Hydrobiologia (2006) 565:165–186 Ó Springer 2006 R.S.E.W. Leuven, A.M.J. Ragas, A.J.M. Smits & G. van der Velde (eds), Living Rivers: Trends and Challenges in Science and Management DOI 10.1007/s10750-005-1912-8

Biogeochemical constraints on the ecological rehabilitation of wetland vegetation in river floodplains

L.P.M. Lamers1,2,*, R. Loeb1, A.M. Antheunisse3, M. Miletto4, E.C.H.E.T. Lucassen1,2, A.W. Boxman1, A.J.P. Smolders1,2 & J.G.M. Roelofs1,2 1Department of Aquatic Ecology and Environmental Biology, Institute for Wetland and Water Research, Radboud University Nijmegen, Toernooiveld 1, 6525 ED Nijmegen, The Netherlands 2B-Ware Research Centre, Toernooiveld 1, 6525 ED Nijmegen, The Netherlands 3Landscape Ecology, Institute of Environmental Biology, Utrecht University, P.O. Box 80084, 3508 TB Utrecht, The Netherlands 4Department of Microbial Wetland Ecology, Netherlands Institute of Ecology (NIOO-KNAW), Centre for Limnology, Rijksstraatweg 6, 3631 AC Nieuwersluis, The Netherlands (*Author for correspondence: E-mail: [email protected])

Key words: biodiversity, biogeochemistry, eutrophication, flooding, nature conservation, review, water management

Abstract The European policy for river management during peak discharge periods is currently changing from exclusion strategies (reinforcement of dykes) to allowing a more natural situation by creating more floodplain space to reduce water levels during peak discharges. In addition, water retention and water storage areas have been created. The new measures are generally being combined with nature development strategies. Up till now, however, ecological targets of broadscale floodplain wetland restoration including sedge marshes, species-rich floodplain forests and carrs, riparian mesotrophic grasslands and other bio- diverse riverine ecosystems, have hardly developed in these areas. Most studies on the conditions needed for sustainable ecological development of floodplains have focused on hydrological and geomorphological rather than biogeochemical issues (including nutrient availability and limitation). There are, however, large differences in the composition of river water and groundwater and in sediment quality between rivers in densely populated areas and those in more pristine areas, which serve as a reference. It is very likely that these factors, in combination with heavily altered hydrological regimes and the narrow areas confined between the dykes on both sides of the rivers, impose major constraints on sustainable ecological devel- opment of riverine areas. Another issue is that existing wetlands are generally considered to be very appropriate for water retention and conservation, although recent research has shown that this may pose a serious threat to their biodiversity. The present paper reviews the biogeochemical constraints on the combination of floodplain rehabilitation, water conservation and the conservation and development of wetlands. It is concluded that biogeochemical problems (mainly related to eutrophication) predominantly arise in less dynamic parts of the river system, to which the flood-pulse concept applies less. Sound knowledge of the biogeochemical processes involved will contribute to greater efficiency and a better prediction of the opportunities for restoration and development of riverine wetlands. This information can be directly applied in nature management, water management, policy-making and consultancy.

Changing concepts, changing rivers Netherlands have to cope with excessive volumes of water in a narrow area confined by dykes. In the During peak discharge periods, regulated river past, the only strategy was to avoid flooding by systems in densely populated areas such as the raising and reinforcing dykes (levees). Since this 166 approach has proved to be insufficient and types that have become rare along rivers in concepts of river management have changed, new densely-populated areas, such as sedge fens and strategies have been designed which allow the alder carrs. After the discussion of constraints re- development of a more natural riverine system lated to surface water, groundwater and soil (Admiraal et al., 1993; De Waal et al., 1995; quality, their possible effects on plant biodiversity Harper et al., 1995; Sparks, 1995; Tockner et al., will be treated. Next, implications for river man- 1998). Additional space is created in floodplains to agement and nature conservation will be outlined. avoid the heading up of water between the embankments, thereby reducing flood risks (‘Space for the river’ or ‘Room for the river’ policy; Van River water: quantity and quality Stokkom et al., 2005). In this approach, flood- plains are mechanically lowered (by excavation) The hydrological changes described above will al- and widened (by dyke displacement), secondary low rivers greater access to their floodplains (in- channels are excavated and large obstacles re- creased connectivity), and increase water level moved, so that rivers are able to expand into side fluctuations in floodplains. In other words, the channels and wetland areas (Mostert, 1998; flood pulse, regulating floodplain functioning, is Nienhuis et al., 1998; Leuven et al., 2000; Smits restored (Junk et al., 1989; Bayley, 1991; Heiler et al., 2000). In strongly regulated rivers lacking et al., 1995; Middleton, 1999; Van Geest et al., space for natural hydrogeomorphological pro- 2005). In the river deltas, more estuarine processes cesses, the concept of ‘cyclic rejuvenation’, mim- (tidal action and brackish water intrusion) will be icking the effects of natural channel migration, is allowed (Smits et al., 2000, 2006). As a result of being applied as a compromise between flood this, large areas will be temporarily or permanently protection and ecological rehabilitation (Simons flooded, or waterlogged. Such water table fluctu- et al., 2001; Baptist et al., 2004; Geerling et al., ations lead to alterations between oxidized and 2006). In addition, inland areas are adapted to reduced soil conditions. In addition, flooding will retain and store excess water temporarily, both in become more frequent in many parts of the natural wetlands and in agricultural areas. floodplains. It is evident that these changes in the These new safety strategies are generally being water table regimes will lead to changes in vege- combined with habitat creation strategies (Cals tation composition, as a result of the susceptibility et al., 1998; Buijse et al., 2002; Nienhuis et al., of plants to waterlogging and flooding. The 2002). The ecological objective is to create flood- duration of the flooding periods, as well as the plain forests, riparian grasslands, sedge fens and seasonality and frequency of flooding, have major other biodiverse ecosystem types (Postma et al., effects on species distribution and succession. This 1996; Bal et al., 2001; Buijse et al., 2005; Nienhuis, may be the result of low oxygen concentrations 2006). Up till now, however, high-biodiversity affecting root growth (Van der Valk, 1981; Jack- vegetation targets have hardly been achieved, son & Drew, 1984; Brock et al., 1987; Crawford, especially for the permanently flooded or water- 1987, 2003; Smits et al., 1990b; Van den Brink logged locations (Pedroli et al., 2001; Buijse et al., et al., 1991; 1995; Armstrong et al., 1994; Blom 2002, 2005). Increased connectivity to the river has et al., 1994; Lenssen et al., 1999; Visser et al., even been shown to decrease species diversity for 2003; Van Eck et al., 2006) or germination (Smits macrophytes, macrozoobenthos, phytoplankton et al., 1990a, 1995), decreased photosynthetic rates and zooplankton in floodplain lakes along the (Vervuren et al., 1999; 2003; Mommer et al., lower Rhine and Meuse rivers (Van den Brink 2004), erosion and sedimentation, or mechanical et al., 1993; Van den Brink & Van der Velde, stress by current and wave exposure (Coops & Van 1994). der Velde, 1996; Coops et al., 1996a). In addition The present paper will focus on biogeochemical the water depth and the fluctuation of the water problems that may impede ecological rehabilita- table strongly control vegetation growth and tion of floodplain vegetation, and on their relation zonation (Coops & Van der Velde, 1995; Coops with the flood-pulse concept. The rehabilitation et al., 1996a; Van Geest et al., 2005). Lastly, discussed includes the development of ecosystem flooding may well increase dispersal of propagules 167 via the surface water (Brock et al., 1987; Coops & floods, however, large amounts of phosphate may Van der Velde, 1995; Amoros, 2001; Boedeltje be released from heavily fertilized floodplain soils et al., 2004). as they dissolve and diffuse to the water layer, as Increased connectivity between rivers and their explained below. This is expected to reduce the floodplains will inevitably lead to a stronger retention capacity for this nutrient, or may even influence of the river water quality on floodplain lead to a net efflux of P from the floodplains to the communities, such as demonstrated for the rela- river. In addition, flooding is known to stimulate tion between the flooding frequency and duration, the loss of C, N and P as a result of increased and their water quality (Van den Brink & Van der decomposition and mineralization in soil litter, for Velde, 1994). The magnitude and location of instance, in floodplain forests (Lockaby et al., sedimentation and erosion directly influence bio- 1996). diversity while, conversely, the vegetation type has Even if river water is eutrophic, mesotrophic a significant effect on hydrogeomorphological vegetation types such as sedge fens may develop on characteristics such as hydraulic conductivity and floodplains. The occurrence of these often biodi- sedimentation (Hupp, 1992; Bornette & Amoros, verse vegetation types depends on a low total influx 1996; Coops et al., 1996b, 1999; Henry et al., 1996; of nutrients. This can be determined by a lower Steiger & Gurnell, 2003). High connectivity seems level of connectivity (by distance), but may also to be a prerequisite for some aquatic species such result from permanent influx of nutrient-poor as as Nymphoides peltata through the prevention groundwater like in floodplains surrounded by of calcium deficiency (Smits et al., 1992). Surface calcareous massifs (Kohler et al., 1973; Carbiener water pollution by phosphate and nitrate increases et al., 1990; Bornette et al., 1996), or from seepage the influx of these nutrients, both in the dissolved of iron-rich groundwater (see below). It can, how- and the particulate fraction, in floodplain soils ever, also be caused by the presence of highly (Van den Brink et al., 1993; Van den Brink & Van productive, eutrophic or hypertrophic vegetation der Velde, 1994; Knowlton & Jones, 1997; types between the biodiverse vegetation types and Kronvang et al., 1998; Spink et al., 1998; Darke & the river. Nutrient-rich, species-poor stands of Walbridge, 2000). This direct pollution is known helophytes such as Phragmites australis, Typha to be able to decrease biodiversity in floodplain latifolia, Phalaris arundinacea and Glyceria streams and ponds through eutrophication (Van maxima, and of species-poor brushwood and den Brink et al., 1993; 1994; Tre` molieres et al., forests adjacent to rivers, act as natural helophyte 1994; Bornette et al., 2001; Bij de Vaate et al., filters, removing nutrients and particles from the 2006). Floodplain waters become dominated by surface water and enabling biodiverse wetland algae and cyanobacteria, causing the deterioration types to develop farther from the river (Wassen of underwater communities. et al., 2002; 2003). However, vast floodplains, On the other hand, nutrient retention and ranging in width from several hundreds of metres removal by denitrification, uptake by plants, to several kilometres (e.g., up to 10 km for the river periphyton and microorganisms, sedimentation Danube), are absent in densely populated lowland and binding to the soil are regarded as important areas such as the Netherlands (Buijse et al., 2002). regulators of nutrient concentrations in the river The actual space for the river will, even after the water (Kok et al., 1990; Sjodin et al., 1997; Reddy implementation of the new plans for riverine areas, et al., 1999; Olde Venterink et al., 2003b; Hogan in most cases be too narrow to allow for et al., 2004). Van der Lee et al. (2004) studied the these lengthy gradients from nutrient-rich to eutrophic river Rhine and found that nitrogen (N) nutrient-poor conditions, including both mineral retention was insignificant (<3% of the annual and more organic soils. All desired wetland types of load), while phosphorus (P) retention amounted to the target situation including marshes, natural 18%. For both nutrients, sedimentation was the grasslands and floodplain forests (Postma et al., most important retention mechanism. As a result 1996; Buijse et al., 2005), have to develop within a of differences in connectivity, flow velocity and ‘straitjacket’, confined by the embankments on vegetation composition, nutrient accretion rates both sides. The better the water quality, the smaller show great variability within floodplains. During the influence of this spatial problem, which is very 168 difficult, if not impossible, to tackle in densely alkalinity (Kok et al., 1990; Roelofs, 1991; Smol- populated areas. This fact underlines the need for ders & Roelofs, 1995) as a result of the reduction cleaner water and cleaner floodplain soils. of electron acceptors such as sulphate or nitrate. In addition to the direct eutrophication effects The effect of the increased sulphate concentrations caused by the influx of nutrients, high sulphate and is, however, strongly dependent on the soil type, chloride concentrations in the inflowing surface and on the concentrations (Lamers et al., 2001). water may cause adverse biogeochemical One of the significant factors is the concentration changes in many Dutch wetlands with organic of free iron that is available for phosphate binding. soils, including former river meanders (Roelofs, For wetland soils, the ratio between the pore water 1991; Koerselman et al., 1993; Smolders & concentrations of iron and phosphate has a high Roelofs, 1993; Lamers et al., 1998b; Beltman indicator value. Phosphate mobilization to the et al., 2000; Lamers et al., 2001; Verhoeven et al., water layer is low at values above 10 (mol mol)1), 2001; Lucassen et al., 2004a). Concentrations of and moderate between 1 and 10 (Smolders et al., both anions in wetlands have greatly increased as a 2001; Geurts et al., unpublished data). At ratios result of anthropogenic input into rivers (indus- below 1, however, iron levels proved to be too low trial, agricultural and mining activities) and the to provide a functional ‘iron trap’. Although it is increased use of river water to compensate for known that the microbial community responsible water shortage in agricultural areas and nature for sulphate reduction strongly changes over a reserves (Roelofs, 1991; Smolders & Roelofs, 1993; salinity gradient (Laanbroek & Pfennig, 1981), Lamers et al., 1998b). Increased chloride loading information on the composition of assemblages of from chloride-polluted river Rhine water may sulphate-reducing prokaryotes in S-polluted negatively influence the vegetation composition of freshwater systems is still limited (Lomans et al., floodplain lakes (Van den Brink & Van der Velde, 1997; Castro et al., 2002). Potential sulphate 1993). Large amounts of sulphate are being reduction rates appear to be high within the first mobilized by the oxidation of geological iron sul- week after the water table is raised in drained soils, phide (FeSx) deposits by oxidation as a result of showing the presence and fast response of a viable drainage (Lamers et al., 1998a; Schuurkes et al., community of sulphate-reducing prokaryotes 1988) and, indirectly, by nitrate pollution of (Miletto, unpublished data). aquifers from the intensive fertilization of arable land (Nordstrom, 1982; Ko¨lle et al., 1985; Paul & Clark, 1989; Appelo & Postma, 1993; Hoffmann Groundwater changes et al., 1998; Lamers et al., 2002b). In the latter case, FeSx is oxidized anaerobically through Floodplains show a hydrological mosaic, caused chemolithotrophic denitrification, resulting in even by differences in lateral fluxes between uplands, stronger increases in sulphate concentrations in floodplains and the river, often including discharge groundwater and, after discharge, surface water. locations. In more pristine areas, discharged Increased sulphate concentrations, between 1 and groundwater will generally be nutrient-poor (De 4 mmol l)1 for Dutch surface waters, may well Mars et al., 1997). This discharge of hypoxic, iron- lead to considerable phosphate mobilization under or calcium-rich groundwater may lead to increased anaerobic conditions (waterlogging and flooding; phosphate retention, decreasing phosphate Fig. 1). The increased production of sulphide in availability under natural wetland conditions. The sulphur (S)-polluted wetlands by increased micro- extent to which this occurs depends on the bial sulphate reduction rates interacts with iron hydrological characteristics and on soil and phosphate binding, leading to the formation of groundwater quality (Stumm & Morgan, 1981). As iron sulphides (including pyrite), the release of a result of large-scale water table draw-down in the phosphate and potential iron deficiency (Bostro¨m catchment, groundwater fluxes have greatly chan- et al., 1982; Caraco et al., 1989, 1993; Roelofs, ged in floodplains. This has led not only to 1991; Smolders & Roelofs, 1993; Smolders et al., quantitative changes, but also to qualitative 1994; Lamers et al., 2002a). In addition, mineral- changes. Discharge of iron has decreased, and ization rates may be stimulated by increased groundwater may contain high concentrations of 169

30,0 sedge meadow +S ) 1

− 25,0 fen meadow +S

alluvial meadow +S

(µmol l 20,0 − 2 4 sedge meadow control 15,0 fen meadow control

10,0 alluvial meadow control

5,0 soil moisture HPO

0,0 −4 0 4 8 12 16 20 24 28 weeks of waterlogging Figure 1. Phosphate mobilization (pore water) in mesocosm experiments by waterlogging with (S) or without (control) sulphate pollution (2 mmol l)1) for three different sediment types. Sedge meadow: grassland abandoned by agriculture approx. 10 years before the onset of the experiment; organic sediment; Caricion nigrae vegetation. Fen meadow: organic sediment; Junco-Molinion vegetation. Alluvial meadow: clay sediment; Alopecurion pratensis vegetation. After 11 weeks (end of the experiment), phosphate mobilization in the alluvial meadow was just as high as for sulphate-treated sedge-meadows, although its initial rate was slower. In contrast to the other two types, however, sulphate did not fortify the eutrophication process. Means are given, with their standard errors (n=6, 5, 5, respectively). (From: Lamers et al., 2001; Loeb & Lamers, 2003). nitrate and sulphate (up to the millimolar range; tural areas, nitrate intrusion has led to sulphate Lucassen et al., 2004a; Baker & Vervier, 2004). As ‘charging’ of the groundwater in this way (Pauwels a result of the high P sorption capacity of soils, et al., 1998; Lamers et al., 2002b). In addition to phosphate concentrations in groundwater are this source, increased sulphate concentrations in considerably lower. Depending on the constitution groundwater originate from sulphide oxidation by of the subsoil (including water table and organic oxygen in drained wetland soils and sulphate efflux matter concentration), a large part of the nitrate from farmlands (Heathwaite, 1990; Freeman may be reduced and lost from the soil by denitri- et al., 1996; Lamers et al., 1998a,b). Values of up fication (Kester et al., 1997; Baker & Vervier, to 8 mmol l)1 have been reported for groundwa- 2004; Hefting et al., 2004), although this strongly ter-feeding carr woods in oxbow lakes (Boxman depends on the actual hydrology of the floodplain et al., 2003). In many Dutch riverine wetlands that (Burt et al., 1999) and soil texture (Pinay et al., still show seepage, groundwater has become the 2000). Higher water tables and concomitant primary source of sulphate pollution (Lucassen absence of oxygen may, however, promote nitrate et al., 2004a). reduction to ammonium, leading to ammonium Sulphate pollution is especially detrimental accumulation in the topsoil. On drained sites, when the efflux of water and dissolved nutrients nitrification rates are high and nitrate may accu- is blocked. The creation of high water tables mulate (Hefting et al., 2004). throughout the year as a measure against In addition to organotrophic denitrification, desiccation was found to be disastrous for alder nitrate may also be reduced by lithotrophic carrs in former river meanders (oxbow lakes) of denitrifiers, as explained above. Although the the river Meuse in the Netherlands. This (for this passage of nitrate-rich groundwater through wetland type) artificial situation differs strongly aquifers rich in iron seems beneficial because it from the natural conditions, which involve water means that nitrate is stripped from the water, a table fluctuations during the year with lower val- significant new problem is created because ues in the summer. Even within the first growing sulphide is oxidized to sulphate. In many agricul- season after the introduction of this measure, the 170 surface water had become completely covered by redox reactions). Additionally, chloride and Lemna spp., while characteristic species such as sulphate concentrations will increase. Increased Caltha palustris, Calla palustris and several Carex sulphate and chloride concentrations in freshwater spp. had been literally drowned. Phosphate con- wetlands are expected to reduce biodiversity by centrations had increased severalfold as a result of eutrophication (nutrient-controlled systems), as prolonged anoxic conditions and accumulation described above. Higher values of both anions (Boxman et al., 2003; Lucassen et al., 2004a, under brackish conditions, however, will probably 2005). Alnus glutinosa even started to die at these lead to the rehabilitation of characteristic (eutro- locations. It may be clear that this artificial phic) brackish wetlands with a high biodiversity. hydrological regime (‘over-rewetting’) is very det- Unlike freshwater wetlands, which are generally rimental and undesirable. The same holds for nutrient-controlled, these are expected to be con- other types of wetlands, such as fen meadows. trolled by high, fluctuating chloride concentrations Aerobic conditions are needed in summer, (salinity controlled), or by the interaction between presumably also to stimulate nitrification. Plants salinity and nutrient availability (Den Hartog, characteristic of this type of habitat show a strong 1974; Bloemendaal & Roelofs, 1988; Moss, 1994). preference for nitrate as an N source, while the Higher salinity will be directly toxic to most accumulation of ammonium leads to toxic condi- freshwater species, but it will also influence them tions (Smolders et al., 1996; De Graaf et al., 1998; indirectly, through the modification of biogeo- Lucassen et al., 2003; Van den Berg et al., 2005). chemical processes like decomposition, minerali- The re-establishment of a more natural water zation, redox reactions and nutrient release regime, with higher water tables during winter and (Christian et al., 1990; Van den Brink & Van der tables below the hydraulic head of the seepage Velde, 1993; Hall & Anderson, 1995; Mendelssohn groundwater (allowing adequate nutrient efflux), et al., 1999; Hyacinthe & Van Cappellen, 2004). have been shown to offer better prospects (Lucassen et al., 2004a, 2005). A positive effect of water table fluctuation on biodiversity was also Soil quality: nutrients found for floodplain lakes along the Lower Rhine (Van Geest et al., 2005). For fen types that need a Research in floodplains has generally focused on constant water level at soil surface, however, this hydrology and river water quality. Decomposition type of water table fluctuation does not provide a rates are, for instance, strongly determined by soil solution. moisture content. In peatlands, water tables are a Interestingly, high nitrate concentrations in the major factor controlling decomposition rates, with groundwater, originating from intensive fertiliza- higher rates at drained sites (Bridgham & Rich- tion of arable land in the catchment, are able to ardson, 2003). However, in addition to hydrolog- reduce or even counteract the adverse effects of ical changes and surface water quality, several stagnating, sulphate-rich conditions. Since nitrate other factors have been drastically changed in is a more favourable electron acceptor than riverine areas. Although river water quality in the sulphate, it appears to be able to prevent sulphate- Netherlands has improved in recent decades, and induced phosphate mobilization and sulphide nutrient concentrations have decreased substan- accumulation (Lucassen et al., 2004b). On the tially (except for N in the Meuse; Admiraal et al., other hand, accumulation of ammonium (Hefting 1993; Bij de Vaate et al., 2006; Fig. 2), most et al., 2004) may well lead to undesirable floodplains are still highly eutrophic, because they dominance of highly competitive graminoid spe- have received massive amounts of nutrients by cies (Lucassen et al., unpublished data). fertilization (manure and fertilizer) for agricultural In addition to the plans for development of purposes. In the Netherlands, influxes may freshwater wetlands, plans have also been amount to 500 kg N and 75 kg P per hectare per announced to restore more saline conditions to year (CBS, 2004). Additionally, approximately 20– regions that had formerly been brackish. This is 30 kg of airborne N per hectare per year is expected to lead to restored tidal water level fluc- deposited in non-forest vegetation (Boxman, tuations, affecting soil biogeochemistry (including 2005). Due to high concentrations of amorphous 171

− discharge (m3 s 1) 25 200 discharge P 3000 N Rhine 0 175 ) ) 1 1 − 20 − discharge 2500 150 0 Meuse 125 15 2000 0 Rhine 100 1500 0 10 75 Meuse

10000 50 5 total concentration (µmol l total concentration (µmol

500 l total concentration (µmol 25

0 0 0 0 0 1975 1980 1985 1990 1995 2000 2005 1975 1980 1985 1990 1995 2000 2005 year year Figure 2. Total concentrations of phosphorus (P) and nitrogen (N) in the rivers Rhine and Meuse, and their discharges, between 1980 and 2003. (Data from: RIVM, 2004).

(i.e. non-crystalline) iron and aluminium (the latter 1974; Khalid et al., 1977; Lamers et al., 1998a, b; especially in clay), the phosphate-binding capacity Baldwin & Mitchell, 2000; Loeb et al., unpub- of wetlands is generally very high. Clay soils in lished data). Phosphate bound to aluminium and particular are known to have an extremely high aluminium-organic compounds will, in contrast, potential for P retention. Phosphate retention ap- not be released, because the metal is not redox- pears to be strongly correlated to the concentra- sensitive (Darke & Walbridge, 2000). This means tion of both compounds, especially aluminium that after retention, the greater part of this frac- (Richardson, 1985; Darke & Walbridge, 2000). tion will stay in the soil. The same holds for Furthermore, pollution of the groundwater with phosphate bound to calcium carbonate nitrate may also lead to the accumulation of ni- (Golterman, 1998; Reddy et al., 1999). For cal- trate and ammonium in floodplain soils (as ex- cium-rich soils, the relatively high proportion of P plained above). This soil eutrophication (P and N) may be sequestered by calcium, preventing phos- appears to impose a more imminent constraint on phate from dissolving during flooding. the ecological rehabilitation of floodplains than Recent research has revealed that many flood- the threat from the river water quality, especially plains along Dutch rivers show high concentra- now that this quality is improving. tions of free iron in the soil moisture (up to It is known that flooding of former agricultural 6000 lmol l)1), preventing the accumulation of land leads to excessive P eutrophication by oxygen free sulphide (Loeb et al., unpublished data). depletion and concomitant phosphate mobiliza- Total soil iron concentrations range from 0.6 to tion (Lamers et al., 1998b; Lamers et al., 2001; 6% (100–1000 lmol g)1 DW; Loeb, unpublished Young & Ross, 2001; Van Dijk et al., 2004). This data), as a result of sedimentation from river is a result of limited access of oxygen to soils and surface water (the iron concentration in suspended the concomitant shift to other electron acceptors matter being 3% for the Rhine; ICPR, 2005). available for soil microorganisms, decreasing the Higher concentrations of iron measured in flood- overall redox potential and generating alkalinity plain soils are mainly the result of the discharge of (Ponnamperuma, 1984; Laanbroek, 1990). Simul- anaerobic, iron-rich groundwater in the present or taneously, the concentration of dissolved free iron past. Upon reaching the (temporarily or perma- increases (due to iron reduction) and phosphate is nently) aerobic top layer, iron is oxidized to iron mobilized (Patrick et al., 1973; Patrick & Khalid, hydroxides, and accumulates. This may even lead 172 to the deposition of iron stone (marsh ore, with Fe mesocosm study using mineral monoliths with concentrations up to 50%), for instance in river their vegetation. While summer floods resulted in valleys whose soils comprise a mixture of sand and phosphate mobilization, especially with sulphate- peat (called dalgronden in Dutch). rich water, winter floods did not produce signifi- At first glance, these high iron concentrations cant changes (Fig. 3; Bruijnzeels & Lamers, seem to be beneficial because of their high phos- unpublished data). It was striking that this soil phate-binding capacity. Unfortunately, this high showed a clear response to sulphate pollution, in binding capacity has, in combination with the contrast to other mineral soils tested, as explained heavy fertilization of floodplains, led to excessive above. After the water table had been lowered, loading of the soil’s phosphate binding sites. As a phosphate concentrations dropped again, as result, the phosphate concentration in pore water expected. In addition, sulphide hardly accumu- may increase to 20–200 times the original con- lated during the winter, while sulphide production centration during flooding, leading to concentra- peaked in spring, both in the sulphate and control tions up to 225 lmol l)1 (Young & Ross, 2001; conditions of the experiment. During water Loeb et al., unpublished data). The phosphate table draw-down, sulphate reduction ceased and saturation of iron-based binding sites has proved sulphide became oxidized. Although the season- to be a powerful diagnostic tool to predict phos- ality of flooding is clearly one of the most impor- phate mobilization during waterlogging or flood- tant variables controlling both floodplain ing (Young & Ross, 2001; Hogan et al., 2004; biogeochemistry and vegetation responses, it is Loeb et al., unpublished data). In combination remarkable that river management options with ammonium loading, this phosphate ‘time involving combined water storage and ecological bomb’ provides a serious biogeochemical pitfall rehabilitation often ignore this fact. Most plants for the creation of biodiverse wetlands, whose will survive winter flooding, and nutrient cycling production is generally limited by the availability rates are modest. While water storage as a flood of N, phosphate, potassium or a combination of prevention measure during winter is relatively these nutrients (Aerts & Chapin, 2000; Olde harmless to most vegetation types of mesotrophic Venterink et al., 2003a; Gu¨sewell, 2004). In iron- wetlands (see also Van Eck et al., 2006), storage rich, mineral floodplain soils, sulphate pollution during the growing season will be disastrous. does not lead to extra phosphate mobilization. The high availability of iron appears to be sufficient to sequester all sulphide produced without phosphate Soil quality: toxicants being mobilized from iron-phosphate complexes (Fig. 1; Loeb & Lamers, 2003). We expect that Next to nutrients and chloride, floodplains in floodplain soils higher in organic matter and lower densely populated areas have also been loaded in iron concentration will react like fen soils when with a large range of contaminants, including polluted with sulphate, and show both extra potentially toxic metals and persistent organic phosphate mobilization and sulphide accumula- micropollutants (Leuven et al., 2005). There is tion (Lamers et al., 1998b; 2001). This is currently considerable variation in concentrations between being tested. Peat-forming systems, similar to fens different floodplain areas along the same river, as outside river floodplains, have become very rare in well as along horizontal and vertical gradients regulated riverine landscapes. This may be attrib- within the same floodplain (Middelkoop, 2000). uted to the lack of space that would allow for less Concentrations of many compounds exceed dynamic aquatic wetland types farther from the quality standards at many locations (Japenga river, showing lower sedimentation and erosion et al., 1990; Hendriks et al., 1995; Gocht et al., rates, and to the young age of most ‘rejuvenated’ 2001). Even if pollutant concentrations in river floodplains, in which the accumulation of organic water decrease, there is still accumulation in the matter has only just started. soil (Japenga & Salomons, 1993; Leuven et al., The extent to which phosphate mobilization 2005; Hollander et al., 2006; Wijnhoven et al., occurs strongly depends on the seasonality of 2006). Although the concentrations of most con- flooding. This has been clearly demonstrated by a taminants in the river water have substantially 173

30

25 waterlogged + S ) 1 − 20 waterlogged

15 control (µmol l (µmol − 2 4 10

HPO 5

0 0 5 10 15 20 5 0 C 20 0 C 20 0 C - drained 100 90 80 ) 1

− 70 60 50 (µmol l (µmol − 40

HS 30 20 10 0 0 5 10 15 20 week number Figure 3. Effect of temperature on the mobilization of phosphate and accumulation of sulphide in pore water as a result of water- logging without or with (S) 2 mmol l)1 sulphate. Mesocosm experiment with calcareous, sandy monoliths (4% organic matter) including their Junco-Molinion vegetation (Bruijnzeels & Lamers, unpublished data).

decreased during the last decades, floodplain soils ging or flooding) or acidic conditions favour the may still contain high concentrations of persistent dissolving of a number of metals. High total contaminants that can accumulate at different concentrations will have a much smaller effect in trophic levels (e.g., Hendriks & Pieters, 1993; calcareous soils than in more acidic soils, because Goodyear & McNeill, 1999). The soil profile may the metals are hardly mobile. Since desiccation of even provide a proxy for the estimation of historic soils generally leads to acidification through water qualities in terms of both P and contami- oxidation processes, including the oxidation of Fe nants, since the age of the different layers can be and S and nitrification (Lamers et al., 1998a; determined by Cs-137 dating (Middelkoop, 2000; Lucassen et al, 2002), the acid neutralizing Winter et al., 2001). For metals, however, the capacity (ANC) is a significant regulator of metal actual toxicity at a particular temperature is toxiciy. If the ANC is sufficiently high, acidificat- closely related to metal speciation (Florence et al., ion (i.e. loss of ANC; Van Breemen et al., 1983) 1992; Forstner, 1993; Tessier & Turner, 1995; does not lead to a drop in pH because all protons Vink, 2002; Schro¨der et al., 2005). While total are neutralized. For S-rich wetlands, therefore, the concentrations in soils may be high (suggesting ratio between the total concentration of S (the great toxicity risks), bioaccumulation may be acidifying component) and that of calcium + adventitiously low (Hobbelen et al., 2004). Metal magnesium (representing ANC) has proved to be a speciation is determined by redox conditions, pH useful diagnostic tool to assess toxicity problems and the concentration and type of dissolved during drainage (Lucassen et al., 2002). Values organic matter. Low redox potentials (waterlog- exceeding 0.7 (mol mol)1) indicate that desiccation 174 would lead to severe acidification and metal ‘resetting’ succession. In addition, the dispersal of mobilization. In conclusion, this means that data plant propagules is promoted, which is relevant in on the total metal concentrations (which are gen- relation to habitat fragmentation. In this way, erally used in policy-making programmes) have to biodiversity is warranted by disturbance caused by be combined with biogeochemical knowledge of hydrological fluctuations. For the upper Rhone the regulatory processes to allow toxicity to be river, it was indeed demonstrated that the most properly predicted (Van der Welle, in press). For a frequently flooded channel had the greatest mac- specific location, current toxicity, bioaccumulation rophyte biodiversity (Bornette et al., 1998). How- and biomagnification can be determined by mea- ever, from extensive field studies it is known that suring contaminant concentrations and physio- plant biodiversity does not necessarily increase logical variables in selected organisms (Kooistra linearly with flooding intensity. Some studies even et al., 2001, 2004), or by estimating the fraction found a negative effect, because the regional spe- that is readily taken up by organisms (Vink, 2002). cies pool of flooding-tolerant plants appeared to In addition to eutrophication and pollution by be relatively small (Grace & Pugesek, 1997; Grace metals and other toxicants, the accumulation of & Jutila, 1999; Olde Venterink et al., 2001). Other natural, potentially toxic substances such as studies showed that species richness was highest at sulphide and ammonium may provide an addi- intermediate levels of flooding (Day et al., 1988; tional problem for wetland development in Pollock et al., 1998). In a study along the lower polluted river systems. For both substances, even Rhine covering 215 floodplain lakes the effect of low concentrations are known to be phytotoxic to inundation appeared to be modest, probably helophyte and aquatic macrophyte communities because flooding during the growing season is rare (Roelofs, 1991; Van Wijck et al., 1992; Armstrong in this region (Van Geest et al., 2003). A stable et al., 1996; Smolders & Roelofs, 1996; Smolders water table in these lakes, however, appeared to et al., 1996; Lamers et al., 1998b). High concen- reduce biodiversity and to promote the undesired trations of iron are also known to be toxic to dominance of Elodea nuttallii (Van Geest et al., certain plants (Smolders et al., 1994; Lucassen 2005). Incidental temporal lowering of the water et al., 2000). For eutrophic floodplain vegetation, level is expected to enhance the ecological status of the accumulation of organic acids such as acetic these lakes, like in peat lakes (Lamers et al., acid might be an additional problem (Laanbroek 2002b). In an earlier study on floodplain lakes et al., 1984; Armstrong et al., 1996). The higher along the lower Rhine, it was also concluded that sensitivity to flooding of plants on organic, highly in-lake processes (including soil resuspension by reductive substrates as compared to mineral benthivorous fish) appeared to be more important substrates (Van den Brink et al., 1995) can be than river dynamics (Roozen et al., 2003). This explained by the higher oxygen consumption rate seems to hold particularly for less dynamic river in the soil and the accumulation of toxic systems, in contrast to highly dynamic systems. A substances mentioned above. complicating factor, not related to connectivity, is the fact that biodiversity can be very low in less dynamic backwaters (connected through back- Biogeochemistry and biodiversity flows) as a result of high local nutrient loading (Bornette et al., 1998; Boedeltje et al., 2005). At the highly dynamic locations near the river, In order to predict the above effects of changed increased connectivity is expected to lead to nutrient biogeochemistry, it is vital to know greater biodiversity as a result of increased whether the target vegetation types are limited by disturbance by flooding, by sedimentation and nutrient availability, and which of the elements erosion, and by increased dispersal potential is actually limiting biomass production for the (Bornette et al., 1998; Galat et al., 1998; Demars total vegetation, or for potentially invading fast & Harper, 2005). In this situation, highly growing species. competitive species are not able to become domi- In aquatic habitats, phosphate availability nant, and niches for regeneration of ruderal spe- initially controls biomass production, except in cies (Grime, 1974) are constantly created, thereby those characterized by extremely low N availabil- 175 ity (like softwater lakes). Increasing concentrations as well as the small-scale heterogeneity in resource of phosphate strongly promote the growth of availability, enable species coexistence and biodi- algae, decreasing light availability. Submerged versity (Tilman, 1994; Gu¨sewell, 2004). In the macrophytes and their associated fauna start to Biebrza Valley (Poland), P fertilization of flood- disappear at phosphate concentrations around plain fens resulted in increased growth of large 5 lmol l)1, and plants possessing floating leaves, sedges, grasses and non-graminoid herbs. Small such as Lemna, take over. At even higher sedges were shown to be outcompeted by large phosphate concentrations (>10 lmol l)1) high sedges because of the low light availability phytoplankton (green algae and cyanobacteria) (Kotowski, 2002). A review of the literature on concentrations develop, or the water becomes European fens showed that light availability was covered by Azolla filiculoides, a fern capable of N closely related to fertility, rather than to the fixation thanks to its symbiosis with cyanobacte- hydrological regime (Kotowski & Van Diggelen, ria. This means that these waters have changed 2004). Contradictory results may also be explained from P-limited to N-limited, providing a compet- by changes in the relative need for nutrients over itive advantage for N-fixing organisms. Eutrophic time. This was demonstrated by Van der Hoek floodplain lakes show clear phytoplankton domi- et al. (2004). After N addition, the total biomass of nance, suppressing submerged vegetation (Van a fen meadow vegetation increased only during the den Brink et al., 1993; Van den Brink & Van der first year of fertilization, without any effect of P Velde, 1994). The growth of the vegetation in fertilization. During the second year, total biomass biodiverse wet meadows, poor fens and rich fens, was unexpectedly boosted by P addition, while the now almost completely absent from Dutch riverine vegetation became dominated by the grass species areas, has been shown to be limited by either P or Holcus lanatus. At higher N availability, P uptake N, or both (Koerselman & Verhoeven, 1995; by the vegetation as a whole had increased, leading Wassen et al., 1995; Van Oorschot et al., 1998; to P shortage. Bedford et al., 1999; Wheeler & Proctor, 2000). In The type of nutrient limitation is strongly the floodplains of the Allier and Loire (France) regulated by the extent of drainage and the vegetation was found to be N-limited, as indicated mowing regime. Annual mowing of floodplains by the positive correlation with N mineralization along the Polish river Jegrznia has resulted in a (Van Oorschot et al., 1998). Species-rich fen shift from no limitation to (periodical) N and P grasslands in the floodplain of the less polluted limitation at slightly drained riparian locations, river Shannon were found to be co-limited by N and to P (and probably also K) limitation at the and P (Spink et al., 1998). heavily drained locations (De Mars et al., 1996). The nature of nutrient limitation is generally Mowing generally leads to a net export of P and K tested by conducting fertilization experiments, by (Koerselman & Verhoeven, 1995). correlation analysis, or by analyzing the N:P ratio The extremely high nutrient concentrations (reflecting the relative availability for a particular (both N and P) in floodplain soils along the rivers species). However, the results of fertilization Rhine and Meuse, due to heavy fertilization, are experiments mostly refer to the biomass response expected to lead to increased aboveground of the actual total vegetation to nutrient addition, biomass production, and thereby to loss of species. and not to the competitive advantage of fast- Highly competitive, tall and fast growing species growing species in the longer term. The differential may outcompete characteristic, more slowly effects of nutrient addition can partly be explained growing species (Day et al., 1988; Taylor et al., by differences in interspecific variation in the 1990). Vegetation types with aboveground uptake of N and P, or in the need for these two biomass values exceeding 500 g m)2 always show nutrients. In conditions of low P availability, small low plant species richness. The conservation of Carex spp. have a competitive advantage because many threatened wetland plant species requires they show higher nutrient-use efficiency than low productive P (co)-limited conditions (Olde species adapted to eutrophic conditions (Chapin, Venterink et al., 2001, 2003a; Wassen et al., 2005). 1980; Aerts & Chapin, 2000). The fact that indi- In Dutch floodplains, the average biomass was vidual species can be limited by different resources, found to be 1000 g m)2, with values up to 176

8000 g m)2 for Phragmites australis stands. Less der Velde, 1993) and aluminium (Rengel, 1996; De than one quarter of the locations monitored Graaf et al., 1997; Rout et al., 2001; Batty et al., showed a standing stock of 500 g m)2 or less 2002). As ammonium and aluminium appear to be (Antheunisse et al., unpublished data). far more toxic at low pH and low Ca concentra- Highly productive species-poor vegetation tion, the toxicity of these compounds will be more types such as those dominated by Glyceria maxima pronounced at locations with a low ANC. In fen (Europe) or Scirpus fluviatilis (USA) appear not to meadows, ammonium and aluminium toxicity be limited by nutrients (Wassen et al., 1995, 1998, have been shown to be quite detrimental to a 2003; Spink et al., 1998). In eutrophic Dutch and number of characteristic, vulnerable species (De American floodplains, nutrient availability was Graaf et al., 1998; Lucassen et al., 2003; Van den found to be a less important regulator of plant Berg et al., 2005). However, high ammonium production than other factors such as flooding and concentrations in anaerobic soils (in which nitrifi- drought (Spink et al., 1998). In this case, nutrient cation rates are low) may also be toxic to aquatic availability as the controlling factor of plant macrophytes in calcareous waters (Onaindia et al., growth is superseded by hydrological variables 1996; Smolders et al., 1996; Pezeshki, 2001). (e.g. flooding, current, sedimentation and drought) Therefore, species biodiversity is expected to be and genetic constraints (including flood tolerance regulated by concentrations of both nutrients and and maximum growth rate). High biomass toxicants. Depending on speciation (regulated by production rates reinforce nutrient accumulation, pH, dissolved organic matter and redox state), due to increased accumulation of nutrients ex- high metal concentrations may accumulate in both tracted from the soil and enhanced sedimentation. riparian and aquatic macrophytes. This stabilizes the existing state and impedes the transition to a state controlled by nutrient avail- ability. For eutrophic floodplains, we expect that From research to policy and measures the rehabilitation of vegetation types dominated by slow-growing species, which show high nutri- As discussed above, we expect several biogeo- ent-use efficiency and a low physiological ability to chemical problems to arise when Dutch rivers are respond to eutrophication, will only be possible allocated more space, and new hydrological after additional measures, as explained in the next regimes are established. We think that it will not section. In slightly acidic grassland abandoned by be possible to simply extrapolate the ecological agriculture, the fast growing common rush, Juncus developments in the more pristine river systems effusus, represents a major frustration to chosen as a reference to greatly disturbed river conservationists who are trying to convert arable systems such as those in the Netherlands. land to biodiverse wetland by raising water tables. Phosphate-poor conditions for the develop- High phosphate availability in the topsoil impedes ment of biodiverse wetland vegetation can only be competition by more slowly growing species created by either sod cutting or soil removal. To (Smolders et al., unpublished data). estimate the biological availability of P, extraction In addition to nutrients, high concentrations of procedures (e.g. Olsen et al., 1954) provide a better toxic organic and inorganic pollutants are also estimate than the total concentrations. Olsen-P expected to impact on floodplain biodiversity. concentrations are very high in Dutch floodplains Although most studies refer to effects on fauna along the rivers Rhine, Meuse and IJssel, compa- communities, vegetation development may also be rable to nature development regions outside river regulated by toxicity. Potentially phytotoxic com- areas (Fig. 4a). Even for Polish floodplains along pounds include iron (Snowden & Wheeler, 1995; the Bug and Narew, P concentrations have been Lucassen et al., 2000; Batty et al., 2002), sulphide found to be quite high, while those for the Odra (Roelofs, 1991; Armstrong et al., 1996; Smolders floodplains were extremely high, due to fertiliza- et al., 1996; Lamers et al., 1998b), ammonium (e.g. tion or pollution from the river. Efforts to deplete Roelofs et al., 1985; Boxman & Roelofs, 1988; the strongly eutrophicated soils by mowing and Smolders et al., 1996; De Graaf et al., 1998; Luc- biomass removal are inadequate. Total P concen- assen et al., 2003), chloride (Van den Brink & Van trations in the soils are, on average, 10–20 times 177 the Olsen-P values. This means that if one wanted excavation. In combination with this analysis, to achieve the phosphate concentrations that are rejuvenation measures may meet extra targets by characteristic of biodiverse reference areas eliminating the eutrophication constraint imposed (amounting to 250 lmol Olsen-P kg)1 DW; by P enrichment of floodplain soils. 2500 lmol total P kg)1 DW), it might take Nitrate is generally a less important problem 100–500 years to create the desired conditions. than phosphate, due to the relatively high mobility This is generally well beyond the scope of policy- of this nutrient (Fig. 5). N concentrations in soil makers, nature managers and water managers. A moisture of Dutch floodplains are generally depth profile for P will indicate the required between 0 and 100 lmol l)1 for ammonium, and 0 removal depth, which is often not equivalent to the and 50 lmol l)1 for nitrate, which is moderate. depth of ploughing. For many habitat creation Both in Poland and in the Netherlands, however, areas, however, phosphate saturation is too deep extremely high values, up to the millimolar range, and soil removal therefore too costly (Fig. 4b). for nitrate (drier soils) and ammonium (wetter soils) This means that the above-mentioned have been measured (Loeb et al., unpublished development of P-limited, biodiverse vegetation results). At locations with seepage of nitrate-rich types is impossible. In this case, regular sod cut- groundwater, soil removal will not exhaust the ting, removing highly competitive species, may available nitrate. At these high N concentrations, favour early successional stages. As deeper soil N-based eutrophication is very likely. As discussed, layers in floodplains may still contain high con- the relative availability of one nutrient will influence centrations of phosphate due to historic loading the response to another. (Winter et al., 2001), it is vital to analyze soil The ecological rehabilitation of floodplains profiles before starting rehabilitation measures like may also be frustrated by soil contamination. As

12000 12000 (a) (b) 11000 11000

10000 10000

9000 9000 )

8000 1 8000 −

7000 7000 ) in topsoil 1 − 6000 6000 Excessive Excessive 5000 5000

4000 Olsen-P (µmol kg 4000 Very high Very high 3000 3000 Olsen-P (µmol kg High High 2000 2000

1000 1000

Df. Df. 0 0 Ref012345678910111213 ND OV OM ME RH W SW GE SH OD BU NA Ref0123456 0-2020-40 40-50 50-60 60-70 Location ND, Depth below surface (cm) Figure 4. (a) Estimation of phosphorus availability in the upper 10–20 cm of sediments and sediments by bicarbonate extraction (Olsen et al., 1954), for various locations in the Netherlands, Poland and Ireland for riverine soils, grey dots represent organic soils, white dots mineral soils. Many locations show high P availability, some even excessive. Df.: P deficiency plausible. Ref: reference biodiverse wetlands; ND: non-alluvial, former agricultural areas destined for nature development (Smolders, unpubl. data). All other locations are alluvial wetlands, most of which were formerly or are currently in agricultural use (Antheunisse, unpublished data; Loeb, unpublished data). River codes: OV: Overijsselse Vecht, the Netherlands (NL); OM: Oude Maas (NL); ME: Meuse (NL); RH: Rhine (NL); W: Waal (NL); SW: Swalm (NL); GE: Geulle (NL); SH: Shannon, Ireland; OD: Odra, Poland (PL); BU: Bug (PL); NA: Narew (PL). (b) Olsen P values for the different depth classes at the ND locations of panel A (Smolders, unpublished data). Defic.: P deficiency plausible. Ref: reference biodiverse wetlands. Based on these estimates, decisions can be made on the extent of sediment removal needed to restore nutrient-poor conditions in ecological rehabilitation programmes. 178

measures require an understanding of biogeo- chemical and biological key processes in riverine areas (Buijse et al., 2002) and other wetland types (Bedford, 1999). Unlike a trial and error approach, this will enable managers and policy- makers to understand the factors controlling degradation or successful restoration (Bedford, 1999; Lamers et al., 2002b; Buijse et al., 2005). Field observations providing correlative infor- mation need to be coupled to experimental data, in order to find causal relationships. Once the key variables and processes have been assessed, diagnoses can be established and results for one location can be extrapolated to other areas. In addition to biogeochemical assessments, vegeta- tion can also be used for an ecological diagnosis of riverine wetlands, because the occurrence and abundance of wetland plant species are highly indicative of their habitat characteristics (Bloe- mendaal & Roelofs, 1988; Demars & Harper, 1998; Dawson & Szoszkiewicz, 1999; Amoros ) Figure 5. Concentrations of plant-available nitrate (NO3 ;H2O et al., 2000; Buijse et al., 2005). Moreover, it 2) extraction) and phosphate (HPO4 ; lactate/acetate extraction) should be possible to predict the degree of suc- in sediments, for different depth classes; comparison between an arable land location, former arable land and reference (biodi- cess for a given location, and to make sound verse nature reserve). All three are recharge areas in terms of choices between potential new areas combining ) hydrology. NO3 concentrations have decreased to reference water safety and ecological targets (risk analy- values after cessation of agricultural activites 10 years ago while sis). In order to be able to include the spatial 2) HPO4 concentrations remain high. (From: Smolders et al., variability within riverine systems, geographic unpublished data). information systems applications are indispens- able (Poiani & Bedford, 1995; Kooistra et al., concentrations of a large number of contaminants 2001). Rehabilitation projects based on trial and were extremely high during the 1970s and have error show a very unfavourable cost-benefit since decreased in the surface water, excavation ratio, which is undesirable given the high costs may well uncover contaminated layers (Winter of ecological rehabilitation. et al., 2001). The targets of ecological rehabilitation are not only related to plant species diversity, but also to Conclusions: how much ecological space does the ecosystem diversity and riverine landscape diver- ‘straitjacket’ provide? sity (Amoros, 2001; Ward et al., 2002). This means that measures should be taken at these different Highly regulated rivers only provide a narrow levels. The biogeochemical constraints we have area to accommodate an elaborate mosaic of discussed operate not only at the species and wetland types, including sedge fens and carr ecosystem level, but also at the landscape level. woods. In the past, high inputs of nutrients and For instance, vegetation development interferes contaminants caused a gradually declining with geohydrology at the landscape level, urging nutrient gradient, depending on connectivity. the need for preservation and rehabilitation of The heavy fertilization of floodplain soils and of biodiversity at higher scales (Sparks, 1995; Tock- groundwater feeding floodplain wetlands with N ner et al., 1998; Godreau et al., 1999). and P impose a serious constraint on the devel- The identification of possible constraints and opment of wetland types controlled by nutrient the prediction of the outcome of restoration availability. For these areas, managers have to 179 be satisfied with lower ecological targets, or take Present state and signs of ecological recovery. Hydrobiologia additional measures such as top layer removal. 265: 97–128. The need for this may be partly obviated by Aerts, R. & F. S. Chapin, 2000. The mineral nutrition of wild plants revisited: A re-evaluation of processes and patterns. removing soil as a part of rejuvenation measures, Advances in Ecological Research 30: 1–67. provided that newly exposed soil layers do not Amoros, C., 2001. The concept of habitat diversity between and contain high concentrations of phosphate or within ecosystems applied to river side-arm restoration. contaminants as a legacy from past eutrophica- Environmental Management 28: 805–817. tion and pollution. For floodplain lakes, water Amoros, C., G. Bornette & C. P. Henry, 2000. A vegetation- based method for ecological diagnosis of riverine wetlands. level fluctuation seems to be vital to reduce Environmental Management 25: 211–227. eutrophication and increase biodiversity. In or- Appelo, C. A. J. & D. Postma, 1993. Geochemistry, Ground- der to tackle nitrate and sulphate pollution via water and Pollution. A.A. Balkema Publishers, Rotterdam. groundwater seepage, measures have to be taken Armstrong, J., F. Afreen-Zobayed & W. Armstrong, 1996. at the landscape scale. For those communities Phragmites die-back: sulphide- and acetic acid induced bud and root death, lignifications, and blockages within aeration primarily controlled by flooding regime (alter- and vascular systems. New Phytologist 134: 601–614. nating flooded and desiccated conditions), Armstrong, W., R. Bra¨ndle & M. B. Jackson, 1994. Mecha- nutrient biogeochemistry will be of minor nisms of flood tolerance in plants. Acta Botanica Neerlan- importance, because flooding regimes overrule all dica 43: 307–358. other factors. Their biodiversity is mainly con- Baker, M. A. & P. Vervier, 2004. Hydrological variability, matter supply and denitrification in the Garonne River trolled by disturbance, either positively or nega- ecosystem. Freshwater Biology 49: 181–190. tively. Next to eutrophication, intoxication by Bal, D., H. M. Beije, M. Fellinger, R. Haveman, A. J. F. M. (micro)pollutants imposes an additional con- van Opstal & F. J. van Zadelhoff, 2001. Handboek Na- straint. After removal of the top layer, soils may tuurdoeltypen. Tweede, Geheel Herziene Druk. EC-LNV, even show higher concentrations as a heritage Wageningen. Baldwin, D. S. & A. M. Mitchell, 2000. The effects of drying from a more polluted past. and re-flooding on the sediment and soil nutrient dynamics We have argued that a trial and error ap- of lowland river-floodplain systems: a synthesis. Regulated proach will not be able to predict the ecological Rivers: Research and Management 16: 457–467. development of particular floodplain locations, Baptist, M. J., W. E. Penning, H. Duel, A. J. M. Smits, G. W. nor will it be possible to select those locations Geerling, G. E. M. van der Lee & J. S. L. van Alphen, 2004. Assessment of the effects of cyclic floodplain rejuvenation on that are most likely to be successful in terms of flood levels and biodiversity along the Rhine river. River ecological rehabilitation. Biogeochemical knowl- Research and Applications 20: 285–297. edge is a prerequisite for successful ecological Batty, L. C., A. J. M. Baker & B. D. Wheeler, 2002. Aluminium rehabilitation, ensuring that space for the river and phosphate uptake by Phragmites australis: the role of also implies space for ecologically sound flood- Fe, Mn and Al root plaques. Annals of Botany 89: 443–449. Bayley, P. B., 1991. The flood pulse advantage and the resto- plains. ration of river-floodplain systems. Regulated Rivers: Research and Management 47: 75–86. Bedford, B. L., 1999. Cumulative effects on wetland landscapes: Acknowledgements links to wetland restoration in the United States and southern Canada. Wetlands 19: 775–788. Bedford, B. L., M. R. Walbridge & A. Aldous, 1999. Patterns in The Netherlands Organisation for Scientific Re- nutrient availability and plant diversity of temperate North search (NWO-TRIAS) and the Limburgs Land- American wetlands. Ecology 80: 2151–2169. schap Foundation funded part of the research Beltman, B., T. G. Rouwenhorst, M. B. van Kerkhoven, T. van presented in this paper. Jan Klerkx provided lin- der Krift & J. T. A. Verhoeven, 2000. Internal eutrophica- guistic advice. This is CWE publication 422. tion in peat soils through competition between chloride and sulphate with phosphate for binding sites. Biogeochemistry 50: 183–194. Bij de Vaate, A., R. Breukel & G. van der Velde, 2006. Long- term developments in the ecological rehabilitation of the References main distributaries in the Rhine Delta: fish and macroin- vertebrates. Hydrobiologia 565: 229–242. Admiraal, W., G. van der Velde, H. Smit & W. G. Cazemier, Bloemendaal, F. H. J. L. & J. G. M. Roelofs, (eds), 1988. 1993. The rivers Rhine and Meuse in the Netherlands – Waterplanten en Waterkwaliteit. Stichting Uitgeverij 180

Koninklijke Nederlandse Natuurhistorische Vereniging, Buijse, A. D., F. Klijn, R. S. E. W. Leuven, H. Middelkoop, Utrecht (in Dutch). F. Schiemer, J. H. Thorp & H. P. Wolfert, 2005. Reha- Blom, C. W. P. M., L. A. C. J. Voesenek, M. Banga, W. M. H. bilitation of large rivers: references, achievements and G. Engelaar, J. H. G. M. Rijnders, H. M. van de Steeg & integration into river management. Archiv fu¨r Hydrobiol- E. J. W. Visser, 1994. Physiological ecology of riverside ogie Supplement 155 – Large Rivers 15: 715–738. species – adaptive responses of plants to submergence. Burt, T. P., L. S. Matchett, K. W. T. Goulding, C. P. Webster & Annals of Botany 74: 253–263. N. E. Haycock, 1999. Denitrification in riparian buffer Boedeltje, G., J. P. Bakker, A. ten Brinke, J. M. van Groen- zones: the role of floodplain hydrology. Hydrological endael & M. Soesbergen, 2004. Dispersal phenology of Processes 13: 1451–1463. hydrochorous plants in relation to discharge, seed release Cals, M. J. R., R. Postma, A. D. Buijse & E. C. L. Marteijn, time and buoyancy of seeds: the flood pulse concept 1998. Habitat restoration along the River Rhine in The supported. Journal of Ecology 92: 786–796. Netherlands: Putting ideas into practice. Aquatic Conser- Boedeltje, G., A. J. P. Smolders, L. P. M. Lamers & J. G. M. vation – Marine and Freshwater Ecosystems 8: 61–70. Roelofs, 2005. Interactions between sediment propagule Caraco, N. F., J. J. Cole & G. E. Likens, 1989. Evidence for banks and sediment nutrient fluxes explain floating plant sulphate-controlled phosphorus release from sediments of dominance in stagnant shallow waters. Archiv fu¨r Hydro- aquatic systems. Nature 341: 156–158. biologie 162: 349–362. Caraco, N. F., J. J. Cole & G. E. Likens, 1993. Sulfate control Bornette, G. & C. Amoros, 1996. Disturbance regimes and of phosphorus availability in lakes – a test and reevaluation vegetation dynamics: Role of floods in riverine wetlands. of Hasler and Einsele model. Hydrobiologia 253: 275–280. Journal of Vegetation Science 7: 615–622. Carbiener, R., M. Tre` molieres, J. L. Mercier & A. Ortscheit, Bornette, G., C. Amoros & N. Lamouroux, 1998. Aquatic plant 1990. Aquatic macrophyte communities as bioindicators of diversity in riverine wetlands: the role of connectivity. eutrophication in calcareous oligosaprobe stream waters Freshwater Biology 39: 267–283. (Upper Rhine plain, Alsace). Vegetatio 86: 71–88. Bornette, G., M. Guerlesquin & C. P. Henry, 1996. Are the Castro, H., K. R. Reddy & A. Ogram, 2002. Composition and Characeae able to indicate the origin of groundwater in function of sulfate-reducing prokaryotes in eutrophic and former river channels? Vegetatio 125: 207–222. pristine areas of the Florida Everglades. Applied and Envi- Bornette, G., H. Piegay, A. Citterio, C. Amoros & V. Godreau, ronmental Microbiology 68: 6129–6137. 2001. Aquatic plant diversity in four river floodplains: a CBS, 2004. Productie van mest en mineralen. http://www.stat- comparison at two hierarchical levels. Biodiversity and line.cbs.nl. CBS, Voorburg/Heerlen. Conservation 10: 1683–1701. Chapin, F. S. I., 1980. The mineral nutrition of wild plants. Bostro¨m, B., M. Jansson & C. Forsberg, 1982. Phosphorus re- Annual Review of Ecology and Systematics 11: 233–260. lease from lake sediments. Archiv fu¨r Hydrobiologie 18: 5–59. Christian, R. R., W. L. Bryant & M. M. Brinson, 1990. Juncus Boxman, A. W., 2005. Effecten van Verminderde Stikstof- roemerianus production and decomposition along gradients Depositie op een Grove Dennen Opstand in Natuurgebied of salinity and hydroperiod. Marine Ecology – Progress ‘De Rouwkuilen’, Limburg. Report by order of the Ministry Series 68: 137–145. of Agriculture, Nature & Food Quality. Radboud University Coops, H., N. Geilen & G. van der Velde, 1999. Helophyte Nijmegen, Nijmegen (in Dutch). zonation in two regulated estuarine areas in the Netherlands: Boxman, A. W., H. H. Bartelink, Ph. Bossenbroek, R. H. vegetation analysis and relationships with hydrological Kemmers & A. H. F. Stortelder, 2003. Uitvoering van factors. Estuaries 22: 657–668. Vernattingsmaatregelen op Praktijkschaal 1997–2003; Re- Coops, H., N. Geilen, H. J. Verheij, R. Boeters & G. van der ferentieproject Koelbroek. EC-LNV, Ede (in Dutch). Velde, 1996b. Interactions between waves, bank erosion and Boxman, A. W. & J. G. M. Roelofs, 1988. Some effects of emergent vegetation: an experimental study in a wave tank. nitrate versus ammonium nutrition on the nutrient fluxes in Aquatic Botany 53: 187–198. Pinus sylvestris seedlings. Effects of mycorrhizal infection. Coops, H., F. W. B. van den Brink & G. van der Velde, 1996a. Canadian Journal of Botany 66: 1091–1097. Growth and morphological responses of four helophyte Bridgham, S. D. & C. J. Richardson, 2003. Endogenous versus species in an experimental water-depth gradient. Aquatic exogenous nutrient control over decomposition and miner- Botany 54: 11–24. alization in North Carolina peatlands. Biogeochemistry 65: Coops, H. & G. van der Velde, 1995. Seed dispersal, germina- 151–178. tion and seedling growth of six helophyte species in relation Brock, T. C. M., G. van der Velde & H. M. van de Steeg, to water-level zonation. Freshwater Biology 34: 13–20. 1987. The effects of extreme water level fluctuations on the Coops, H. & G. van der Velde, 1996. Effects of waves on wetland vegetation of a nymphaeid-dominated oxbow lake helophyte stands: mechanical characteristics of stems of in the Netherlands. Archiv fu¨r Hydrobiologie Supplement Phragmites australis and Scirpus lacustris. Aquatic Botany 27: 57–73. 53: 175–185. Buijse, A. D., H. Coops, M. Staras, L. H. Jans, G. J. van Geest, Crawford, R. M. M., 1987. Plant Life in Aquatic and Amphibi- R. E. Grift, B. W. Ibelings, W. Oosterberg & F. C. J. M. ous Habitats. Blackwell Science Publication, Oxford. Roozen, 2002. Restoration strategies for river floodplains Crawford, R. M. M., 2003. Seasonal differences in plant along large lowland rivers in Europe. Freshwater Biology 47: responses to flooding and anoxia. Canadian Journal of 889–907. Botany 81: 1224–1246. 181

Darke, A. K. & M. R. Walbridge, 2000. Al and Fe biogeo- Geerling, G. W., A. M. J. Ragas, R. S. E. W. Leuven, J. H. van chemistry in a floodplain forest: implications for P retention. den Berg, M. Breedveld, D. Liefhebber & A. J. M. Smits, Biogeochemistry 51: 1–32. 2006. Succession and rejuvenation in floodplains along the Dawson, F. H. & K. Szoszkiewicz, 1999. Relationships of some River Allier (France). Hydrobiologia 565: 71–86. ecological factors with the associations of vegetation in Gocht, T., K. M. Moldenhauer & W. Puttmann, 2001. His- British rivers. Hydrobiologia 415: 117–122. torical record of polycyclic aromatic hydrocarbons (PAH) Day, R. T., P. A. Keddy, J. McNeill & T. Carleton, 1988. and heavy metals in floodplain sediments from the Rhine Fertility and disturbance gradients: a summary model for River (Hessisches Ried, Germany). Applied Geochemistry riverine marsh vegetation. Ecology 69: 1044–1054. 16: 1707–1721. De Graaf, M. C. C., R. Bobbink, J. G. M. Roelofs & P. J. Godreau, V., G. Bornette, B. Frochot, C. Amoros, E. Castella, M. Verbeek, 1998. Differential effects of ammonium B. Oertli, F. Chambaud, D. Oberti & E. Craney, 1999. and nitrate on three heathland species. Plant Ecology 135: Biodiversity in the floodplain of Saoˆ ne: a global approach. 185–196. Biodiversity and Conservation 8: 839–864. De Graaf, M. C. C., R. Bobbink, P. J. M. Verbeek & J. G. M. Golterman, H. L., 1998. The distribution of phosphate over Roelofs, 1997. Aluminium toxicity and tolerance in three iron-bound and calcium-bound phosphate in stratified sedi- heathland species. Water, Air and Soil Pollution 98: 229– ments. Hydrobiologia 364: 75–81. 239. Goodyear, K. L. & S. McNeill, 1999. Bioaccumulation of Demars, B. O. L. & D. M. Harper, 1998. The aquatic macro- heavy metals by aquatic macro-invertebrates of different phytes of an English lowland river system: assessing response feeding guilds: a review. Science of the Total Environment to nutrient enrichment. Hydrobiologia 384: 75–88. 229: 1–19. Demars, B. O. L. & D. M. Harper, 2005. Distribution of Grace, J. B. & H. Jutila, 1999. The relationship between species aquatic vascular plants in lowland rivers: separating the density and community biomass in grazed and ungrazed effects of local environmental conditions, longitudinal coastal meadows. Oikos 85: 398–408. connectivity and river basin isolation. Freshwater Biology Grace, J. B. & B. H. Pugesek, 1997. A structural equation 50: 418–437. model of plant species richness and its application to a De Mars, H., M. J. Wassen & W. H. M. Peeters, 1996. The coastal wetland. The American Naturalist 149: 436–460. effect of drainage and management on peat chemistry and Grime, J. P., 1974. Vegetation classification by reference to nutrient deficiency in the former Jegrznia-floodplain strategies. Nature 250: 26–31. (NE-Poland). Vegetatio 126: 59–72. Gu¨sewell, S., 2004. (Tansley review) N : P ratios in terrestrial De Mars, H., M. J. Wassen & H. O. Venterink, 1997. Flooding plants: variation and functional significance. New Phytolo- and groundwater dynamics in fens in eastern Poland. Jour- gist 164: 243–266. nal of Vegetation Science 8: 319–328. Hall, L. W. & R. D. Anderson, 1995. The influence of salinity Den Hartog, C., 1974. Brackish-water classification, its on the toxicity of various classes of chemicals to aquatic development and problems. Hydrobiological Bulletin 8: 15– biota. Critical Reviews in Toxicology 25: 281–346. 28. Harper, D., C. Smith, P. Barham & R. Howell, 1995. The De Waal, L. C., A. R. G. Large, C. J. Grippel & P. M. Wade, ecological basis for the management of the natural river 1995. River and floodplain rehabilitation in Western Europe: environment. In Harper, D. M. & A. J. D. Ferguson (eds), opportunities and constraints. Archiv fu¨r Hydrobiologie The Ecological Basis for River Management. Wiley, Supplement 101 – Large Rivers 9: 679–693. Chichester, 219–238. Florence, T. M., G. M. Morrison & J. L. Stauber, 1992. Heathwaite, A. L., 1990. The effect of drainage on nutrient Determination of trace-element speciation and the role of release from fen peat and its implications for water-quality – speciation in aquatic toxicity. Science of the Total Envi- a laboratory simulation. Water, Air and Soil Pollution 49: ronment 125: 1–13. 159–173. Forstner, U., 1993. Metal speciation – General concepts and Hefting, M., J. C. Clement, D. Dowrick, A. C. Cosandey, S. applications. International Journal of Environmental Ana- Bernal, C. Cimpian, A. Tatur, T. P. Burt & G. Pinay, 2004. lytical Chemistry 51: 5–23. Water table elevation controls on soil nitrogen cycling in Freeman, C., G. Liska, N. J. Ostle, M. A. Lock, B. Reynolds & riparian wetlands along a European climatic gradient. J. Hudson, 1996. Microbial activity and enzymic decompo- Biogeochemistry 67: 113–134. sition processes following peatland water table drawdown. Heiler, G., T. Hein, F. Schiemer & G. Bornette, 1995. Hydro- Plant and Soil 180: 121–127. logical connectivity and flood pulses as the central aspects Galat, D. L., L. H. Fredrickson, D. D. Humburg, K. J. for the integrity of a river-floodplain system. Regulated Bataille, J. R. Bodie, J. Dohrenwend, G. T. Gelwicks, Rivers: Research and Management 11: 351–361. J. E. Havel, D. L. Helmers, J. B. Hooker, J. R. Jones, Hendriks, A. J., W. C. Ma, J. J. Brouns, E. M. De Ruiter-Dijkman M. F. Knowlton, J. Kubisiak, J. Mazourek, A. C. & R. Gast, 1995. Modeling and monitoringorganochlorine and McColpin, R. B. Renken & R. D. Semlitsch, 1998. heavy-metal accumulation in soils, earthworms, and shrews in Flooding to restore connectivity of regulated, large-river Rhine-delta floodplains. Archives of Environmental Contam- wetlands – natural and controlled flooding as comple- ination and Toxicology 29: 115–127. mentary processes along the lower Missouri River. Bio- Hendriks, A. J. & H. Pieters, 1993. Monitoring concentrations science 48: 721–733. of microcontaminants in aquatic organisms in the Rhine 182

delta – a comparison with reference values. Chemosphere 26: Koerselman, W., M. B. van Kerkhoven & J. T. A. Verhoeven, 817–836. 1993. Release of inorganic N, P and K in peat soils – effect of Henry, C. P., C. Amoros & G. Bornette, 1996. Species traits temperature, water chemistry and water-level. Biogeochem- and recolonization processes after flood disturbances in istry 20: 63–81. riverine macrophytes. Vegetatio 122: 13–27. Koerselman, W. & J. T. A. Verhoeven, 1995. Eutrophication of Hobbelen, P. H. F., J. E. Koolhaas & C. A. M. van Gestel, fen ecosystems: external and internal nutrient sources 2004. Risk assessment of heavy metal pollution for detriti- and restoration strategies. In Wheeler, B. D., S. C. Shaw, W. vores in floodplain soils in the Biesbosch, the Netherlands, J. Fojt & R. A. Robertson (eds), Restoration of Temper- taking bioavailability into account. Environmental Pollution ate Wetlands. John Wiley and Sons Ltd, Chichester, 91– 129: 409–419. 112. Hoffmann, C. C., M. L. Pedersen, B. Kronvang & L. Ovig, Kohler, A., R. Wonneber & G. Zeltner, 1973. Chemical data 1998. Restoration of the rivers Brede, Cole and Skerne: a and aquatic vascular plants as indicators for pollution in joint Danish and British EU-LIFE demonstration project, Moosach River system near Munich. Archiv fu¨r Hydrobi- IV – Implications for nitrate and iron transformation. ologie 72: 533–549. Aquatic Conservation – Marine and Freshwater Ecosystems Kok, C. J., G. van der Velde & K. M. Landsbergen, 1990. 8: 223–240. Production, nutrient dynamics and initial decomposition of Hogan, D. M., T. E. Jordan & M. R. Walbridge, 2004. Phos- floating leaves of Nymphaea alba L and Nuphar lutea (L) Sm phorus retention and soil organic carbon in restored and (Nymphaeaceae) in alkaline and acid waters. Biogeochem- natural freshwater wetlands. Wetlands 24: 573–585. istry 11: 235–250. Hollander, A., M. A. J. Huijbregts, A. M. J. Ragas & D. van de Ko¨lle, W., O. Strebel & J. Bo¨ttcher, 1985. Formation of sulfate Meent, 2006. BasinBox: a generic multimedia fate model for by microbial denitrification in a reducing aquifer. Water predicting the fate of chemicals in river catchments. Hy- Supply 3: 33–40. drobiologia 565: 21–38. Kooistra, L., R. S. E. W. Leuven, P. H. Nienhuis, R. Wehrens Hupp, C. R., 1992. Riparian vegetation recovery patterns fol- & L. M. C. Buydens, 2001. A procedure for incorporating lowing stream channelization – a geomorphic perspective. spatial variability in ecological risk assessment of Dutch Ecology 73: 1209–1226. river floodplains. Environmental Management 28: 359–373. Hyacinthe, C. & P. van Cappellen, 2004. An authigenic iron Kooistra, L., E. A. L. Salas, J. G. P. W. Clevers, R. Wehrens, phosphate phase in estuarine sediments: composition, R. S. E. W. Leuven, P. H. Nienhuis & L. M. C. Buydens, formation and chemical reactivity. Marine Chemistry 91: 2004. Exploring field vegetation reflectance as an indicator of 227–251. soil contamination in river floodplains. Environmental ICPR, 2005. International Commission for the Protection of Pollution 127: 281–290. the Rhine, http://www.iksr.org/. Kotowski W., 2002. Fen communities. Ecological mechanisms Jackson, M. B. & M. C. Drew, 1984. Effects of flooding on and conservation strategies. Ph.D Thesis, University of growth and metabolism of herbaceous plants. In Kozlowski, Groningen, Groningen. T. T. (ed.), Flooding and Plant Growth. Academic Press Kotowski, W. & R. van Diggelen, 2004. Light as an environ- Inc, New York, 47–128. mental filter in fen vegetation. Journal of Vegetation Science Japenga, J. & W. Salomons, 1993. Dyke-protected floodplains – 15: 583–594. a possible chemical time bomb. Land Degradation and Kronvang, B., L. M. Svendsen, A. Brookes, K. Fisher, B. Rehabilitation 4: 373–380. Moller, O. Ottosen, M. Newson & D. Sear, 1998. Restora- Japenga, J., K. H. Zschuppe, A. J. de Groot & W. Salomons, tion of the rivers Brede, Cole and Skerne: a joint Danish and 1990. Heavy-metals and organic micropollutants in flood- British EU-LIFE demonstration project, III – channel plains of the River Waal, a distributary of the River Rhine, morphology, hydrodynamics and transport of sediment and 1958–1981. Netherlands Journal of Agricultural Science 38: nutrients. Aquatic Conservervation – Marine and Freshwa- 381–397. ter Ecosystems 8: 209–222. Junk, W. J., P. B. Bayley & R. E. Sparks, 1989. The floodpulse Laanbroek, H. J., 1990. Bacterial cycling of minerals that affect concept in river floodplain systems. Canadian Journal of plant-growth in waterlogged soils – A review. Aquatic Fisheries and Aquatic Sciences 106: 110–127. Botany 38: 109–125. Kester, R. A., M. E. Meijer, J. A. Libochant, W. de Boer & H. Laanbroek, H. J., H. J. Geerligs, L. Sijtsma & H. Veldkamp, J. Laanbroek, 1997. Contribution of nitrification and deni- 1984. Competition for sulfate and ethanol among Desul-

trification to the NO and N2O emissions of an acid forest fobacter, Desulfobulbus, and Desulfovibrio species isolated soil, a river sediment and a fertilized grassland soil. Soil from intertidal sediments. Applied and Environmental Biology and Biochemistry 29: 1655–1664. Microbiology 47: 329–334. Khalid, R. A., W. H. Patrick & R. D. DeLaune, 1977. Phos- Laanbroek, H. J. & N. Pfennig, 1981. Oxidation of short-chain phorus sorption characteristics of flooded soils. Soil Science fatty-acids by sulfate-reducing bacteria in fresh-water and in Society of America Journal 41: 305–310. marine-sediments. Archives of Microbiology 128: 330–335. Knowlton, M. F. & J. R. Jones, 1997. Trophic status of Lamers, L. P. M., S. J. Falla, E. M. Samborska, L. A. R. van Missouri River floodplain lakes in relation to basin type and Dulken, G. van Hengstum & J. G. M. Roelofs, 2002a. connectivity. Wetlands 17: 468–475. Factors controlling the extent of eutrophication and toxicity 183

in sulfate-polluted freshwater wetlands. Limnology and Lucassen, E. C. H. E. T., A. J. P. Smolders & J. G. M. Oceanography 47: 585–593. Roelofs, 2002. Potential sensitivity of mires to drought, Lamers, L. P. M., S. M. E. van Roozendaal & J. G. M. Roelofs, acidification and mobilisation of heavy metals: the sedi- 1998a. Acidification of freshwater wetlands: combined ment S/(Ca+Mg) ratio as diagnostic tool. Environmental effects of non-airborne sulfur pollution and desiccation. Pollution 120: 635–646. Water, Air and Soil Pollution 105: 95–106. Lucassen, E. C. H. E. T., A. J. P. Smolders, J. van de Crom- Lamers, L. P. M., A. J. P. Smolders & J. G. M. Roelofs, 2002b. menacker & J. G. M. Roelofs, 2004a. Effects of stagnating The restoration of fens in the Netherlands. Hydrobiologia sulphate-rich groundwater on the mobility of phosphate in 478: 107–130. freshwater wetlands: a field experiment. Archiv fu¨r Hydro- Lamers, L. P. M., G. E. ten Dolle, S. T. G. van den Berg, S. P. biologie 160: 117–131. J. van Delft & J. G. M. Roelofs, 2001. Differential responses Lucassen, E. C. H. E. T., A. J. P. Smolders, A. L. van der Salm of freshwater wetland soils to sulphate pollution. Biogeo- & J. G. M. Roelofs, 2004b. High groundwater nitrate chemistry 55: 87–102. concentrations inhibit eutrophication of sulphate-rich Lamers, L. P. M., H. B. M. Tomassen & J. G. M. Roelofs, freshwater wetlands. Biogeochemistry 67: 249–267. 1998b. Sulfate-induced eutrophication and phytotoxicity in Mendelssohn, I. A., B. K. Sorrell, H. Brix, H. H. Schierup, freshwater wetlands. Environmental Science and Technology B. Lorenzen & E. Maltby, 1999. Controls on soil cellu- 32: 199–205. lose decomposition along a salinity gradient in a Phrag- Lenssen, J., F. Menting, W. van der Putten & K. Blom, 1999. mites australis wetland in Denmark. Aquatic Botany 64: Control of plant species richness and zonation of functional 381–398. groups along a freshwater flooding gradient. Oikos 86: 523– Middelkoop, H., 2000. Heavy-metal pollution of the river 534. Rhine and Meuse floodplains in the Netherlands. Nether- Leuven, R. S. E. W., A. J. M. Smits & P. H. Nienhuis, 2000. lands Journal of Geosciences – Geologie en Mijnbouw 79: From integrated approaches to sustainable river manage- 411–427. ment. In Smits, A. J. M., P. H. Nienhuis & R. S. E. W. Middleton, B. A., 1999. Wetland Restoration, Flood Pulsing Leuven (eds), New Approaches to Sustainable River and Disturbance Dynamics. John Wiley and Sons Inc., New Management. Backhuys Publishers, Leiden, 329–347. York. Leuven, R. S. E. W., S. Wijnhoven, L. Kooistra, R. J. W. de Mommer, L., O. Pedersen & E. J. W. Visser, 2004. Accli- Nooij & M. A. J. Huijbregts, 2005. Toxicological constraints mation of a terrestrial plant to submergence facilitates gas for rehabilitation of riverine habitats along lowland rivers. exchange under water. Plant Cell and Environment 27: Archiv fu¨r Hydrobiologie Supplement 155, Large Rivers 15: 1281–1287. 657–676. Moss, B., 1994. Brackish and fresh-water shallow lakes – Dif- Lockaby, B. G., A. L. Murphy & G. L. Somers, 1996. ferent systems or variations on the same theme. Hydrobio- Hydroperiod influences on nutrient dynamics in decompos- logia 276: 1–14. ing litter of a floodplain forest. Soil Science Society of Mostert, E., 1998. River basin management in the European America Journal 60: 1267–1272. Union. How it is done and how it should be done. European Loeb, R., & L. P. M. Lamers, 2003. The effects of river water Water Management 1: 26–35. quality on the development of wet floodplain vegetation Nienhuis, P.H., 2006. Water and values: ecological research as types in the Netherlands. Proceedings of the International the basis for water management and nature management. Ecoflood Conference ‘Towards natural flood reduction Hydrobiologia 565: 261–275. strategies’, Warsaw, Poland. http://levis.sggw.waw.pl/ Nienhuis, P. H., A. D. Buijse, R. S. E. W. Leuven, A. J. M. ecoflood/. Smits, R. J. W. de Nooij & E. M. Samborska, 2002. Eco- Lomans, B. P., A. J. P. Smolders, L. M. Intven, A. Pol, H. J. M. logical rehabilitation of the lowland basin of the river Rhine den op Camp & C. van der Drift, 1997. Formation of (NW Europe). Hydrobiologia 478: 53–72. dimethyl sulfide and methanethiol in anoxic freshwater Nienhuis P. H., R. S. E. W. Leuven, & A. M. J. Ragas (eds), sediments. Applied and Environmental Microbiology 63: 1998. New Concepts for Sustainable Management of River 4741–4747. Basins. Backhuys Publishers, Leiden. Lucassen, E. C. H. E. T., R. Bobbink, A. J. P. Smolders, P. Nordstrom, D. K., 1982. Aqueous pyrite oxidation and con- J. M. van der Ven, L. P. M. Lamers & J. G. M. Roelofs, sequent formation of secondary iron minerals in acid sul- 2003. Interactive effects of low pH and high ammonium phate weathering. Soil Science Society of America Journal levels responsible for the decline of Cirsium dissectum (L.) 10: 37–56. Hill. Plant Ecology 165: 45–52. Olde Venterink, H., M. J. Wassen, J. D. M. Belgers & J. T. A. Lucassen, E. C. H. E. T., A. J. P. Smolders, L. P. M. Lamers & Verhoeven, 2001. Control of environmental variables on J. G. M. Roelofs, 2005. Water table fluctuations and species density in fens and meadows: importance of direct groundwater supply are important in preventing phosphate- effects and effects through community biomass. Journal of eutrophication in sulphate-rich fens: consequences for Ecology 89: 1033–1040. wetland restoration. Plant and Soil 269: 109–115. Olde Venterink, H., M. J. Wassen, A. W. M. Verkroost & P. C. Lucassen, E. C. H. E. T., A. J. P. Smolders & J. G. M. Roelofs, de Ruiter, 2003a. Species richness-productivity patterns dif- 2000. Increased groundwater levels cause iron toxicity in fer between N-, P-, and K-limited wetlands. Ecology 84: Glyceria fluitans (L.). Aquatic Botany 66: 321–327. 2191–2199. 184

Olde Venterink, H., F. Wiegman, G. E. M. van der Lee & J. E. Roelofs, J. G. M., A. J. Kempers, A. L. F. M. Houdijk & J. Vermaat, 2003b. Role of active floodplains for nutrient Jansen, 1985. The effect of ar-borne ammonium-sulfate on retention in the river Rhine. Journal of Environmental Pinus nigra var maritima in the Netherlands. Plant and Soil Quality 32: 1430–1435. 84: 45–56. Olsen, S. R., C. W. Cole, R. Watanabe & L. A. Dean, 1954. Roozen, F. C. J. M., G. J. van Geest, B. W. Ibelings Estimation of Available Phosphorus in Soils by Extraction R. Rooijackers, M. Scheffer & A. D. Buijse, 2003. Lake age with Sodium Bicarbonate. US Dpt. of Agriculture circular and water level affect the turbidity of floodplain lakes along 939. the lower Rhine. Freshwater Biology 48: 519–531. Onaindia, M., B. G. Debikuna & I. Benito, 1996. Aquatic Rout, G. R., S. Samantaray & P. Das, 2001. Aluminium plants in relation to environmental factors in northern Spain. toxicity in plants: a review. Agronomie 21: 3–21. Journal of Environmental Management 47: 123–137. Schro¨der, T. J., T. Hiemstra, J. P. M. Vink & S. E. A. T. M. van Patrick, W. H., S. Gotoh & B. G. Williams, 1973. Strengite der Zee, 2005. Modeling of the solid-solution partitioning of dissolution in flooded soils and sediments. Science 179: 564– heavy metals and arsenic in embanked flood plain soils of the 565. rivers Rhine and Meuse. Environmental Science and Tech- Patrick, W. H. & R. A. Khalid, 1974. Phosphate release and nology 39: 7176–7184. sorption by soils and sediments – effect of aerobic and Schuurkes, J. A. A. R., A. J. Kempers & C. J. Kok, 1988. anaerobic conditions. Science 186: 53–55. Aspects of biochemical sulfur conversions in sediments of a Paul, E. A. & F. E. Clark, 1989. Soil Biology and Biochemistry. shallow soft-water lake. Journal of Freshwater Ecology 4: 2nd edn. Academic Press, San Diego, London. 369–381. Pauwels, H., W. Kloppmann, J. C. Foucher, A. Martelat & V. Simons, J. H. E. J., C. Bakker, M. H. I. Schropp, L. H. Jans, Fritsche, 1998. Field tracer test for denitrification in a pyrite- F. R. Kok & R. E. Grift, 2001. Man-made secondary bearing aquifer. Applied Geochemistry 13: 767–778. channels along the River Rhine (the Netherlands); results of Pedroli, B., G. de Blust, K. van Looy & S. van Rooij, 2001. post-project monitoring. Regulated Rivers: Research and Setting targets in strategies for river restoration. Landscape Management 17: 473–491. Ecology 17: 5–18. Sjodin, A. L., W. M. Lewis & J. F. Saunders, 1997. Denitrifi- Pezeshki, S. R., 2001. Wetland plant responses to soil flooding. cation as a component of the nitrogen budget for a large Environmental and Experimental Botany 46: 299–312. plains river. Biogeochemistry 39: 327–342. Pinay, G., V. J. Black, A. M. Planty-Tabacchi, B. Gumiero Smit, H., G. van der Velde, R. Smits & H. Coops, 1997. Eco- & H. Decamps, 2000. Geomorphic control of denitrifica- system responses in the Rhine-Meuse delta during two dec- tion in large river floodplain soils. Biogeochemistry 50: ades after enclosure and steps toward estuary restoration. 163–182. Estuaries 20: 504–520. Poiani, K. A. & B. L. Bedford, 1995. GIS based nonpoint- Smits, A. J. M., R. M. J. C. Kleukers, C. J. Kok & G. van der source pollution modeling – considerations for wetlands. Velde, 1990b. Alcohol dehydrogenase isozymes in the roots Journal of Soil and Water Conservation 50: 613–619. of some nymphaeid and isoetid macrophytes. Adaptations to Pollock, M. M., R. J. Naiman & T. A. Hanley, 1998. Plant hypoxic sediment conditions?. Aquatic Botany 38: 19–27. species richness in riparian wetlands – A test of biodiversity Smits, A. J. M., P. H. Nienhuis & R. S. E. W. Leuven (eds), theory. Ecology 79: 94–105. 2000. New Approaches to River Management. Bakhuys Ponnamperuma, F. N., 1984. Effects of flooding on soils. In Publishers, Leiden. Kozlowski, T. T. (ed.), Flooding and Plant Growth. Aca- Smits, A. J. M., P. H. Nienhuis & H. L. F. Saeijs, 2006. demic Press Inc., Orlando: 9–45. Changing estuaries, changing views. Hydrobiololgia 565: Postma, R., M. J. J. Kerkhofs, G. B. M. Pedroli & J. G. M. 339–355. Rademakers, 1996. Een stroom natuur. Natuurstreefbeelden Smits, A. J. M., P. H. van Avesaath & G. van der Velde, 1990a. voor Rijn en Maas. Watersysteemverkenningen 1996:1–102. Germination requirements and seed banks of some nym- RIZA nota 95.060 (in Dutch). phaeid macrophytes – Nymphaea alba L, Nuphar lutea Reddy, K. R., R. H. Kadlec, E. Flaig & P. M. Gale, 1999. (L) Sm and Nymphoides peltata (Gmel.) O. Kuntze. Fresh- Phosphorus retention in streams and wetlands: a review. water Biology 24: 315–326. Critical Reviews in Environmental Science and Technology Smits, A. J. M., G. W. H. Schmitz & G. van der Velde, 1992. 29: 83–146. Calcium-dependent lamina production of Nymphoides pel- Rengel, Z., 1996. (Tansley review) Uptake of aluminium by tata (Gmel.) O. Kuntze (Menyanthaceae) – implications for plant cells. New Phytologist 134: 389–406. distribution. Journal of Experimental Botany 43: 1273–1281. Richardson, C. J., 1985. Mechanisms controlling phosphorus Smits,A.J.M.,G.H.W.Schmitz,G.vanderVelde&L.A.C.J. retention capacity in fresh-water wetlands. Science 228: Voesenek, 1995. Influence of ethanol and ethylene on the 1424–1427. seed-germination of three nymphaeid water plants. Fresh- RIVM, 2004. Water quality Rhine and Meuse, 1980–2003 water Biology 34: 39–46. (figures from RIZA monitoring programme, various years), Smolders, A. J. P., C. den Hartog, C. B. L. van Gestel & J. G. http: //www.rivm.nl/milieuennatuurcompendium/. M. Roelofs, 1996. The effects of ammonium on growth, Roelofs, J. G. M., 1991. Inlet of alkaline river water into peaty accumulation of free amino acids and nutritional status of lowlands: effects on water quality and Stratiotes aloides L. young phosphorus deficient Stratiotes aloides plants. Aqua- stands. Aquatic Botany 39: 267–293. tic Botany 53: 85–96. 185

Smolders, A. J. P., L. P. M. Lamers, M. Moonen, K. Zwaga & ammonium toxicity in combination with low pH. New J. G. M. Roelofs, 2001. Controlling phosphate release from Phytologist 166: 551–564. phosphate-enriched sediments by adding various iron com- Van den Brink, F. W. B., J. P. H. M. de Leeuw, G. van der pounds. Biogeochemistry 54: 219–228. Velde & G. M. Verheggen, 1993. Impact of hydrology on the Smolders, A. J. P. & J. G. M. Roelofs, 1993. Sulphate mediated chemistry and phytoplankton development in floodplain iron limitation and eutrophication in aquatic systems. lakes along the lower Rhine and Meuse. Biogeochemistry 19: Aquatic Botany 46: 247–253. 103–128. Smolders, A. J. P. & J. G. M. Roelofs, 1995. Internal eutro- Van den Brink, F. W. B., M. M. J. Maenen, G. van der phication, iron limitation and sulphide accumulation due to Velde & A. bij de Vaate, 1991. The (semi-)aquatic vege- the inlet of river Rhine water in peaty shallow waters in the tation of still waters within the floodplains of the rivers Netherlands. Archiv fu¨r Hydrobiologie 133: 349–365. Rhine and Meuse in The Netherlands: historical changes Smolders, A. J. P. & J. G. M. Roelofs, 1996. The roles of and the role of inundation. Verhandlungen der Interna- internal iron hydroxide precipitation, sulphide toxicity and tionale Vereinigung fu¨r theoretische und angewandte oxidizing ability in the survival of Stratiotes aloides roots at Limnologie 24: 2693–2699. different iron concentrations in sediment pore water. New Van den Brink, F. W. B. & G. van der Velde, 1993. Growth and Phytologist 133: 253–260. morphology of four fresh-water macrophytes under the im- Smolders, A., J. G. M. Roelofs & G. van der Velde, 1994. Iron pact of raised salinity level of the Lower Rhine. Aquatic deficiency in Nymphoides peltata owing to the exhaustion of Botany 45: 285–297. dissolved iron in anaerobic sediments. Aquatic Botany 47: Van den Brink, F. W. B. & G. van der Velde, 1994. Impact of 349–353. hydrology and water chemistry on floodplain lake commu- Snowden, R. E. D. & B. D. Wheeler, 1995. Chemical changes in nities along the lower Rhine and Meuse. Water Science and selected wetland plant species with increasing Fe supply, Technology 29: 57–60. with specific reference to root precipitates and Fe tolerance. Van den Brink, F. W. B., G. van der Velde, W. W. Bosman & New Phytologist 131: 503–520. H. Coops, 1995. Effects of substrate parameters on growth- Sparks, R. E., 1995. Need for ecosystem management of large responses of eight helophyte species in relation to flooding. rivers and their floodplains. BioScience 45: 168–182. Aquatic Botany 50: 79–97. Spink, A., R. E. Sparks, M. van Oorschot & J. T. A. Van den Brink, F. W. B., M. M. van Katwijk & G. van der Verhoeven, 1998. Nutrient dynamics of large river flood- Velde, 1994. Impact of hydrology on phytoplankton and plains. Regulated Rivers: Research and Management 14: zooplankton community composition in floodplain lakes 203–216. along the lower Rhine and Meuse. Journal of Plankton Steiger, J. & A. M. Gurnell, 2003. Spatial hydrogeomorpho- Research 16: 351–373. logical influences on sediment and nutrient deposition in Van der Hoek, D., A. J. E. M. van Mierlo & J. M. Van riparian zones: observations from the Garonne River, Groenendael, 2004. Nutrient limitation and nutrient-driven France. Geomorphology 49: 1–23. shifts in plant species composition in a species-rich fen Stumm, W. & J. J. Morgan, 1981. Aquatic Chemistry, an meadow. Journal of Vegetation Science 15: 389–396. Introduction Emphasizing Chemical Equilibria in Natural Van der Lee, G. E. M., H. O. Venterink & N. E. M. Asselman, Waters (2nd ed.). John Wiley and Sons Inc., New York. 2004. Nutrient retention in floodplains of the Rhine dis- Taylor, D. R., L. W. Aarssen & C. Loehle, 1990. On the rela- tributaries in The Netherlands. River Research and Appli- tionship between R/K selection and environmental carrying- cations 20: 315–325. capacity – A new habitat templet for plant life-history Van der Valk, A. G., 1981. Succession in wetlands – a Glea- strategies. Oikos 58: 239–250. sonian approach. Ecology 62: 688–696. Tessier, A. & D. R. Turner (eds), 1995. Metal Speciation and Van der Welle, M. E. W., M. L. C. Cuppens, L. P. M. Lamers Bioavailability in Aquatic Systems. John Wiley & Sons Inc., & J. G. M. Roelofs, (in press). Detoxifying toxicants: inter- Chichester. actions between iron and sulfide toxicity in freshwater wet- Tilman, D., 1994. Competition and biodiversity in spatially lands. Environmental Toxicology and Chemistry. structured habitats. Ecology 75: 2–16. Van Dijk, J., M. Stroetenga, L. Bos, P. M. van Bodegom, H. A. Tockner, K., F. Schiemer & J. V. Ward, 1998. Conservation by Verhoef & R. Aerts, 2004. Restoring natural seepage con- restoration: the management concept for a river-floodplain ditions on former agricultural grasslands does not lead to system on the Danube River in Austria. Aquatic Conserva- reduction of organic matter decomposition and soil nutrient tion – Marine and Freshwater Ecosystems 8: 71–86. dynamics. Biogeochemistry 71: 317–337. Tre` molieres, M., R. Carbiener, A. Ortscheit & J. P. Klein, 1994. Van Eck, W. H. J. M., J. P. M. Lenssen, H. M. van de Steeg, Changes in aquatic vegetation in Rhine floodplain streams in C. W. P. M. Blom & H. de Kroon, 2006. Seasonal dependent Alsace in relation to disturbance. Journal of Vegetation effects of flooding on plant species survival and zonation: a Science 5: 169–178. comparative study of 10 terrestrial grassland species. Van Breemen, N., J. Mulder & C. T. Driscoll, 1983. Acidifi- Hydrobiologia 565: 59–69. cation and alkalinization of soils. Plant and Soil 75: 283–308. Van Geest, G. J., H. Coops, R. M. M. Roijackers, A. D. Buijse Van den Berg, L. J. L., E. Dorland, P. Vergeer, M. A. C. Hart, & M. M. Scheffer, 2005. Succession of aquatic vegetation R. Bobbink & J. G. M. Roelofs, 2005. Decline of acid-sen- driven by reduced water-level fluctuations in floodplain sitive plant species in heathland can be attributed to lakes. Journal of Applied Ecology 42: 251–260. 186

Van Geest, G. J., F. C. J. M. Roozen, H. Coops, R. M. M. shifts in plant species composition in a species-rich fen Roijackers, A. D. Buijse, E. T. H. M. Peeters & M. Scheffer, meadow. Journal of Vegetation Science 15: 389–396. 2003. Vegetation abundance in lowland flood plan lakes Ward, J. V., K. Tockner, D. B. Arscott & C. Claret, 2002. determined by surface area, age and connectivity. Freshwa- Riverine landscape diversity. Freshwater Biology 47: 517– ter Biology 48: 440–454. 539. Van Oorschot, M., C. Hayes & I. van Strien, 1998. The influ- Wassen, M. J., H. G. M. O. Veterink & E. O. A. M. de Swart, ence of soil desiccation on plant production, nutrient uptake 1995. Nutrient concentrations in mire vegetation as a mea- and plant nutrient availability in two French floodplain sure of nutrient limitation in mire ecosystems. Journal of grasslands. Regulated Rivers: Research and Management Vegetation Science 6: 5–16. 14: 313–327. Wassen, M. J., R. E. van der Vliet & J. T. A. Verhoeven, 1998. Van Stokkom, H. T. C., A. J. M. Smits & R. S. E. W. Leuven, Nutrient limitation in the Biebrza fens and floodplain (Po- 2005. Flood defence in the Netherlands: a new era, a new land). Acta Botanica Neerlandica 47: 241–253. approach. Water International 30: 76–87. Wassen, M. J., H. Olde Venterink, E. D. Lapshina & Van Wijck, C., C. J. de Groot & P. Grillas, 1992. The effect of F. Tanneberger, 2005. Endangered plants persist under anaerobic sediment on the growth of Potamogeton pectinatus phosphorus limitation. Nature 437: 547–550. l – The role of organic-matter, sulfide and ferrous iron. Wassen, M. J., W. H. M. Peeters & H. O. Venterink, 2002. Aquatic Botany 44: 31–49. Patterns in vegetation, hydrology, and nutrient availability Verhoeven, J. T. A., D. F. Whigham, R. van Logtestijn & in an undisturbed river floodplain in Poland. Plant Ecology J. O’Neill, 2001. A comparative study of nitrogen and 165: 27–43. phosphorus cycling in tidal and non-tidal riverine wetlands. Wassen, M. J., W. H. M. Peeters & H. Olde Venterink, 2003. Wetlands 21: 210–222. Patterns in vegetation, hydrology, and nutrient availability Vervuren, P. J. A., C. W. P. M. Blom & H. de Kroon, 2003. in an undisturbed river floodplain in Poland. Plant Ecology Extreme flooding events on the Rhine and the survival and 165: 27–43. distribution of riparian plant species. Journal of Ecology 91: Wheeler, B. D. & M. C. F. Proctor, 2000. Ecological gradients, 135–146. subdivisions and terminology of north–west European mires. Vervuren, P. J. A., S. M. J. H. Beurskens & C. W. P. M. Blom, Journal of Ecology 88: 187–203.

1999. Light acclimation, CO2 response and long-term Wijnhoven, S., G. van der Velde, R.S.E.W. Leuven & A. J. M. capacity of underwater photosynthesis in three terrestrial Smits, 2006. Modelling recolonisation of heterogeneous river plant species. Plant Cell and Environment 22: 959–968. floodplains by small mammals. Hydrobiologia 565: 135– Vink, J. P. M., 2002. Measurement of heavy metal speciation 152. over redox gradients in natural water-sediment interfaces Winter, L. T., I. D. L. Foster, S. M. Charlesworth & J. A. Lees, and implications for uptake by benthic organisms. Envi- 2001. Floodplain lakes as sinks for sediment-associated ronmental Science and Technology 36: 5130–5138. contaminants - A new source of proxy hydrological data? Visser, E. J. W., L. A. C. J. Voesenek, B. B. Vartapetian & Science of the Total Environment 266: 187–194. M. B. Jackson, 2003. Flooding and plant growth. Annals of Young, E. O. & D. S. Ross, 2001. Phosphate release from Botany 91: 107–109. seasonally flooded soils: a laboratory microcosm study. Van der Hoek, D., A. J. E. M. van Mierlo & J. M. van Journal of Environmental Quality 30: 91–101. Groenendael, 2004. Nutrient limitation and nutrient-driven