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Linköping Studies in Science and Technology Dissertation No 1041

Nitrogen removal in treatment – Factors influencing spatial and temporal variations

Sofia Bastviken

Department of Biology, IFM Linköping University, SE-581 83 Linköping, Sweden

Linköping 2006

Nitrogen removal in treatment wetlands – Factors influencing spatial and temporal variations

ISBN 91-85523-12-7 ISSN 0345-7524 © Sofia Bastviken

Front cover photo by Jonas Andersson, WRS i Uppsala

Printed by LiU Tryck, Linköping, 2006 ABSTRACT

Decreasing the nitrogen transport from land to surrounding seas is a major task throughout the world to limit of the coastal areas. Several approaches are currently used, including the establishment of wetlands, to decrease the transport of nitrogen. Wetlands represent where the nitrogen removal from water can be efficient given that they are appropriately designed. The aim of this thesis was to investigate and quantify the effect of critical factors that regulate the nitrogen removal in wetlands, and to develop better guidelines for design. Studies were performed at different scales, from microcosms to full scale wetlands, and methods included modelling, mass balance calculations and process studies. A first-order rate model was used to simulate the nitrogen transformations in two large wetlands treating wastewater containing both and nitrogen. It was found that the dynamics of the main nitrogen transformation processes could not be satisfactorily described using this approach. Large wetlands containing vegetation are complex ecosystems, and the process rates vary in both time and space. The great diversity of microenvironments favours different nitrogen processes, and large differences in potential and denitrification rates were found between different surface structures within a wetland. The results from microcosms measurements showed that the highest potential for nitrification was on surfaces in the water column, while the denitrification capacity was highest in the sediment. For the sediment denitrification capacity, the plant community composition was shown to be of major importance primarily by supplying litter serving as a carbon and energy source, and/or attachment surfaces, for denitrifying . Denitrification rates may be affected more than three fold by different types of litter and in the sediments. Intact sediment cores from stands of the emergent plants Glyceria maxima and Typha latifolia had higher denitrification potential than sediment cores from stands of the submersed plant Potamogeton pectinatus. However, the quality of the organic material for the was highest in G. maxima and P. pectinatus stands. All sediment cores from the wetland were limited by carbon, and the lower denitrification capacity of the submersed plant, P. pectinatus, was likely due to lower amounts of . However, in another wetland, intact cores from stands of the submersed plant Elodea canadensis had a higher denitrification capacity than the cores from stands of T. latifolia and Phragmites australis. This was possibly due to a larger biomass, and better quality, of the organic matter from that submersed specie, or to epiphytic biofilms on the living plants. Those microcosms studies showed that both the quality of the organic matter as a substrate for the microbial communities, and the amount of organic material produced were important for the denitrification capacity. In pilot-scale wetlands, the composition of the plant community was also a more important factor for high nitrate removal than the differences in hydraulic loads (equivalent of 1 or 3 d retention time), despite the cold climate. The greatest removal was found in wetlands with emergent vegetation dominated by P. australis and G. maxima, rather than in wetlands with submersed vegetation. In brief, the results presented in this thesis emphasize the importance of dense emergent vegetation for high annual nitrate removal in treatment wetlands.

LIST OF PAPERS

The following papers are included in the thesis. They are referred to in the text by their Roman numerals.

I Kallner S. and Wittgren H. B. 2001. Modelling nitrogen transformations in free water surface wastewater treatment wetlands in Sweden. Wat. Sci. Tech. 44 (11-12): 237-244.

II Kallner Bastviken S., Eriksson P. G., Martins I., Neto J. M., Leonardson L. and Tonderski K. 2003. Potential nitrification and denitrification on different surfaces in a constructed treatment wetland. J. Environ. Qual. 32:2414-2420 and J. Environ. Qual. 33:411 (2004) (errata).

III Kallner Bastviken S., Eriksson P. G., Premrov A. and Tonderski K. S. 2005. Potential denitrification in wetland sediments with different plant species detritus. Ecol. Eng. 25:183-190.

IV Kallner Bastviken S., Eriksson P. G., Ekstöm A. and Tonderski K. S. Seasonal denitrification potential in wetland sediments with organic matter from different plant species. Submitted to Water, Air and Pollution

V Kallner Bastviken S., Weisner S. E. B., Thiere G., Svensson J. M., Ehde P. M. and Tonderski K. S. Effects of vegetation and hydraulic load on seasonal nitrate removal in treatment wetlands. Manuscript

Papers I, II and III are reprinted with permission from the publisher.

CONTENTS

Introduction...... 1 The nitrogen transformation processes in wetlands...... 2 The role of vegetation for nitrification and denitrification ...... 6 The role of nutrient and hydraulic load for nitrification and denitrification ...... 7 Constructed wetlands in the landscape...... 9

Rationale of this thesis...... 11

Methods...... 12 Microcosm process studies...... 12 Model approach for describing nitrogen removal...... 14

Spatial variability of the nitrogen removal processes ...... 17

The effect of organic matter on denitrification...... 19

Critical hydraulic load for nitrate removal...... 23

Comparing wetland performance ...... 24

Conclusions and future studies...... 27 Conclusions ...... 27 Future studies...... 28

Acknowledgement...... 28

References...... 29

Tack...... 35

INTRODUCTION

Eutrophication is a major problem in many coastal areas and lakes of the world including the coastal areas surrounding Sweden. Nitrogen and phosphorus limit the primary production of different parts of the seas and it is likely that an increase of nitrogen and phosphorus transports from the drainage areas has enhanced the eutrophication process (Boesch et al. 2006; Granéli et al. 1990; Håkansson 2002). The direct effect of nitrogen and phosphorus loads on the seas is hard to measure as the primary production of the seas varies with weather conditions and water circulations. However, to improve the trophic status of the coastal areas surrounding Sweden in a long-term perspective, measures to limit the load of nitrogen and phosphorus are critical.

At the Swedish west coast, Kattegat and Skagerrak, a decrease of particularly the nitrogen transport is thought to be critical to prevent the eutrophication process (Boesch et al. 2006; Håkansson 2002; Nielsen et al. 2002). Håkansson (2002) estimated that the load of nitrogen from point sources and rivers to the Swedish west coast, Kattegat and Skagerrak, has increased by 40 % from the 1970s. Agriculture has been estimated to be the largest single source of nitrogen to the sea (SEPA 1997). During the last two decades, measures have been taken to reduce the nitrogen transport originating from agricultural activities. Firstly, best management practises are increasingly used to reduce the input of nitrogen to the agricultural fields and the nitrogen leakage from the fields. Secondly, different water treatment approaches, such as wetlands and , are used to decrease the transport of nitrogen from land to the sea (Fleischer et al. 1994; Svensson et al. 2004).

Drainage of large areas from the 19th century has caused decreased residence times of the water from land to sea. The sometimes very short residence times in catchment areas limit the contact between the nutrients in the water and the environments that cycle these nutrients, which has led to increased transport of nutrients to the sea. These large drainage projects led to less lakes and wetlands in the catchment areas. Lakes and ponds are able to store water volumes, and consequently the high floods and the low flows become less pronounced downstreams. In this way, lakes and wetlands increase the water residence times in catchment areas during high floods and provide environments for nutrient cycling and removal from the water phase. The year-round presence of water and high biomass production make wetlands among the most biologically active ecosystems on earth. Due to the high biological activity, wetlands transform many compounds at a high rate and have increasingly been

1 used for wastewater treatment (Kadlec and Knight 1996). In the recent decade, research has focused on how wetlands can be used to remove nitrogen, both in agricultural drainage water (Braskerud 2002; Phipps and Crumpton 1994; Raisin and Mitchell 1995; Spieles and Mitsch 2000) and municipal wastewater (Andersson et al. 2005; Kadlec and Knight 1996). Wetlands have been constructed with different design and the two main groups are free water surface wetlands, where the water is passing over the sediment surface, and sub surface wetlands, where the water is transported within the soil medium of a planted bed. Free water surface wetlands are the most commonly constructed wetlands in Sweden and are also in focus in this thesis.

Free water surface wetlands have been constructed using different designs and with different environmental conditions, i.e. with different water quality, water flow and load, surface area, depth and containing dense emergent or submersed vegetation or no macrophytes at all. Physical, chemical and biological factors influence the nitrogen removal both spatially and seasonally and the nitrogen removal efficiency varies among wetlands and throughout the year. Today, we can not satisfactorily explain the observed differences in nitrogen removal between wetlands and more knowledge is needed to develop better guidelines to improve nitrogen removal in constructed wetlands. The aim of this thesis was to investigate and quantify critical factors that regulate the nitrogen removal in wetlands, and to develop better guidelines for wetland design. Studies were performed at different scales, and methods included modelling, mass balance calculations and process studies.

The nitrogen transformation processes in wetlands

In the biosphere, nitrogen is continuously transformed between organic, soluble inorganic and gaseous nitrogen forms (Fig. 1). The is very complex, and it is hard to control even the most basic transformations within a wetland. How much nitrogen that is removed, i.e. transformed or removed from the water phase, in a wetland and which nitrogen processes or nitrogen fluxes that are the most important depend on water chemistry and other wetland conditions, such as climate, vegetation, water depth and water flow. Ammonification is the microbial mineralisation of organic nitrogen to ammonium. This process can be caused by heterotrophic bacteria and fungi. Ammonium is then transformed to nitrate by the bacterial process nitrification and nitrate or is finally reduced to gaseous end products, nitrous gas and dinitrogen gas, through the bacterial process denitrification. Nitrate can

2 also be transformed to ammonium during low conditions (Prescott et al. 1990; Vymazal 2001). is a bacterial process, which transfers dinitrogen gas to ammonium. However, nitrogen fixation is generally not significant in nitrogen rich waters such as constructed wetlands for water treatment (Kadlec and Knight 1996).

Nitrogen assimilation is the transformation of inorganic nitrogen to organic nitrogen in cells and tissue. Plant assimilation only accounts for a small percent of the total nitrogen removal when the nitrogen load is high (Tanner et al. 1995), which is the case in most free water surface treatment wetlands (Kadlec and Knight 1996). Further, plant uptake does not play an important role in the annual nitrogen removal in the long run as the nitrogen assimilated by vegetation usually is released during decomposition of litter (Johnston 1991). However, plant uptake can contribute to the seasonal dynamic of nitrogen removal and account for a significant part of the wetland nitrogen removal during the growth period with rapid nitrogen uptake. In temperate climates and in wetlands receiving riverine runoff, this often coincides with periods of low flow and nitrogen load, thus making the plant uptake more important temporary.

Ammonia volatilisation is a physicochemical process where ammonia in the ammonium-ammonia equilibrium is transported to the gas phase. Ammonia volatilisation can be important in wetlands with high temperature and pH, although volatilisation losses of ammonia usually are small if pH is below 8 (Reddy and Patrick 1984). Other nitrogen fluxes that can occur in a wetland are sedimentation and resuspension of the various nitrogen forms. Burial of organic nitrogen in the sediment will make the nitrogen less available to living plants and organisms, while release of nitrogen from biomass during decomposition will make the nutrients available again. Ammonium can also readily adsorb to sediment particles and litter, ammonium adsorption, as ammonium is a positively charged ion. The adsorbed ammonium concentration is in equilibrium with the ammonium concentration in the water, and can be released if a change in the water chemistry or hydrology occurs.

3 Nitrogen fixation N2

NH3 Air

Ammonia volatilisation Denitrification Ammonification (Anoxic) Nitrification ON Assimilation AN NN

Assimilation

Sedimentation Diffusion and adsorption Diffusion Water Nitrification Ammonification (Oxic) Sediment

ON Assimilation AN NN

Assimilation

Figure 1. A simplified picture of the nitrogen processes and the flows of different nitrogen forms in a wetland (modified from Wittgren H. B., unpublished). ON is organic nitrogen, AN ammonium nitrogen and NN nitrate nitrogen.

Nitrification followed by denitrification is usually the most important processes for the permanent nitrogen removal in wetlands receiving ammonium rich waters (Johnston 1991). In wetlands receiving nitrate as the dominant nitrogen form, such as in agricultural drainage water, the most important process is usually denitrification (Johnston 1991; Trepel and Palmeri 2002; Vymazal 2001). In this thesis, both nitrification and denitrification have been studied.

Nitrification and denitrification is influenced by factors at different spatial scales. In Figure 2, factors affecting denitrification have been described at different spatial scales; the process scale, the wetland scale and the landscape scale. At the process scale, the process rates performed by the bacteria are directly regulated by factors like temperature and redox conditions. These process scale factors can be affected by factors on the wetland scale, such as water flow, nutrient load and different plant communities. Finally, the wetland scale factors are affected by landscape

4 factors such as climate and land use, e.g. higher nitrogen load the more agricultural areas there is in the upstream catchment area.

Figure 2. Factors affecting the microbiological process of denitrification on different spatial scales (Trepel 2002).

On the process scale (Fig. 2), the rates of nitrification and denitrification are affected by several factors. The require , an inorganic carbon source, as well as ammonium, and are favoured by high soil temperature (optimum 30-40 oC) (Reddy and Patrick 1984) and high pH (7.5-8.0) (Prosser 1989). In the denitrification process, commonly organic material serves as an electron donor while nitrate is used as an electron acceptor. Denitrification rates will be favoured by anoxic conditions, high nitrate availability and bioavailable organic carbon. The bacteria are also favoured by high temperature (optimum 60-75 oC) and pH 6-8.5 (Knowles 1982; Reddy and Patrick 1984). Most denitrifying bacteria are , but some are autotrophs and uses H2, CO2 and reduced sulphur compounds (Knowles 1982).

5 The role of vegetation for nitrification and denitrification

On the wetland scale, plants can play an important role for the total removal of nitrogen in wetlands. Wetlands containing plants have been shown to remove larger quantities of nitrate than unplanted wetlands (Bachand and Horne 2000; Lin et al. 2002; Tanner et al. 1995; Zhu and Sikora 1995). The vegetation can affect the nitrifying and denitrifying bacteria in the wetland in many ways. In wetlands, the organic carbon is supplied mainly by the vegetation. This organic matter is used directly as a carbon and energy source for heterotrophic bacteria such as the denitrifying bacteria. The effect of vegetation on denitrification is dependent on the quality and the amount of the plant litter (Hume et al. 2002). The decomposition rate, or the availability of the organic carbon as an energy source for the heterotrophic bacteria, is linked to the initial composition of nitrogen and lignin in the plant (Hume et al. 2002). Submersed and free-floating plants generally have a lower initial lignin to N ratio than emergent plant species (Godshalk and Wetzel 1978; Hume et al. 2002), and consequently, the organic matter of a submersed plant is usually more available for decomposition, and hence for denitrifying bacteria, than the organic matter of an emergent plant. However, the primary production is higher for the emergent plants (Westlake et al. 1998) and hence, the organic matter lasts longer when emergent plants are decomposed than for the submersed macrophytes. In addition, the seasonal pattern of litter accumulation may significantly affect the availability of organic matter, and the redox conditions in the sediments and water. Studies have shown that dry dead leaves remain on the stems till late autumn and only then fall off in a littoral stand of P. australis (Dykyjová and Kvet 1978). The dry dead plant parts become available to the bacteria first when it reaches the water. In a study of G. maxima, the amount of loose litter, not attached to dead shoots was highest in November to March, and then rapidly decreased to a minimum in May (Westlake 1966). However, questions remain about what plant litter that is most favourable for denitrification in wetlands.

Organic material can also have an indirect effect on the nitrifying and denitrifying bacteria. The supply of litter on top of the sediment may limit the oxygen diffusion to lower sediment layers, which causes anaerobic sediments that is favourable for the denitrification process, but not for the nitrifying bacteria. Also, the supply of organic carbon raises the total heterotrophic activity, which consumes oxygen and thus indirectly favours denitrification by lowering oxygen concentrations in the sediment (Nielsen et al. 1990a).

6 Plant tissue provides a large amount of surface areas for microbial growth. Attached bacteria are more active and more abundant than free- living bacteria in aquatic systems (Hamilton 1987). This attachment of microorganisms to submersed solid surfaces, such as sediment and aquatic macrophytes, leads to the formation of biofilms. Different surfaces can affect the presence of nitrifiers and denitrifiers by being more or less suitable as attachment surfaces. Surfaces in the water phase, to which bacteria can attach and where oxygen and ammonium concentrations are high would favour nitrification. However, it is not yet known if some surface structures in wetlands are more important than others for the different nitrogen transformations. Eriksson and Andersson (1999) showed that litter of different emergent plants was of varying quality as habitats for nitrifying bacteria, with higher ammonium and nitrite oxidation in litter of Scirpus sylvaticus and Carex rostrata than in Equisetum fluviatile and Typha latifolia. Commonly, the sediment has been considered to be the site where most of the denitrification occurs in wetlands e.g. Seitzinger (1988), but later studies have shown that denitrification in periphytic communities on submerged plants can make a significant contribution to overall denitrification in a wetland (Eriksson and Weisner 1997; Toet 2003).

Living plants can also increase the oxygen concentrations in the water and thus create conditions that do not favour denitrification (Nielsen et al. 1990b; Toet et al. 2005), but are favourable for nitrification. Submersed plants can release oxygen to the water through their photosynthesis, while emergent plants can transport oxygen in aerenchymatous stem tissue to the rhizomes and roots, that may eventually diffuse to the sediment and water (Reddy et al. 1989; Vretare 2001). Hence, living and dead vegetation can influence the presence and activity of nitrifying and denitrifying bacteria in several different ways on the process scale. However, what type of vegetation within a wetland that is most favourable and provides the highest potential for nitrification and denitrification is not fully understood.

The role of nutrient and hydraulic load for nitrification and denitrification

The water flow regime of constructed treatment wetlands is mainly of two types. Either the wetland receives a relatively constant water flow, e.g. wetlands receiving municipal wastewater, or a water flow that is determined by the runoff in the catchment area, e.g. wetlands receiving drainage or storm water. At the wetland scale, the nitrogen removal in

7 treatment wetlands is usually assumed to be regulated by the supply of nitrogen, i.e. the dominant nitrogen fraction, the concentration, and the hydraulic load, and by the water temperature. A lower nitrogen removal has been observed in wetlands receiving large concentrations of ammonium nitrogen, i.e. municipal wastewater, than in those receiving nitrogen in the form of nitrate (Andersson et al. 2005). The reason is that in the former both nitrification and denitrification has to occur, and the conditions in the wetland have to be suitable for both processes. In wetlands receiving runoff water from agriculturally dominated catchment areas, the dominating nitrogen fraction is nitrate (Ekologgruppen 2001) and the nitrogen removal is mainly regulated by the rate of denitrification.

The hydrology of a wetland that is situated in a stream is strongly dependent on the runoff in the catchment area. The hydraulic load, the water volume entering the wetland divided by the wetland surface area, and the water velocity can have extreme variations and occasionally be very high. A high hydraulic load increases the amount of ammonium and nitrate passing the wetland and thus the availability of these nutrients for the bacteria. The nitrate removal rate of wetlands receiving runoff water from agricultural catchment areas is generally thought to be regulated by the nitrate load. The nitrate load is commonly correlated with the hydraulic load, and mass balance studies (inlet and outlet measurements) have shown that higher hydraulic loads result in higher area-specific nitrate removal (kg per hectare and year) (Bachand and Horne 1999; Kadlec 2005). However, Arheimer and Wittgren (2002) used data sets from 8 different wetlands to model nitrate removal, and found that area- specific nitrate removal was almost negligible at very high hydraulic loads. Other studies have shown that at occasional high hydraulic loads there can be larger outflows of nitrate than inflows (Raisin and Mitchell 1995; Spieles and Mitsch 2000). However, at present it is not known at what flow velocities such negative effects will occur and override the positive effects of an increased supply of nitrate.

The negative effects of high hydraulic loads on the nitrogen removal could be caused by several factors. High water flows and water velocities increases the turbulence and exchange of water within the wetland, which favours the supply of ammonium and nitrate to the surfaces of the wetland through turbulent diffusion. On the other hand, this also increases the oxygen concentrations of the water, which would favour nitrification, but limit the denitrification. In flowing water, oxygen may penetrate through the entire biofilm community and inhibit the development of anoxic microzones (Eriksson 2000). Christensen et al. (1989), showed that denitrification in a stream was restricted to a depth of 1-5 mm in the

8 biofilm of the sediment. In the upper level in the sediment, the denitrification was restricted by high oxygen concentrations, and at the lower levels it was restricted by nitrate availability with low nitrate concentrations in the deeper sediment. Very high water flows might also resuspend the organic material in the upper sediment, and thus decrease the amount of energy and carbon available for the denitrifying bacteria. High water velocities may also restrict the biofilm development and maturation due to physical disturbances (Claret and Fontvieille 1997). For example, Eriksson and Weisner (1996) showed that the epiphytic abundance on submersed plants and the denitrifying capacity was higher at sites that were less exposed to wave turbulence or water currents, than at more exposed sites.

The hydraulics, the internal flow patterns, can be of importance for how efficiently the wetland area is used for nitrogen transformations (Persson 1999). Different flow patterns can cause shortcuts in the system and result in much shorter nitrogen residence times, higher water velocities and less efficient wetland area. This affects the contact between the nutrients in the water and the bacteria and consequently the nitrate removal. The hydraulics can depend on the morphology (Persson 1999) and vegetation (Braskerud 2001) of the wetland.

Negative high flow effects on nitrogen removal have been shown to have a large impact on the annual nitrogen removal of wetland basins (Spieles and Mitsch 2000). Hence, to achieve high nitrogen removal on an annual basis, high flow effects that limit the nitrogen removal in a wetland should be avoided. The magnitude of these high flow effects are probably dependent on the conditions in the wetlands, e.g. vegetation, depth, hydraulics etc., and it is of great importance to identify the critical hydraulic loads at which such nitrogen release occur in different kinds of wetlands.

Constructed wetlands in the landscape

At the landscape scale, wetlands are affected by climate, nitrate concentrations and water flow. Under boreal temperate conditions, the hydraulic load and thus the nitrate transport is usually high during the colder period of the year, when the biological activity is low, and low during the warmer period when the biological activity is high (Fig. 3).

9 Q Temp. Conc.

60000 25

50000 C) o

) 20 -1 40000 d 3 30000 15 ), Temp. ( -1 20000 10 10000 Water flow (m 5 0 Conc. (mg L -10000 0 jul-95 jul-99 jan-97 jun-97 jan-98 jun-98 feb-99 okt-95 okt-96 apr-98 apr-99 nov-97 nov-98 sep-94 dec-94 maj-95 dec-95 maj-96 aug-96 aug-97 sep-98 mar-97 mar-95 mar-96

Figure 3. Seasonal variations of water flow (Q), nitrogen concentration and temperature in the outlet of the wetland basin Råbytorp, southern Sweden, in 1994 to 1999.

To reduce the nitrogen transport to the coastal areas, an important task is thus to achieve high nitrogen removal during the colder period. The activity of the nitrogen transforming processes is slower in cold temperatures, and thus the hydraulic load must be lower during winter to achieve the same percentage nitrogen removal (%) as in summer. Nevertheless, studies have shown that on an annual basis the area-specific nitrate removal, expressed as kg ha-1 y-1, is higher at higher loads of nitrate (in kg ha-1 y-1), also in colder climate (Fleischer et al. 1994). However, the relative nitrogen removal, expressed as percentage of the inflow of nitrogen, can be very low at high nitrogen loads. To remove large percentages of the nitrate load from the agricultural runoff water, the residence times of the runoff from land to sea need to be increased. To achieve longer residence times and efficient removal, wetlands should be constructed with a sufficiently large area to remove the amount or percentage of nitrate that is of interest. The relation between hydraulic load and wetland area is usually decided through estimating the size of the catchment area, and predicting the hydraulic load from that catchment size. Of great importance to achieve an efficient result is the position of the wetland in the landscape. This includes land use, geology and the distance from the recipient. The nitrogen processes are more efficient at higher nitrogen concentrations in the water. Large agricultural fields, e.g. sandy planted with potatoes, can loose high concentrations of nitrogen to the water (Hoffmann 1999) and could thus be a good choice to include in the catchment for a planned efficient wetland. Arheimer (1998) showed that nitrogen that was discharged to the surface waters of

10 a catchment close to the recipient of interest, e.g. the coastal areas, would have much greater impact than the nitrogen that was discharged far upstream from the recipient. The reason is that nitrogen that is transported a long distance by rivers and streams will have a long residence time before it reaches the recipient and thus there is a long time for the nitrogen transformation processes to act. On the opposite, nitrogen in the runoff discharged close to the recipient has a very short residence time and a larger proportion will affect the recipient. This leads to the conclusion that wetlands that are established close to the recipient are more efficient than those established far upstream.

RATIONALE OF THIS THESIS

The aim of the study in Paper I was to investigate if a fairly simple first- order area-based kinetic model could describe the fate of nitrogen satisfactorily. Another intention was to discuss the fate of nitrogen in two wetlands, receiving different forms of nitrogen, and analyse what environmental factors that could be most critical for the nitrogen removal processes in the respective wetland. The model in Paper I was site- specific and insufficient to describe the fate of the different nitrogen forms in the wetlands. This indicated that the spatial heterogeneity within the wetlands, i.e. differences within and between the wetland basins, were important. Using the first-order area-based equation, the nitrogen transformations are assumed to occur only at the sediment surface in a wetland. However, substantial nitrification and denitrification have also been documented in biofilms of surfaces in the water column, such as submersed plants (Eriksson and Weisner 1997; Eriksson and Weisner 1999). Hence, in Paper II, the spatial distribution of the nitrogen processes within a wetland was investigated. This was studied using a microcosm approach.

The aim of Paper II was to investigate the capacity of nitrification and denitrification on different surfaces in a wetland and also to discuss the importance of these surfaces for the overall nitrogen removal in the wetland. The results showed great spatial variability with highest nitrification on twigs laying in the water and highest denitrification in the sediments. This shows that the first-order area based model was a good choice for describing denitrification in the wetlands as the sediment had the greatest potential for denitrification and very low denitrification potential was found at surfaces in the water column. However, different characteristics of the sediments might cause large variations in denitrification capacity within and between wetlands. The high

11 denitrification in the sediment compared to the other surfaces indicated that organic matter availability may regulate the distribution of the denitrifying bacteria when nitrate concentrations are high.

This emphasised the importance of organic matter in the sediment. Plants supply the denitrifying bacteria with organic material when decomposing in the sediments and constructed wetlands contain different types and species of vegetation. Consequently, the aim of Paper III and IV was to investigate if the denitrification capacity differed between wetland sediments formed in stands of different plant species and if these denitrification rates varied with season. The study showed a large variation in denitrification capacity between the sediments from different plant species and between seasons. However, these studies were performed in microcosms under laboratory conditions and other more important factors could override the effect of vegetation under field conditions. Hence, in Paper V we wanted to quantify the effect of vegetation relative to another important factor for nitrate removal, i.e. hydraulic load, at the field scale.

METHODS

Microcosm process studies

To study the nitrogen transformation processes, laboratory microcosm studies where environmental factors can be controlled are useful. Using this approach, selected samples are collected from the wetland and the processes are measured during laboratory incubations. These methods can be carried out during in situ conditions to measure the actual process rates in field, or under optimal conditions, to measure potential process rates.

Nitrification rates are commonly measured as an increase of nitrite and nitrate per hour. The isotope-dilution technique (Koike and Hattori 1978; Rysgaard et al. 1993) is one common method to measure the production of nitrate. 15N-nitrate is added to the aquatic phase of the samples. Nitrifying bacteria dilute the 15N-nitrate pool by oxidizing 14N- ammonium to 14N-nitrate. The dilution of 15N-nitrate is considered proportional to nitrification. The 14N- and 15N-nitrate content can be measured according to Davidsson et al. (1997) after converting nitrate to 28 nitrogen gas using a denitrifying culture (Pseudomonas nautica). N2, 29 30 N2 and N2 are measured using a masspectrometer (Hewlett-Packard

12 MS engine quadropole mass spectrometer). In the study of Paper II, potential nitrification was measured using this isotope-dilution technique.

There are several methods to measure denitrification activity. Among the most common ones are the acetylene inhibition technique, the 15N tracer technique and the N2 flux technique (Leonardson 1992; Seitzinger et al. 1993). The acetylene inhibition technique is a method that is used to indirectly measure the denitrification rate. Acetylene blocks the last step in the denitrification process, the reduction of to dinitrogen gas. If denitrification occurs in a sample, addition of acetylene to the sample causes an accumulation of nitrous oxide. This nitrous oxide can easily be measured using a gas chromatograph. However, the acetylene also inhibits the nitrification. If the denitrification process is mainly supplied with nitrate from the nitrification process in the sample, this can 15 result in severe underestimations and the N tracer technique or the N2 flux technique is preferable.

The 15N tracer technique is a method where both denitrification and nitrate produced through nitrification can be measured (Nishio et al. 15 1983). The isotope NO3 is added to the denitrifying sample and the dilution of the isotope is measured as described in the isotope-dilution technique for nitrification. This is measured with masspectrometer by converting all nitrate to dinitrogen gas and measure the weight of the gas. This method needs high concentrations of nitrate and is thus not a good method when measuring actual activity of samples in which denitrification is limited by nitrate concentrations. The 15N tracer technique is also a costly method. The N2 flux technique is a direct measurement of the end product in the denitrification process, the dinitrogen gas (Seitzinger et al. 1993). The background concentration of dinitrogen gas is lowered from 79 to 1% using helium and molecular oxygen and then it is possible to measure an increase in dinitrogen gas with a gas chromatograph. This technique needs gas tight incubations and is sensitive to contamination, as there is high concentration of dinitrogen gas in the surroundings. In summary, when the nitrate concentrations are high in the water and the denitrification process is not mainly supplied with nitrate from the nitrification process, the acetylene inhibition method is an easy and cheap method to use.

In Paper II, III and IV, the acetylene inhibition technique was used as the nitrate concentrations were very high and potential denitrification only, i.e. no coupled nitrification and denitrification, was of interest. In the incubations of intact cores, the acetylene and nitrate might diffuse more or less rapidly into the sediment depending on the sediment

13 characteristics, the water movements of the overlying water and the time of incubation. This might both underestimate and overestimate the potential denitrification that would occur in field. In the experiments of Paper III and IV, the potential denitrification might be overestimated by faster stirring of the overlying water, causing a more rapid diffusion than in situ, while this might have been underestimated in Paper II. In all incubations, the incubation time might influence the diffusion of acetylene and nitrate. The longer time, the lower down into the sediment acetylene and nitrate would diffuse. However, even though this compromise comparison between the rates measured in the different experiments, the relationship between the samples within each experiment is not likely to be affected.

Model approach for describing nitrogen removal

The nitrogen removal in treatment wetlands is usually described using first-order equations based on input and output data of the wetland, i.e. mass balance models. Higher nitrate concentrations in the inflow result in a steeper concentration gradient in the sediment which would enhance the diffusion rate of nitrate. Increased rate of nitrate diffusion into the sediments where the denitrification occur increase the nitrogen removal when nitrate is the limiting factor. This has been shown in flooded soils where diffusion of nitrate into the sediments was limiting denitrification, and the nitrate removal rate was best described as being linearly related to the nitrate concentration (first-order) (Reddy and Patrick 1984). Hence, the first-order equation (Eq. 1) is most commonly used in many wetland models to describe overall total nitrogen removal when diffusion is not considered explicitly in the model (Kadlec and Knight 1996; Knight et al. 2000).

J = k ⋅C (Equation 1) where J is the rate of the nitrogen transformation process (g m-2 y-1), k is the first-order area-based reaction rate coefficient (m y-1) and C is the concentration of nitrogen (g m-3).

First-order design models are either area dependent (Eq. 1), and thus estimate nitrogen removal per unit wetland area, or volume dependent, and thus estimate nitrogen removal per unit wetland volume (Kadlec 2000; Kadlec and Knight 1996; Knight et al. 2000; Spieles and Mitsch 2000). Models used in a cold climate region usually also include temperature as a factor e.g. Arheimer and Wittgren (1994). Temperature

14 effects on nitrogen removal are most commonly described using a modified Arrhenius temperature relation (Paper I), where the temperature has stronger effect on the nitrogen process at lower temperatures. In the models, the water column is usually assumed to be totally mixed every time step or to follow a plug flow, where the water volume of each time step is treated separated from the others. These assumptions are however simplifications of the water flows in a wetland and tracer studies have shown that the water flows in wetlands usually follow patterns that can be described as a combination of these two assumptions. After evaluating tracer studies in wetlands, the tank-in-series model was developed to better mimic the hydraulics (Kadlec and Knight 1996). In the tank-in- series model, the wetland basin is modelled as consisting of several cells, tanks, that are assumed to be totally mixed every time step. These cells are connected and the nitrogen discharged from the first cell is entering the second. Tracer studies can be used to analyse the hydraulic efficiency, i.e. how well the water is distributed over the wetland (Persson 1999). Using tracer studies, the number of cells (N) that would be suitable for a certain wetland can be estimated and used in the first-order equation (Paper V).

The first-order reaction rate coefficients are then used to compare the efficiency of different wetlands. However, the use of this approach has been questioned (Kadlec 2000). Published area-based nitrogen removal rates vary largely between wetlands (0.5 < k < 70 m yr-1) (Kadlec and Knight 1996) making the prediction of nitrogen removal in specific wetlands quite uncertain. The large variation has several causes. All nitrogen transformation processes and affecting factors are lumped together and seasonal changes, like temperature, are not considered in all calculations. Studies conducted in wetlands receiving mainly, or high concentrations of, nitrate might be more comparable as one nitrogen process, denitrification, is usually dominating for the nitrate removal. But also the nitrate removal rates in such wetlands have resulted in a very large variation (10 > k > 360 m yr-1) (Arheimer and Wittgren 1994; Burgoon 2001; Kadlec and Knight 1996). These rates are, however, also influenced by several different factors that could be limiting nitrate removal, such as organic carbon availability and hydraulic efficiency, which are not included in the calculations.

Researchers have tried to describe the seasonal variation of nitrogen removal using the lumped model approach. Spieles and Mitsch (2000) were able to explain 80-96% of the average annual nitrate removal in three wetlands using the four parameters, hydraulic loading, nitrate loading, wetland volume and temperature. The model was on the other

15 hand poor in describing the seasonal variation (R2=0.40), probably due to flooding events. Arheimer and Wittgren (1994) used the same model approach as in Paper I and succeeded to describe the seasonal variation very accurately (R2=0.92). However, the wetlands had occasionally very short residence time and a very low relative nitrate removal, which made the incoming concentrations closely related to the outlet concentrations and thus the description of the nitrogen processes was not as critical. Other studies have presented more complex models for wetlands receiving both ammonium and nitrate, containing more detailed hydraulics and several nitrogen transformation processes (Dørge 1994; Jørgensen 1994; Martin and Reddy 1997; Werner and Kadlec 2000). A serious drawback of these more complex approaches is that a lot of data of the processes are needed to properly calibrate and validate the models. There are not many studies that have used models of intermediate complexity, between lumped mass balance nitrogen removal models and complex process focused models. To improve the treatment wetland design modelling, models containing more internal processes than in the lumped total nitrogen removal models, and less than in the complex models, are needed. In Paper I, one such model was developed from data commonly available from monitoring programs. The model used was intermediate in both complexity and realism between lumped mass balance nitrogen removal models and complex process focused models. First-order area-based equations were set up to describe three major processes in two different wetlands; ammonification, nitrification and denitrification. The water in the wetland of Oxelösund was modelled as six basins with completely mixed conditions within each basin, while the wetland of Hässleholm was considered as only one totally mixed basin. An advantage of this model approach was that it recognised the reactivity of individual nitrogen transformations, and for the wetland of Oxelösund, also that the water regime was considered.

The study of Paper V was performed in the experimental wetlands of Halmstad. These are pilot-scale wetland basins, and a great advantage is that replicates have been used, providing unique opportunities to statistically test the effects of different factors. The area-based first-order kinetic coefficients for nitrate removal were calculated using the tank-in- series model approach. The first-order rate coefficients were used to compare the performance of the basins with different treatments. All basins were small and similar in shape and the hydraulics was assumed to be of minor importance for the differences between the basins. Further, different temperature dependencies were applied and the effect of these on the first-order rate coefficients was analysed.

16 SPATIAL VARIABILITY OF THE NITROGEN REMOVAL PROCESSES

Large constructed wetlands with large nitrogen removal (kg ha-1 y-1) are usually very heterogeneous with basins, or parts of basins, containing different vegetation and having different water flow patterns. Nitrogen transformations in such wetlands are complex in both time and space, and different factors are limiting the nitrogen transformation processes in different parts of the wetland and during different times of the day or year. The model used in Paper I was site-specific for the two different wetlands. This model approach was an improvement of a lumped total nitrogen model, by separating three different nitrogen processes; ammonification, nitrification and denitrification, but it did not have the flexibility in predicting outflow concentrations over a wide variety of conditions. The calibrated model described the seasonal variation of both ammonium and nitrate removal well in Hässleholm, but poorly in Oxelösund. Also, the calibrated reaction rate coefficients for ammonium removal and nitrate removal were very different (Paper I).

The reaction rate coefficients for nitrate removal in Oxelösund were very high compared with other studies, but similar rates have been shown in wetlands with similar water quality, i.e. high ammonium concentrations and low nitrate concentrations (Gerke et al. 2001). This could be a result of a very intense denitrification process, but more likely it reflected that the first-order equation was not suitable for such low nitrate concentrations since diffusion was not included in the model. The choice of equation has been discussed in other studies (Liehr et al. 2000; Reddy and Patrick 1984), also pointing out the dependence on nitrate -1 concentration. At low nitrate concentrations (<2 mg NO3-N L ), denitrification can be regulated by nitrate concentration (Kadlec and Knight 1996) and diffusion, and nitrogen removal might be better explained using a second-order equation. In a wetland receiving very high -1 nitrate concentrations (>400 mg NO3-N L ), the first-order model clearly overestimated the nitrogen removal (Liehr et al. 2000), and the authors suggested the use of a zero-order model under such conditions. At such high nitrate concentrations, denitrification may be nitrate saturated and hence regulated by neither concentration nor diffusion.

The wetland of Oxelösund could be considered more complex than that of Hässleholm both in water quality and vegetation, as the wetland of Oxelösund in large areas was covered by dense emergent vegetation. The Oxelösund wetland was also receiving mainly ammonium nitrogen and thus both nitrification and denitrification has to occur to remove nitrogen.

17 In contrast, the Magle wetland in Hässleholm received almost equal amounts of ammonium and nitrate, had mainly open water surface areas and, in parts of the wetland, dense submersed vegetation. The varying environment in these wetlands probably favour different nitrogen processes, and it is likely that the spatial heterogeneity within a wetland is an important component when trying to describe the nitrogen transformation processes in a model. In less complex wetlands, where only one factor is limiting the wetland system within the whole wetland and throughout the year, e.g. nitrate concentrations or temperature, simple models might better describe the nitrogen removal.

In Paper II, the spatial differences within a wetland of different nitrogen processes were investigated. Large differences in potential nitrification and denitrification rates were found between different structures in a wetland, i.e. submersed macrophytes, filamentous algae, twigs and the sediment. This indicated a substantial spatial variability for the nitrogen transformations in wetlands. The results showed that the potential for nitrification was highest on different surfaces in the water column, while the denitrification capacity was highest in the sediment. Hence, the use of an area-based rate coefficient as in Paper I, was less suitable for describing the transformation of ammonium, while for denitrification the area-based assumption was acceptable, although sediments might be very diverse and denitrification may occur in thin or thick layers of the sediment (Nielsen et al. 1990a).

The nitrification potential was shown to be highest on twigs, followed by sediment and plants (Fig. 4). In the sediment, nitrifying bacteria might be affected by the presence of other heterotrophic bacteria through competition for oxygen which would explain the lower potential (Zhang and Bishop 1996). Surprisingly, the rates (kg N ha-1 d-1) on twigs were 100 times higher than those noted for Myrriophyllum spicatum. The photosynthetic activity of plants producing oxygen could be assumed to favour the nitrifying bacteria. Interestingly, other investigators have reported values representing nitrification on litter of emergent macrophytes that were also much higher than the corresponding values for the biofilms on the living submersed plants in the present study (comparison based on activity per dry weight of substrate) (Eriksson and Andersson 1999). It has been suggested that plants might restrict biofilm development and nitrification in epiphytic microbial communities by release of inhibitory chemical compounds (Lodhi 1982; Rice and Pancholy 1972). Such impeding effects might explain the varying nitrification capacity displayed on twigs, litter and living submersed plants. For denitrification, the sediment was the most important site,

18 followed by twigs and plants (Fig. 5). The submersed plants showed very low denitrifying capacity expressed per shoot area compared to the other surfaces. However, expressed per wetland area and assuming the highest abundance of submersed vegetation, the denitrification capacity of the submersed plants with epiphytic biofilms was in the same range as the sediments, which has also been shown by Eriksson and Weisner (1997) and Toet (2003). The greater denitrification capacity in the sediment was probably due to a high amount of available organic matter.

Figure 4. Potential nitrification (black bars), potential denitrification (striped bars), and nitrate removal (white bars) on the surfaces of Eurasian watermilfoil (Myrriophyllum spicatum), twigs, sediment, and macroalgae (biofilm area-1 in A; dry wt.-1 in B). The samples were collected in Basins B and D (n = 5) in Magle wetland, Hässleholm. Error bars show SD. (Reprinted from Paper II).

THE EFFECT OF ORGANIC MATTER ON DENITRIFICATION

Organic carbon is supplied to the sediment by plants in a wetland and several studies have shown that the supply of the organic matter often limits the denitrification in treatment wetlands that receive high concentrations of nitrate (Burgoon 2001; Ingersoll and Baker 1998;

19 Kozub and Liehr 1999; Zhu and Sikora 1995). Sediments have different characteristics depending on their organic content. Wetland sediments with low organic content have been shown to have much lower nitrogen removal and denitrification capacity compared to those with high organic content (Craft 1997). The vegetation is thus important as a source of the detritus that fuels the denitrification and by creating a favourable environment in other ways, e.g. lowering the oxygen concentrations through increased heterotrophic activity in the sediment, or supplying surfaces for microbial growth.

In the study of Paper III and IV, the effect of organic matter from different plant species on denitrification was investigated. The results of Paper IV showed that the organic matter in stands of Glyceria maxima and Potamogeton pectinatus was more available, i.e. easily degradable, to the denitrifying bacteria than that of the stands of Typha latifolia and Phragmites australis. The litter from a submersed plant is usually more easily degradable than the litter from an emergent plant (Godshalk and Wetzel 1978). However, when measuring the denitrification capacity in intact sediment cores, more closely resembling wetland conditions, the organic sediment of the emergent plant G. maxima showed the highest denitrification capacity, followed by T. latifolia (Fig. 5). The denitrification in the sediment cores was limited by organic carbon and the lower denitrification capacity in the stands of the submersed plant P. pectinatus was likely due to less amount of organic matter and not the availability of the organic matter.

The potential denitrification rates in the intact cores of Paper IV were highest in autumn (November), while when measuring potential denitrification in slurries, the highest denitrification rates were found in May and August. This indicates that the organic matter was more available, i.e. easily degradable, in May and August, and that the higher denitrification rates in November for the intact cores were due to larger quantities of organic material. The higher quality of the organic matter of P. pectinatus was, in this study, not enough to compensate for the much lower organic matter production. However, in a study of another wetland (Paper III), intact cores of the submersed plant Elodea canadensis had a higher denitrification capacity than the cores from the emergent plants T. latifolia and P. australis in November. This was probably explained by larger amounts of available organic matter from that submersed plant. Further, some cores of E. canadensis included fresh plants with epiphytic biofilms that could have contributed to high denitrification. The impact of submersed plants and epiphytic biofilms was addressed by Weisner et al. (1994) showing that microcosms containing the fresh submersed plant P.

20 pectinatus with epiphytic biofilm, had higher denitrification rates than litter of G. maxima. Altogether, these studies suggest that both the availability and the amount of organic material are critical for the denitrification capacity in a wetland, and emphasize the importance of dense vegetation, and thus high amounts of available organic material, for high annual wetland nitrate removal.

) ) -1 -1 0.4 , h , h -2 -2 0.35 incubation 1 incubation 2 incubation 3 0.3 O-N m O-N O-N m 2 2 0.25

0.2

0.15

0.1 0.05 0 Incubation 3 10( g N x Incubation (g 2 1 and N cc c ccc c ccc c c May Aug Nov Nov Apr Apr May Aug Nov Nov Apr Apr May Aug Nov Nov Apr Apr G. maxima P. pectinatus T. latifolia

Figure 5. Potential denitrification rates (g N m−2 h−1) in intact cores containing organic material of different plant species from the sampled Källby wetland (Experiment I). C denotes acetate addition in incubations 2 and 3. In incubation 1, no external carbon source was added. Error bars indicate SD. (Reprinted from Paper IV).

The plants in a wetland make the site very complex and diverse biologically, chemically and physically. In Paper III and IV the effect of the organic matter on denitrification was investigated in a laboratory environment, and the denitrification rates were measured during optimum environmental conditions. In the field, several environmental factors vary and affect the denitrifying bacteria. The quantitative importance of vegetation might thus be different depending on the type of vegetation and the season, and in addition, other more important factors could override the effect of vegetation under in situ wetland conditions.

In the study of Paper V, nitrate removal was investigated in pilot-scale wetlands with different vegetation and water flows. The mass balances (inflow and outflow measurements) showed vegetation to be an important factor for high nitrate removal despite high hydraulic load and cold

21 climate, and the highest removal was found in dense stands of emergent vegetation dominated by P. australis and G. maxima (Fig. 6). Also other studies have shown that plants are important under in situ nitrate removal in wetlands (Bachand and Horne 2000; Lin et al. 2002). The wetland systems in Paper V achieved higher nitrate removal as the wetlands matured and the density of vegetation increased. This effect has been suggested by Craft (1997), who observed higher nitrate removal in wetlands where organic material had accumulated for about 10 years than for 1-5 years. The organic matter was probably important for the denitrifying bacteria as a source of organic carbon, which was also shown in the microcosm studies of Paper IV. Further, thicker organic matter layer at the sediment would probably lower the oxygen concentrations, through high heterotrophic activity (Nielsen et al. 1990a). In the study of Paper V, the first-order area based rate coefficients were highest during autumn, which was similar to the results presented in Paper IV, where the denitrification potentials in intact sediment cores were highest in November, probably due to more organic matter in the sediments.

Emergent Submersed Free development

0.7 )

-1 0.6 d

-2 0.5 0.4 0.3 0.2 0.1 0 -0.1 Nitrate-N removal (g m -0.2 low low low low low low low low low low low low low high high high high high high high high high high high high high 1234123412341

2003 2004 2005 2006

Figure 6. Nitrate nitrogen removal (g m-2 d-1) in wetlands containing emergent, submersed and freely developed vegetation with different hydraulic loads (high: 0.13 m d-1 and low: 0.39 m d-1) and during different seasons (Period 1: December to February; period 2: March to May; period 3: June to August; period 4: September to November). Error bars show standard errors.

22 CRITICAL HYDRAULIC LOAD FOR NITRATE REMOVAL

Mass balance studies (inlet and outlet measurements) have shown that higher hydraulic loads, and usually also higher nitrate loads, result in higher area-specific nitrate removal (kg per hectare and year) (Bachand and Horne 1999; Kadlec 2005). Surprisingly, the mass balances in Paper V showed no significant difference in either nitrate removal (kg ha-1 y-1) or first-order rate coefficients between the basins of different hydraulic loads. Very high hydraulic load and short residence times of the water has been shown to be crucial for the nitrate removal, causing negative effects for nitrate removal (Raisin and Mitchell 1995; Spieles and Mitsch 2000). Arheimer and Wittgren (2002) suggested that the critical mean residence time for positive annual nitrogen removal would be more than two days, based on data sets from Swedish wetlands receiving hydraulic loads of 0.26-6.8 m d-1. The two different hydraulic loads in Paper V were thus set to 0.13 and 0.39 m d-1, which resulted in residence times of 3 and 1 days, respectively. The results showed no negative effects of those hydraulic loads. However, the wetlands in Paper V received a constant hydraulic load, while the wetlands in the study of Arheimer and Wittgren (2002) received very varying hydraulic loads during the year. The 2 days critical residence time for significant nitrate removal suggested by Arheimer and Wittgren (2002), was an annual mean value, and the residence times could occasionally have been much shorter in all the studied wetlands, causing such negative effects that the resulting annual nitrate removal was insignificant in some of them.

Other studies, summarized in Table 1, also suggest that there is an upper limit for how large the hydraulic loads can be to achieve positive nitrate removal. Critical hydraulic loads for positive nitrate removal in these wetlands was lower than 1.5 m d-1 (Braskerud 2002; Koskiaho et al. 2003; Svedin 2006). One mass balance (inflow and outflow measurements) study of a wetland receiving pre-treated municipal wastewater was performed in September and November (Svedin 2006), and the results of that study showed no significant nitrate removal in November at hydraulic loads more than 1.4-2.0 m d-1, which was equivalent to 6 to 7 hours residence times (Table 1). Braskerud (2002) also showed no significant nitrate removal in wetlands with hydraulic loads larger than 1.7-1.8 m d-1. However, the critical residence time might vary with the wetland conditions. The hydraulic load, residence times and water velocities is affected by the hydraulic efficiency in the wetland, i.e. the wetland area that can be used for nitrogen transformations. It is thus important to achieve high hydraulic efficiency to attain considerable nitrogen removal in wetlands. Estimated critical residence times might

23 also be related to the precision in the measurements of nitrate concentrations and water flows. As the difference between inlet and outlet concentration is very small when the residence time is short, precise measurements are very difficult to perform.

Table 1. Characteristics of Nordic systems receiving high hydraulic loads.

Site Age Area Depth Hydraulic Conc. in Conc. in Nitrate Vegetation Ref load NO3 Tot-N removal .

(y) (ha) (m) (m d-1) (mg l-1) (mg l-1)

Finland 8 0.6 0.5-1 0.41 7.9 9.8 Positive Scirpus sylvaticus, T. 1 removal latifolia (Biomass: 44 g m-2) 10 0.48 0.5-1 2.9 6.8 8.4 No T. latifolia, Juncus removal filiformis (Biomass: 91 g m-2)

Norway 7 0.03- 0.2- 0.7-1.6 2.2 3.2-3.5 Positive Scirpus lacustris, 2 0.09 0.8 removal Acorus calamus

3-4 0.03- 0.2- 1.7-1.8 0.75-2.8 1.6-5.1 No G. fluitans, 0.09 0.8 removal Sparganium erectum

Sweden 9 0.3 0.2- 1.3-1.5 19 20 Positive P. australis 3 0.3 removal

9 0.3 0.2- 1.4-2.0 9.4 20 No P. australis 0.3 removal

1) Koskiaho et al. (2003) 2) Braskerud (2002) 3) Svedin (2006)

COMPARING WETLAND PERFORMANCE

In Table 2, some design characteristics are summarized in -like free water surface wetlands in Sweden with similar temperate boreal climatic conditions. All these wetlands receive water with high concentrations of nitrate and thus denitrification is likely to be the limiting process for nitrate removal.

24 Table 2. Characteristics and nitrate removal rate coefficients of Swedish constructed wetland systems. Summarised below are wetland age, area and vegetation, mean water depth, hydraulic load and inlet nitrate and total nitrogen concentrations, first-order area-based rate coefficients for nitrate removal and temperature coefficients.

Site Age Area Depth Hydraulic Conc. in Conc. in Temp Rate Vegetation Ref load NO3 Tot-N coeff. coeff. kA20

(y) (ha) (m) (m d-1) (mg l-1) (mg l-1) (m yr-1)

Eskilstuna 7 28 1 0.16 15 20 1.08 30 Emergent 1 (May- (20%); G. Oct) maxima and Typha sp.

Southern 0.1-3.0 0.7- 0.26-6.8 4.6-17 4.6-17 1 17 No or 2 Sweden 2.1 submersed vegetation

Hässleholm 11 20 0.5-2 0.057 14 21 1.116 29 Submersed; M. 3 (Paper I) picatum and E. canadensis

Nykvarn 13 0.6 0.3- 0.03-0.05 5.8-14 21-26 1.08 33 Emergent 4 0.5 (80%); P. australis, T. latifolia and submersed; E. canadensis

Halmstad 4 0.002 0.4 0.13 and 11 11 1.088 21 Emergent; P. 5 (Paper V) 0.39 (emerge australis, G. nt), 15 maxima and (submer submersed; E. sed) canadensis

NADB 1.088 1-34 Woddy 6 vegetation, forested

NADB 1.088 26-55 Nonwoddy, 6 emergent vegetation

1) Andersson et al. (2005) 2) Arheimer and Wittgren (2002) 3) Kallner and Wittgren (2001), Paper I 4) Tonderski (2000) 5) Paper V 6) Kadlec (2005)

The highest first-order rate coefficients were in the wetland of Nykvarn (Tonderski 2000). This wetland contained the highest density of emergent

25 vegetation and probably had the highest accumulation of organic matter in the sediments. Surprisingly, the wetland of Hässleholm (Kallner and Wittgren 2001) showed similar performance as Nykvarn, though this wetland contained submersed vegetation. Also, the nitrate removal rates in Eskilstuna were in the same range, containing much larger areas of open water than in Nykvarn. The wetland of Eskilstuna was only measured during the warm season, and thus the annual removal should be lower.

The first-order rate coefficients of the wetlands in Arheimer and Wittgren (2002) and in Paper V were low. The wetlands in Arheimer and Wittgren (2002) received hydraulic loads with strong seasonal variation, which would decrease the removal rate coefficient as the main hydraulic load occurs during the cold season. In addition, the wetlands in Arheimer and Wittgren (2002) and Paper V contained more sparse vegetation and were not as matured systems, and organic material might not have had the time to accumulate in the sediments. The latter was most likely the reason for the lower rate coefficients in the wetlands in Paper V, as there were higher nitrate removal rate coefficients in the wetlands of dense emergent vegetation than in the wetlands of submersed vegetation between replicated wetland basins. Further, the wetland-specific hydraulics were not considered in the first-order area-based rate coefficients in Table 2. In the large complex wetlands, short-circuiting of the water, i.e. less efficient wetland area, can cause much lower wetland performance (Persson 2005).

All these Swedish wetlands have a nitrate removal, described with first- order rate coefficients, that are in the lower range of the wetlands described in Kadlec (2005). Possible reasons for this could be the climatic differences. Wetlands that are located in warmer climate have a longer period with high biological activity. This is critical for the primary production of the vegetation, and the supply of organic carbon would be larger in warmer climates than in colder. This indicates that the nitrate removal of these wetlands might be possible to increase and that design is crucial. To make better predictions of nitrate removal in constructed wetlands, the accumulation of organic matter in the sediment might be important to consider.

26 CONCLUSIONS AND FUTURE STUDIES

Conclusions

Simple models not accounting for spatial heterogeneity and vegetation performs poorly to describe the seasonal nitrogen removal in complex wetlands (Paper I). In such complex wetlands, different factors are limiting the nitrogen transformation processes in different parts of the wetland and during different times of the day or year. In less complex wetlands, where only one or a few factors are limiting the nitrogen removal within the whole wetland and throughout the year, e.g. nitrate concentrations or temperature, simple models might better describe the nitrogen removal.

The results in Paper II emphasised the importance of spatial differences for the nitrogen processes, and suggested that the most active sites for nitrification and denitrification were separated spatially in a wetland. Nitrification took place largely on twigs in the water column, and this process might thus be possible to enhance through supplying the bacteria with more dead structures in the wetland water column. However, the use of an area-based rate coefficient as in Paper I was not accurate for ammonium removal, as the sediment was not the main site for nitrification. For denitrification, the most important site in the wetland was the sediment, and thus the area-based assumption would be acceptable, although different characteristics of the sediments might explain differences in rate coefficients between wetlands.

Sediments have different characteristics depending on their organic matter content. Denitrification capacities may be affected more than three fold by different types of litter and detritus in the sediments (Paper III and IV), and the highest denitrification capacities were found in sediments with organic matter from E. canadensis (Paper III) and G. maxima (Paper IV). The vegetation was thus important by supplying organic matter that fuels the denitrification, or by creating a favourable environment in other ways, e.g. supplying surfaces for microbial growth.

The vegetation provided highest quality of the organic matter during summer, but the wetland area based potential denitrification rates, measured in intact cores, were greatest during autumn, when the organic matter was most abundant (Paper IV). Hence, both the availability and the amount of organic material were critical for the denitrification capacity in a wetland.

27

The vegetation composition was a critical factor for nitrate removal in wetlands receiving high hydraulic loads (Paper V). The nitrate removal was higher in the basins of emergent plants than in the basins of submersed plants and freely developed vegetation, and the removal increased with wetland age and more dense vegetation. However, no difference in nitrate removal in basins with 1 and 3 days residence times could be detected. In summary, these studies emphasised the importance of dense emergent plants for high annual wetland nitrate removal.

Future studies

For better prediction of nitrate removal in future wetlands, models are great tools. To improve the prediction capacity of the models, evaluation of existing wetlands is of major importance for the understanding of the treatment performance during field and full-scale conditions. Analyzing nitrate removal at field and full-scale conditions can allow better understanding of the effect of hydraulics and vegetation and how the wetlands change with age. The studies in this thesis suggest that the characteristics of dead organic matter in the sediment and water column could be important to include in the first-order area-based kinetic model for nitrate removal. The organic matter changes between seasons and a description of this might improve the general applicability of the model. To further test the effect of different factors that could improve the treatment performance, e.g. different plant species, pilot-scale wetlands are very useful. Interesting studies would be to investigate the amount of available carbon in relation to total carbon biomass of different plant species required for optimal denitrification at different conditions, e.g. amount of nitrate. Finally, the possibility of optimizing different wetland services in one multipurpose wetland, e.g. nitrogen, phosphorous and BOD removal, biodiversity and biomass production, would be interesting to explore.

ACKNOWLEDGEMENT

I thank Karin Tonderski, Lars Leonardson and David Bastviken for valuable comments on this thesis. This thesis was financed by MISTRA (the Foundation for Strategic Environmental Research) through VASTRA (Swedish Water Management Research Programme).

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31 Lodhi MAK (1982) Additional evidence of inhibition of nitrifiers and possible cycling of inhibitors produced by selected plants in a climax community. Bulletin of the torrey botanical club 109:199- 204 Martin JF, Reddy KR (1997) Interaction and spatial distribution of wetland nitrogen processes. Ecol. Model 105:1-21 Nielsen LP, Christensen PB, Revsbech NP, Sørensen J (1990a) Denitrification and oxygen respiration in biofilms studied with a microsensor for nitrous oxide and oxygen. Microb. Ecol. 19:63-72 Nielsen LP, Christensen PB, Revsbech NP, Sørensen J (1990b) Denitrification and photosynthesis in stream sediment studied with microsensor and whole-core techniques. Limnol. Oceanogr. 35:1135-1144 Nielsen SL, Sand-Jensen K, Borum J, Geertz-Hansen O (2002) Depth colonization of eelgrass (Zostera marina) and macroalgae as determined by water transparency in Danish coastal waters. Estuaries 25:1025-1032 Nishio T, Koike I, Hattori A (1983) Estimates of denitrification and nitrification in coastal and estuarine sediments. Appl. Environ. Microbiol. 45:444-450 Persson J (1999) Hydraulic efficiency in pond design. Phd-thesis. Department of hydraulics. Chalmers university of technology, Göteborg Persson J (2005) The use of design elements in wetlands. Nordic Hydrology 36:113-120 Phipps RG, Crumpton WG (1994) Factors affecting nitrogen loss in experimental wetlands with different hydrologic loads. Ecol. Eng. 3:399-408 Prescott LM, Harley JP, Klein DA (1990) Microbiology, second edn. Wm. C. Brown Communications, Inc., Dubuque, IA Prosser JI (1989) Autotrophic nitrification in bacteria. Adv. microb. physiol. 30:125-179 Raisin GW, Mitchell DS (1995) The use of wetlands for the control of non-point source pollution. Wat. Sci. Tech. 32:177-186 Reddy KR, Patrick WH (1984) Nitrogen transformations and loss in flooded soils and sediments. Critical Reviews in Environmental Control 13:273-309 Reddy KR, Patrick WH, Lindau CW (1989) Nitrification-denitrification at the plant root-sediment interface in wetlands. Limnol. Oceanogr. 34:1004-1013 Rice LE, Pancholy SK (1972) Inhibition of nitrification by climax ecosystems. American journal of Botany 59:1033-1040

32 Rysgaard S, Risgaard-Petersen N, Nielsen LP, Revsbech NP (1993) Nitrification and denitrification in lake and estuarine sediments measured by the 15N dilution technique and isotope pairing. Appl. Environ. Microbiol. 59:2093-2098 Seitzinger SP (1988) Denitrification in freshwater and coastal marine ecosystems: Ecological and geochemical significance. Limnol. Oceanogr. 33:702-724 Seitzinger SP, Neilsen LP, Caffrey J, Christensen PB (1993) Denitrification measurements in aquatic sediments: A comparison of three methods. Biogeochemistry 23:147-167 SEPA (1997) Kväve från land till hav. Naturvårdsverket, Stockholm Spieles DJ, Mitsch WJ (2000) The effects of season and hydrologic and chemical loading on nitrate retention in constructed wetlands: a comparison of low- and high-nutrient riverine systems. Ecol. Eng. 14:77-91 Svedin C (2006) Nitrogen removal rates in a full-scale wetland dominated by emergent vegetation. Master thesis. Department of physics, chemistry and biology. Linköping university, Linköping, Sweden Svensson JM, Strand J, Sahlén G, Weisner S (2004) Utvärdering av våtmarker anlagda inom lokala investeringsprogram och med LBU-stöd avseende närsaltsretention och biologisk mångfald. Högskolan i Halmstad, Våtmarkscentrum, Halmstad Tanner CC, Clayton JS, Upsdell MP (1995) Effect of loading rate and planting on treatment of dairy farm wastewaters in constructed wetlands - II. Removal of nitrogen and phosphorus. Wat. Res. 29:27-34 Toet S (2003) A treatment wetland used for polishing tertiary effluent from a treatment plant: performance and processes. Phd- thesis. Biology. Utrecht, Utrecht Toet S, Logtestijn RSPV, Schreijer M, Kampf R, Verhoeven JTA (2005) The functioning of a wetland system used for polishing effluent from a plant. Ecol. Eng. 25:101-124 Tonderski KS (2000) Temporal variations in N and P removal in a SF wastewater treatment wetland in temperate climate. Preprint. Wetland systems for water pollution control, Florida Trepel M, Palmeri L (2002) Quantifying nitrogen retention in surface flow wetlands for environmental planning at the landscape-scale. Ecol. Eng. 19:127-140 Vretare V (2001) Internal oxygen transport to below-ground parts. Phd- thesis. Department of Ecology. Lund University, Lund Vymazal J (2001) Types of constructed wetlands for wastewater treatment: Their potential for nutrient removal. In: Vymazal J (ed)

33 Transformations of nutrients in natural and constructed wetlands. Backhuys Publishers, Leiden, pp 1-93 Weisner SEB, Eriksson PG, Granéli W, Leonardson L (1994) Influence of macrophytes on nitrate removal in wetlands. Ambio 23:363-366 Werner TM, Kadlec RH (2000) Wetland residence time distribution modeling. Ecol. Eng. 15:77-90 Westlake DF (1966) The biomass and productivity of Glyceria maxima. I. Seasonal changes in biomass. Journal of Ecology 54:745-753 Westlake J, Kvêt J, Szczepanski A (1998) The production ecology of wetlands. Cambridge university press, Cambridge Zhang TC, Bishop PL (1996) Evaluation of substrate and pH effects in a nitrifying biofilm. Wat. Environ. Res. 68:1107-1115 Zhu T, Sikora FJ (1995) Ammonium and nitrate removal in vegetated and unvegetated gravel bed microcosm wetlands. Wat. Sci. Tech. 32:219-228

34 TACK

Jag vill först tacka min handledare Karin Tonderski och mina bihandledare HB Wittgren och Lars Leonardson, för att ha introducerat mig in i våtmarkernas värld. Jag är väldigt tacksam för ert tålamod, engagemang, alla råd och våra mycket roliga och givande diskussioner. Tack också till Peder Eriksson för ett givande samarbete och att du tillsammans med flera på limnologen i Lund lärde mig metoder för att mäta kväveprocesser.

Jag är mycket glad för att ha varit en del i VASTRA. Genom våra diskussioner i doktorandgruppen har jag fått en bredare syn på miljöproblemen och förstått mer hur samhällsvetare kan se på saker och ting.

En viktig och inspirerande del av mitt arbete har varit mer tillämpad. Stort tack till Jonas Andersson, WRS i Uppsala, för ett samarbete som gett mig inspiration och engagemang och en bättre förståelse för hur våtmarker fungerar i verkligheten.

Mina rumskompisar under min tid som doktorand, Elisabeth Lundkvist, Jenny Bäckman, Hristina Bojcevska och Nicklas Jansson ska ha ett stort tack för att ni gör vårt rum till en så rolig plats. Tack även till mina kollegor på biologiavdelningen för många trevliga fikaraster.

Till sist, men mest av allt vill jag tacka min familj, David, Emanuel, Valdemar och Selma. David, min vetenskapliga rådgivare, men även min bästa vän. Tack för allt ditt arbete med att läsa manuskript, analysera problem och för att du alltid stöttar mig. Emanuel, Valdemar och Selma, tack för att ni finns.

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