Gren et al. 1994, Russi et al. 2013) but especially highlight that valuation exercises should be broadened to include more than monetizeable benefits. First, not all ES are increasing, and societal trade-offs have to be made between benefits and losses of Flemish wetlands. Second, benefits and losses for different users should be disentangled to account for the governance issues involved in actual realization of a certain scenario. Third, a broader value typology to integrate intrinsic values, instrumental values, and relational values should be applied to go beyond an eye-opening study toward actual decision support (Jacobs et al. 2016a).
Projected losses in food production also consider the current production model, which involves substantial financial support from public budgets, as well as issues concerning food waste and caloric efficiency of meat production. Even a slightly different production model might easily compensate projected losses in wetland areas, or provide ways of farming that can be combined with the multiple services provided in these landscapes (Jacobs et al. 2014b, Van Gossum et al. 2014). Rather than retreating to the typical historical struggle for monofunctional land-service allocation and grinding on trade-offs between services and stakeholder groups, the many existing synergies on a practical and local level could offer concrete solutions for a multifunctional, biodiversity-rich wetland use. Such an approach could evoke a more transparent and rational debate on restoration of natural habitats in intensively used areas and might be more effective in obtaining biodiversity goals.
To understand the current state of ecosystems in any country or region and develop new policies, insights into past and current policies are essential. With almost 75% of its wetlands lost since the 1950s, Flanders ranks highest amongst the European regions (see data in European Commission 2007). The high population density and inappropriate spatial planning and urbanization policy were important drivers, as well as the lack of coordination of water management, which is traditionally very complex with many actors on different government and administrative levels. The European Water Framework Directive (2000), the Floods Directive (2007), and increasing socioeconomic costs of flood events in urbanized areas (VMM 2014a) were important turning points in the mind setting of the Flemish water policy makers. Nowadays some of the most prestigious nature restoration projects in Flanders go hand in hand with flood protection, e.g., for the large rivers Schelde, Grensmaas, and IJzer). The once common practice of widening and straightening of rivers and urbanization of flood-prone areas has virtually stopped. A more detailed overview on Flanders-specific water management practice can be found in Appendix 8.
Our integrated and spatially explicit approach delivers data that can be useful in the societal debate on more and better wetland restoration and is helpful to develop guidance for future decision making. In the case of Flanders we found that 35% of the remaining wetlands have no spatial planning or protection status, while 49,000 ha (33%) of potential wetlands lack investments for restoration despite their appropriate status (see Appendix 8). As was demonstrated in many other countries (e.g., Birol et al. 2009, Buijs 2009, Scholte et al. 2016) flood protection is more widely accepted as a motivation for wetland restoration than biodiversity conservation. In the Flemish floodplains this is demonstrated by the still dominant, more or less intensive agricultural use with fertilization and active drainage of wet grasslands. Outside the floodplains, restoration projects of nutrient-poor wet grasslands and heaths on temporary or permanently wet soils remain rare and small in scale. They are mainly restricted to nature reserves in the upstream and interfluvial areas. Conflicts with the surrounding land use in terms of water levels and water quality often hamper these projects. It is the public perception that such wetlands would not contribute to flood prevention and therefore they stay beyond the reach of the (traditionally much bigger) budgets of water management administrations. In general, the lack of interest in the restoration of wetland biodiversity is also reflected in the rather low ambition level for expansion of Natura 2000 wetland habitat types and habitats for Natura 2000 wetland species, particularly those of open landscapes.
For wetland restoration in general, and for the Flanders case specifically, we conclude that more awareness raising beyond direct biodiversity values will be essential to implement a more effective long-term restoration policy for the different types of biodiversity-rich wetlands. Fostering public support is not only essential, different stakeholder groups will need different kinds of information and opportunities for participation (see also Johansson and Henningsson 2011, Tolvanen et al. 2013, Aggestam 2014, Scholte et al. 2016). Studies like ours are essential to identify priority areas for restoration and create a more robust ecological network of wetlands (see also Gibbs 2000, Vos et al. 2010).
According to the most plausible scenarios described in the report of the Flanders Environment Agency (Brouwers et al. 2015) the mean temperature in Flanders may rise by up to 7.2% by 2100, which will lead to more extreme hot days and heat stress. Summers will get drier with more concentrated heavy rain events. Precipitation will be higher during winters. The sea level may rise up to 1 m. Combined with the predicted increase in population size and further urbanization of open space, the flood risk will further increase. Flood plain areas will hence become less valuable for agriculture and inhabitation, which is potentially facilitating their transition to (semi)natural wetlands.
Apart from reducing economic damage caused by floods (see e.g., Bullock and Acreman 2003, Acreman 2012, Acreman and Holden 2013, Walters and Babbar-Sebens 2016), both natural and artificial wetlands could produce additional adaptation services as water buffer areas to ensure sufficient water supply for the production of food crops in hot and dry periods (e.g., Chester and Robson 2013, Downard and Endter-Wada 2013). Artificial wetlands in cities and urban areas will become more important to reduce heat island effects and to buffer heavy rain (e.g., Persson et al. 1999, Sun et al. 2012). More wetlands will also help to remove increased nutrient runoff from cultivated catchments in regions with increased rainfall or more intensive agriculture (e.g., Gilliam 1994, Gren 1995, Woltemade 2000, Verhoeven et al. 2006, Thiere et al. 2009, Jeppesen et al. 2011, Hefting et al. 2013, Ockenden et al. 2014). To avoid depletion of ground water acquifers the creation of more temporary and permanent wetlands can increase the infiltration rate of rain water (e.g., Winter 1999). On the other hand it is possible that suitable areas for wetland restoration or even existing wetlands get lost or suffer from increased pressures in regions with increased droughts, with agricultural expansion as a possible secondary effect (e.g., Hartig et al. 1997). In general, climate change is expected to increase demand for wetland types of eutrophic, temporary wet and tidal soils (e.g. Nicholls 2004, Temmerman et al. 2013). The combination of increased evapotranspiration and more extreme weather events will probably challenge the restoration of specific wetland types that depend on more or less stable, high water levels, especially those of oligo-mesotrophic conditions (e.g., Cusell et al. 2013). Of course, restoration success will also depend on how the local geographical location is impacted by climate change (Čížková et al. 2013).
All in all, the urgency of climate change and the obvious role wetlands can play to increase resilience in multifunctional landscapes, provide arguments for their further protection and restoration. The application of accurate maps and ecosystem service assessments like this study are needed to underpin these arguments scientifically and help them to get implemented into spatial planning.
We show that despite dramatic historic wetland loss and the unfavorable status of remaining wetlands, Flanders still has a large biophysical and ecological potential for wetland restoration with the proper spatial planning or protection status already in place to justify more action in the field. Based on the best available data, we demonstrated that restoring or creating wetland habitat will result in a strongly increased supply of several important regulating and cultural ecosystem services, and in a slight decrease of food production. Benefits supplied by restored or created wetlands and avoided costs of economic damage due to flood hazards might outweigh the costs involved and the loss in food production. Different policies, specific designs and local implementation examples could offer opportunities for multifunctional use, even with producing services, of restored wetlands. An exhaustive area-wide approach, supported by innovative GIS modeling and ecosystem service valuation techniques, provide a robust tool for assisting evidence-based policy decision making that could be applied for other ecosystem types or areas.
We would like to sincerely thank Carine Wils for her help with the GIS mapping and analyses, and the reviewers and editors of Ecology and Society and Dries Bonte for their valuable comments. SJ thanks the EU-FP7 OpenNESS project and the IWT ECOPLAN project for financial support and both project teams for inspiring moments and discussions.
Acreman, M. C. 2012. Wetlands and water storage: current and future trends and issues. Ramsar Convention Secretariat, Gland, Switzerland.
Acreman, M., and J. Holden. 2013. How wetlands affect floods. Wetlands 33:773-786. http://dx.doi.org/10.1007/s13157-013-0473-2
Aggestam, F. 2014. Wetland restoration and the involvement of stakeholders: an analysis based on value-perspectives. Landscape Research 39:680-697. http://dx.doi.org/10.1080/01426397.2013.819076
ALBON. 2014. Bodemkaart: bodemtypes, substraten, fasen en varianten van het moedermateriaal en de profielontwikkeling. ALBON, Brussels, Belgium. [online] URL: http://www.geopunt.be/catalogus/datasetfolder/5c129f2d-4498-4bc3-8860-01cb2d513f8f
Batie, S. S., and L. A. Shabman. 1982. Estimating the economic value of wetlands: principles, methods, and limitations. Coastal Zone Management Journal 10:255-278. http://dx.doi.org/10.1080/08920758209361920
Birol, E., N. Hanley, P. Koundouri, and Y. Kountouris. 2009. Optimal management of wetlands: quantifying trade-offs between flood risks, recreation, and biodiversity conservation. Water Resources Research 45(11). http://dx.doi.org/10.1029/2008WR006955
Boeraeve, F., N. Dendoncker, S. Jacobs, E. Gómez-Baggethun, and M. Dufrêne. 2015. How (not) to perform ecosystem service valuations: pricing gorillas in the mist. Biodiversity and Conservation 24:187-197. http://dx.doi.org/10.1007/s10531-014-0829-9
Broekx, S., L. De Nocker, I. Liekens, L. Poelmans, J. Staes, K. Van der Biest, P. Meire, and K. Verheyen. 2013a. Estimate of the benefits delivered by the Flemish Natura 2000 network. 2013/RMA/R/87. University of Antwerp and University of Ghent, The Netherlands.
Broekx, S., I. Liekens, W. Peelaerts, L. De Nocker, D. Landuyt, J. Staes, P. Meire, M. Schaafsma, W. Van Reeth, O. Van den Kerckhove, and T. Cerulus. 2013b. A web application to support the quantification and valuation of ecosystem services. Environmental Impact Assessment Review 40:65-74. http://dx.doi.org/10.1016/j.eiar.2013.01.003
Brouwers, J., B. Peeters, M. Van Steertegem, N. van Lipzig, H. Wouters, J. Beullens, M. Demuzere, P. Willems, K. De Ridder, B. Maiheu, R. De Troch, P. Termonia, T. Vansteenkiste, M. Craninx, S. W. Maeten, W. Defloor, and K. Cauwenberghs. 2015. MIRA Climate report 2015, about observed and future climate changes in Flanders and Belgium. Flanders Environment Agency in collaboration with KU Leuven, VITO and RMI, Aalst, Belgium.
Buijs, A. E. 2009. Public support for river restoration. A mixed-method study into local residents’ support for and framing of river management and ecological restoration in the Dutch floodplains. Journal of Environmental Management 90:2680-2689. http://dx.doi.org/10.1016/j.jenvman.2009.02.006
Bullock, A., and M. Acreman. 2003. The role of wetlands in the hydrological cycle. Hydrology and Earth System Sciences 7:358-389 http://dx.doi.org/10.5194/hess-7-358-2003
Chester, E. T., and B. J. Robson. 2013. Anthropogenic refuges for freshwater biodiversity: their ecological characteristics and management. Biological Conservation 166:64-75. http://dx.doi.org/10.1016/j.biocon.2013.06.016
Čížková, H., J. Květ, F.A. Comín, R. Laiho, J. Pokorný, and D. Pithart. 2013. Actual state of European wetlands and their possible future in the context of global climate change. Aquatic Sciences 75:3-26. http://dx.doi.org/10.1007/s00027-011-0233-4
Clare, S., and I. F. Creed. 2014. Tracking wetland loss to improve evidence-based wetland policy learning and decision making. Wetlands Ecology and Management 22:235-245. http://dx.doi.org/10.1007/s11273-013-9326-2
Costanza, R., S. C. Farber, and J. Maxwell. 1989. Valuation and management of wetland ecosystems. Ecological Economics 1:335-361 http://dx.doi.org/10.1016/0921-8009(89)90014-1
Cusell, C., L. P. M. Lamers, G. van Wirdum, and A. Kooijman. 2013. Impacts of water level fluctuation on mesotrophic rich fens: acidification vs. eutrophication. Journal of Applied Ecology 50:998-1009. http://dx.doi.org/10.1111/1365-2664.12096
Davidson, N. C. 2014. How much wetland has the world lost? Long-term and recent trends in global wetland area. Marine and Freshwater Research 65:934-941. http://dx.doi.org/10.1071/MF14173
De Keersmaeker, L., N. Rogiers, R. Lauriks, and B. De Vos. 2001. Ecosysteemvisie Bos Vlaanderen, Ruimtelijke uitwerking van de natuurlijke bostypes op basis van bodemgroeperingseenheden en historische boskaarten. Eindverslag van project VLINA C97/06, studie uitgevoerd binnen het kader van het Vlaams Impulsprogramma Natuurontwikkeling. The Institute for Forestry and Game Management, Brussels, Belgium.
De Meyer, A., D. Tirry, H. Gulinck, and J. Van Orshoven. 2011. Actualisatie MIRA Achtergronddocument Bodem. Thema Bodemafdichting. Eindrapport. Studie uitgevoerd in opdracht van de Vlaamse Milieumaatschappij. MIRA, SADL & Departement Aard- en Omgevingswetenschappen, K.U. Leuven, Belgium.
Dendoncker, N., H. Keune, S. Jacobs, and E. Gomez-Baggethun. 2014. Inclusive ecosystem service valuation. Pages 3-12 in S. Jacobs, N. Dendoncker, and H. Keune, editors. Ecosystem services: global issues, local practices. Elsevier, New York, New York, USA. http://dx.doi.org/10.1016/b978-0-12-419964-4.00001-9
Downard, R., and J. Endter-Wada. 2013. Keeping wetlands wet in the western United States: adaptations to drought in agriculture-dominated human-natural systems. Journal of Environmental Management 131:394-406. http://dx.doi.org/10.1016/j.jenvman.2013.10.008
European Commission. 1995. Wise use and conservation of wetlands. COM (95) 189. European Commission, Brussels, Belgium.
European Commission. 2007. LIFE and Europe’s wetlands: restoring a vital ecosystem. Office for Official Publications of the European Communities, Luxembourg.
European Commission. 2015. Reporting under the EU Habitats and Birds Directives 2007-2012. The State of Nature in the EU. Office for Official Publications of the European Union, Luxembourg.
European Environmental Agency (EEA). 2010. EU 2010 biodiversity baseline. EEA, Copenhagen, Denmark.
European Topic Centre for Biological Diversity (ETC/BD). 2008. Habitats Directive Article 17 Report 2000-2006. European Topic Centre for Biological Diversity. [online] URL: http://bd.eionet.europa.eu/activities/Reporting/Article_17/Reports_2007/index_html
Farber, S. 1987. The value of coastal wetlands for protection of property against hurricane wind damage. Journal of Environmental Economics and Management 14:143-151. http://dx.doi.org/10.1016/0095-0696(87)90012-X
Folke, C. 1991. The societal values of wetland life-support. Pages 141-171 in C. Folke and T. Kåberger, editors. Linking the natural environment and the economy: essays from the Eco-Eco Group. Kluwer Academic, Dordrecht, The Netherlands. http://dx.doi.org/10.1007/978-94-017-6406-3_8
Gibbs, J. P. 2000. Wetland loss and biodiversity conservation. Conservation Biology 14:314-317. http://dx.doi.org/10.1046/j.1523-1739.2000.98608.x
Gilliam, J. W. 1994. Riparian wetlands and water quality. Journal of Environmental Quality 23:896-900. http://dx.doi.org/10.2134/jeq1994.00472425002300050007x
Gren, I.-M. 1995. The value of investing in wetlands for nitrogen abatement. European Review of Agricultural Economics 22:157-172. http://dx.doi.org/10.1093/erae/22.2.157
Gren, I.-M., C. Folke, K. Turner, and I. Batemen. 1994. Primary and secondary values of wetland ecosystems. Environmental and Resource Economics 4:55-74. http://dx.doi.org/10.1007/BF00691932
Hartig, E. K., O. Grozev, and C. Rosenzweig. 1997. Climate change, agriculture and wetlands in Eastern Europe: vulnerability, adaptation and policy. Climatic Change 36:107-121. http://dx.doi.org/10.1023/a:1005304816660
Hefting, M. M., R. N. van den Heuvel, and J. T. A. Verhoeven. 2013. Wetlands in agricultural landscapes for nitrogen attenuation and biodiversity enhancement: opportunities and limitations. Ecological Engineering 56:5-13. http://dx.doi.org/10.1016/j.ecoleng.2012.05.001
INBO (Research Institute for Nature and Forest). 2015. Biologische Waarderingskaart versie 2. INBO, Brussels, Belgium. [online] URL: http://www.geopunt.be/catalogus/applicationfolder/biologische-waarderingskaart
Jacobs, S., B. Burkhard, T. Van Daele, J. Staes, and A. Schneiders. 2015. ‘The Matrix Reloaded’: a review of expert knowledge use for mapping ecosystem services. Ecological Modelling 295:21-30. http://dx.doi.org/10.1016/j.ecolmodel.2014.08.024
Jacobs, S., N. Dendoncker, B. Martín-López, D. N. Barton, E. Gomez-Baggethun, F. Boeraeve, L. F. McGrath, K. Vierikko, D. Geneletti, J. K. Sevecke, et al. 2016a. A new valuation school: integrating diverse values of nature in resource and land use decisions. Ecosystem Services. http://dx.doi.org/10.1016/j.ecoser.2016.11.007
Jacobs, S., T. Spanhove, L. De Smet, T. Van Daele, W. Van Reeth, P. Van Gossum, M. Stevens, A. Schneiders, J. Panis, H. Demolder, H. Michels, M. Thoonen, I. Simoens, and J. Peymen. 2016b. The ecosystem service assessment challenge: reflections from Flanders-REA. Ecological Indicators 61:715-727. http://dx.doi.org/10.1016/j.ecolind.2015.10.023
Jacobs, S., T. Spanhove, and J. Panis. 2014a. Hoofdstuk 5 - Toestand en trend van ecosysteemdiensten in Vlaanderen (INBO.R.2014.6160407). In M. Stevens, H. Demolder, S. Jacobs, H. Michels, A. Schneiders, I. Simoens, T. Spanhove, P. Van Gossum, and W. Van Reeth, editors. Natuurrapport - Toestand en trend van ecosystemen en ecosysteemdiensten in Vlaanderen. Technisch rapport. . Vol. INBO.M.2014.1988582, Instituut voor Natuur- en Bosonderzoek, Brussels, Belgium.
Jacobs, S., T. Spanhove, M. Thoonen, L. De Smet, A. Boerema, K. Van der Biest, and D. Landuyt. 2014b. Hoofdstuk 9 - Interacties tussen aanbod, gebruik en vraag van ecosysteemdiensten in Vlaanderen (INBO.R. 2014.6160569). In M. Stevens, H. Demolder, S. Jacobs, H. Michels, A. Schneiders, I. Simoens, T. Spanhove, P. Van Gossum, and W. Van Reeth, editors. Natuurrapport - Toestand en trend van ecosystemen en ecosysteemdiensten in Vlaanderen. Technisch rapport. . Vol. INBO.M.2014.1988582, Instituut voor Natuur- en Bosonderzoek, Brussels, Belgium.
Jacobs, S., W. Verheyden, and N. Dendoncker. In press. Why to map? Guidelines for critical and effective ecosystem service mapping. In M. G. Burkhard, editor. Mapping ecosystem services. Pensoft, Sofia, Bulgaria.
Jeppesen, E., B. Kronvang, J. E. Olesen, J. Audet, M. Søndergaard, C. C. Hoffmann, H. E. Andersen, T. L. Lauridsen, L. Liboriussen, S. E. Larsen, M. Beklioglu, M. Meerhoff, A. Özen, and K. Özkan. 2011. Climate change effects on nitrogen loading from cultivated catchments in Europe: implications for nitrogen retention, ecological state of lakes and adaptation. Hydrobiologia 663:1-21. http://dx.doi.org/10.1007/s10750-010-0547-6
Joao, E. 1998. Causes and consequences of map generalization. Taylor and Francis, London, UK.
Johansson, M., and M. Henningsson. 2011. Social-psychological factors in public support for local biodiversity conservation. Society & Natural Resources 24:717-733. http://dx.doi.org/10.1080/08941920903530925
Louette, G., D. Adriaens, G. De Knijf, and D. Paelinckx. 2013. Staat van instandhouding (status en trends) habitattypen en soorten van de Habitatrichtlijn (rapportageperiode 2007-2012). Instituut voor Natuur- en Bosonderzoek, Brussels, Belgium.
Nicholls, R. J. 2004. Coastal flooding and wetland loss in the 21st century: changes under the SRES climate and socio-economic scenarios. Global Environmental Change 14:69-86. http://dx.doi.org/10.1016/j.gloenvcha.2003.10.007
Ockenden, M. C., C. Deasy, J. N. Quinton, B. Surridge, and C. Stoate. 2014. Keeping agricultural soil out of rivers: evidence of sediment and nutrient accumulation within field wetlands in the UK. Journal of Environmental Management 135:54-62. http://dx.doi.org/10.1016/j.jenvman.2014.01.015
Persson, J., N. L. G. Somes, and T. H. F. Wong. 1999. Hydraulics efficiency of constructed wetlands and ponds. Water Science and Technology 40:291-300. http://dx.doi.org/10.1016/s0273-1223(99)00448-5
Russi, D., P. Ten Brink, A. Farmer, T. Badura, D. Coates, J. Föster, R. Kumar, and N. Davidson. 2013. The economics of ecosystems and biodiversity for water and wetlands. IEEP, London, UK and Brussels, Belgium and Ramsar Secretariat, Gland, Switzerland.
Scholte, S. S. K., M. Todorova, A. J. A. van Teeffelen, and P. H. Verburg. 2016. Public support for wetland restoration: What is the link with ecosystem service values? Wetlands 36:467-481. http://dx.doi.org/10.1007/s13157-016-0755-6
Stevens, M., H. Demolder, S. Jacobs, H. Michels, A. Schneiders, I. Simoens, T. Spanhove, P. Vangossum, and W. Vanreeth. 2015. Flanders regional ecosystem assessment: state and trend of ecosystems and their services in Flanders. Synthesis. Research Institute for Nature and Forest, Brussels, Belgium.
Sun, R., A. Chen, L. Chen, and Y. Lü. 2012. Cooling effects of wetlands in an urban region: the case of Beijing. Ecological Indicators 20:57-64. http://dx.doi.org/10.1016/j.ecolind.2012.02.006
Temmerman, S., P. Meire, T. J. Bouma, P. M. J. Herman, T. Ysebaert, and H. J. De Vriend. 2013. Ecosystem-based coastal defence in the face of global change. Nature 504:79-83. http://dx.doi.org/10.1038/nature12859
Thibodeau, F. R. and B. D. Ostro. 1981. An economic analysis of wetland protection. Journal of Environmental Management 12:19-30
Thiere, G., S. Milenkovski, P.-E. Lindgren, G. Sahlén, O. Berglund, and S. E. B. Weisner. 2009. Wetland creation in agricultural landscapes: biodiversity benefits on local and regional scales. Biological Conservation 142:964-973. http://dx.doi.org/10.1016/j.biocon.2009.01.006
Tolvanen, A., A. Juutinen, and R. Svento. 2013. Preferences of local people for the use of peatlands: the case of the richest peatland region in Finland. Ecology and Society 18(2):19. http://dx.doi.org/10.5751/ES-05496-180219
United Nations World Water Assessment Programme (UNWWAP). 2003. Water for people, water for life. UNWWAP, Paris, France.
Van Gossum, P., S. Danckaert, T. Spanhove, and C. Wils. 2014. Hoofdstuk 11 - Ecosysteemdienst voedselproductie. (INBO.R.2014.1987588). In M. Stevens, H. Demolder, S. Jacobs, H. Michels, A. Schneiders, I. Simoens, T. Spanhove, P. Van Gossum, and W. Van Reeth, editors. Natuurrapport - Toestand en trend van ecosystemen en ecosysteemdiensten in Vlaanderen. Technisch rapport. Vol. INBO.M.2014.1988582, Instituut voor Natuur- en Bosonderzoek, Brussels, Belgium.
Verhoeven, J. T. A., B. Arheimer, C. Yin, and M. M. Hefting. 2006. Regional and global concerns over wetlands and water quality. Trends in Ecology & Evolution 21:96-103. http://dx.doi.org/10.1016/j.tree.2005.11.015
VITO. 2014. The Nature Value Explorer v2.2. VITO, Mol, Belgium. [online] URL: https://natuurwaardeverkenner.be/role.jsf
Vlaamse Milieumaatschappij (VMM). 2014a. Milieurapport Vlaanderen: waterkwantiteit. VMM, Aalst, Belgium. [online] URL: http://www.milieurapport.be/nl/feitencijfers/milieuthemas/waterkwantiteit/
Vlaamse Milieumaatschappij (VMM). 2014b. Overstromingsgevoelige gebieden 2014 (Watertoets). VMM, Aalst, Belgium. [online] URL: http://www.geopunt.be/download?container=overstromingsgevoelig&title=Overstromingsgevoelige%20gebieden%202014%20(Watertoets
Vos, C. C., D. C. J. van der Hoek, and M. Vonk. 2010. Spatial planning of a climate adaptation zone for wetland ecosystems. Landscape Ecology 25:1465-1477. http://dx.doi.org/10.1007/s10980-010-9535-5
Vriens, L., H. Bosch, G. De Knijf, S. De Saeger, R. Guelinckx, P. Oosterlynck, M. Van Hove, and D. Paelinckx. 2011. De Biologische Waarderingskaart: biotopen en hun verspreiding in Vlaanderen en het Brussels Hoofdstedelijke Gewest. Instituut voor Natuur- en Bosonderzoek, Brussels, Belgium.
Walters, K. M. and M. Babbar-Sebens. 2016. Using climate change scenarios to evaluate future effectiveness of potential wetlands in mitigating high flows in a Midwestern U.S. watershed. Ecological Engineering 89:80-102 http://dx.doi.org/10.1016/j.ecoleng.2016.01.014
Winter, T. C. 1999. Relation of streams, lakes, and wetlands to groundwater flow systems. Hydrogeology Journal 7:28-45. http://dx.doi.org/10.1007/s100400050178
Woltemade, C. J. 2000. Ability of restored wetlands to reduce nitrogen and phosphorus concentrations in agricultural drainage water. Journal of Soil and Water Conservation 55:303-309.
Wouters, J., K. Decleer, F. Vanderhaeghe, and M. Hens. 2013. PotNat, een GIS-tool voor het bepalen van de abiotische kansrijkdom van natuurtypen: Deel 1: Methodologie. Instituut voor Natuur- en Bosonderzoek, Brussels, Belgium.