Published in J. Environ. Qual. 32:2414-2420 (2003).
© ASA, CSSA, SSSA
677 S. Segoe Rd., Madison, WI 53711 USA
TECHNICAL REPORTS
Wetlands and Aquatic Processes
Potential Nitrification and Denitrification on Different Surfaces in a Constructed Treatment Wetland
Sofia Kallner Bastviken*,a,
Peder G. Erikssonb,
Irene Martinsc,
João M. Netoc,
Lars Leonardsonb and
Karin Tonderskia
a Dep. of Biology, Linköping Univ., SE-581 83 Linköping, Sweden
b Dep. of Ecology/Limnology, Lund Univ., SE-223 62 Lund, Sweden
c Dep. of Zoology, Univ. of Coimbra, 3004-517 Coimbra, Portugal
* Corresponding author (sofka{at}ifm.liu.se).
Received for publication April 28, 2002.
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ABSTRACT
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Improved understanding of the importance of different surfaces in supporting attached nitrifying and denitrifying bacteria is essential if we are to optimize the N removal capacity of treatment wetlands. The aim of this study was therefore to examine the nitrifying and denitrifying capacity of different surfaces in a constructed treatment wetland and to assess the relative importance of these surfaces for overall N removal in the wetland. Intact sediment cores, old pine and spruce twigs, shoots of Eurasian watermilfoil (Myriophyllum spicatum L.), and filamentous macro-algae were collected in July and November 1999 in two basins of the wetland system. One of the basins had been constructed on land that contained lots of wood debris, particularly twigs of coniferous trees. Potential nitrification was measured using the isotope-dilution technique, and potential denitrification was determined using the acetylene-inhibition technique in laboratory microcosm incubations. Nitrification rates were highest on the twigs. These rates were three and 100 times higher than in the sediment and on Eurasian watermilfoil, respectively. Potential denitrification rates were highest in the sediment. These rates were three times higher than on the twigs and 40 times higher than on Eurasian watermilfoil. The distribution of denitrifying bacteria was most likely due to the availability of organic material, with higher denitrification rates in the sediment than on surfaces in the water column. Our results indicate that denitrification, and particularly nitrification, in treatment wetlands could be significantly increased by addition of surfaces such as twigs.
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INTRODUCTION
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IN TREATMENT WETLANDS RECEIVING WATER that is rich in ammonium and nitrate, the turnover of N is controlled primarily by nitrification and denitrification. These processes are generally confined to microbial communities associated with the sediment or other submersed surfaces such as macrophytes and algae. However, a variety of surfaces may support diverse types of microbial communities that have varying capacities for nitrification and denitrification. Due to the physical complexity of wetlands, it is difficult to quantify nitrification and denitrification in the systems (Reddy and Patrick, 1984; Martin and Reddy, 1997; Kadlec, 2000; Kallner and Wittgren, 2001). Better understanding of the capacity of different surfaces to support nitrifying and denitrifying bacteria would facilitate more accurate estimation of nitrification and denitrification rates and allow improvement of the N removal capacity of treatment wetlands.
Bacteria attached to surfaces are usually more numerous and active than free-living bacteria (Hamilton, 1987). Attached bacteria form microbial communities that are embedded in polysaccharide matrixes, e.g., biofilms, and the bacterial activity within these biofilms is regulated by diffusion of nutrients into the biofilm and by internal processes within this layer (Atlas and Bartha, 1998). The characteristics of the surfaces on which biofilms develop influence the development of the microbial communities (Hamilton, 1987), and may thereby also have an impact on the potential for N transformations. In wetlands, surfaces suitable for attachment of bacteria are found on litter, wood, macrophytes, and algae. All of these surfaces are in close contact with the flowing water, and there is accumulating evidence that these surfaces are as important as the sediment for the N turnover processes.
Studies of N removal processes in freshwater systems have shown that submersed plants can enhance N removal by offering surfaces for both nitrifying and denitrifying bacteria (Eighmy and Bishop, 1989; Körner, 1999). Eriksson and Weisner (1999) observed substantially higher nitrifying activity in incubations that included the submersed macrophyte sago pondweed (Potamogeton pectinatus L.) than in incubations without this plant. The elevated nitrification was attributed to the epiphytic biofilm growing on the shoots of the macrophytes. Submersed plants are located in the water column, where there can be a good supply of ammonium and oxygen, and this environment should promote nitrifying bacteria.
Denitrification has also been shown to take place in biofilms on submersed plants (Law et al., 1993; Eriksson and Weisner, 1996). However, the main site of denitrification in natural systems is probably the sediment, because sediments provide more organic C and anaerobic environments than epiphytic biofilms and may therefore have a higher capacity to promote denitrification. However, in a system receiving nitrate-rich wastewater, denitrification on submersed macrophytes and macro-algae has been shown to be as important as the sediment for overall nitrate removal (Eriksson and Weisner, 1997; Eriksson and Weisner, 1999).
This emphasizes the spatial separation of nitrification and denitrification and the need for additional information about the types of surfaces that are important for the processes. The purpose of the present study was to compare the nitrifying and denitrifying capacity of microbial communities growing on the surfaces of filamentous macro-algae, submersed macrophytes, old pine and spruce twigs, and sediment in two basins of a constructed treatment wetland. Further, we wanted to assess the importance of these surfaces for overall N removal in the wetland.
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METHODS
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Site Description
The study was performed in a surface-flow wastewater treatment wetland, which was constructed in 1995 in southern Sweden (Magle, Hässleholm). The wetland was put into operation immediately after construction. It consists of six basins (A–F) with a total area of 20 ha (Fig. 1)
, and is surrounded by a small catchment with an area of 10 ha. Basins B through E have been divided into three sections separated by ridges, with a few shallow overflow areas. The water depth is on the average 45 cm with a maximum depth of about 2 m. The wetland receives biologically (trickling filters) and chemically treated (FeCl2 additions to remove P) wastewater from the wastewater plant of the municipal community of Hässleholm (18000 inhabitants). In 1996–1999, the wetland received an average of 9600 m3 d-1 wastewater. The average total N concentration in the incoming water was 21 mg L-1, dominated by NO-3–N and NH+4–N (14 and 7 mg L-1, respectively). Mean annual total N retention was 22%, or 788 kg ha-1 yr-1, and the mean P retention was 21%, or 5.3 kg ha-1 yr-1. The wetland was constructed on organic soil (loss on ignition 70 ± 10 SD % of dry wt., n = 5) partly on land that was formerly forested (Basins B and C), and partly on land that was previously used for agriculture (Basins D and E). The forested areas had peaty soil with a peat depth of about 4 to 5 m. In the areas with agriculture, the soil contained little peat, but had a significant content of sand. Basins B and D were the focus of this study. The inflows to Basins B–E are located at different distances from the wastewater inlet in Basin A; therefore, the nutrient concentrations were higher in the water entering Basin B than in that entering Basin D. The mean residence times for the B and D basins in 1999, based on tracer studies (lithium chloride), were 8 and 9 d, respectively, in July, and 14 and 10 d, respectively, in November. Basin B contained a large amount of twigs and wood from pine and spruce trees, whereas only a few twigs were present in Basin D. Twigs were defined as cylindrical pieces of wood with a diameter <2 cm. Although some twigs could be found in the water column of the wetland, this material was mostly confined to the bottom of the system. All basins contained submersed vegetation, mainly Eurasian watermilfoil—abundant and in dense communities in Basin D, but sparse in Basin B. Based on aerial photographs, submersed vegetation was estimated to cover about 40% of the area of Basin D but was not found in Basin B. In summer 1999, the biomass of Eurasian watermilfoil in Basin D was 51 ± 65 SD g dry wt. m-2 (n = 67, max. value: 320 g dry wt. m-2) within the volume of a cylinder (i.d.: 50 cm, entire water depth). Currently, we do not have information on why the abundance of submersed vegetation differed between Basins B and D. Emergent vegetation occurred in both B and D basins, but only at the shoreline of the wetland basins. Macro-algae—dominated by Ulothrix sp., Cladophora sp., Spirogyra sp., and Aulacoseira sp.—covered parts of the water surface in all basins, and were also observed attached to Eurasian watermilfoil and twigs. The biomass of these macro-algae in Basin D was 7 ± 6 SD g dry wt. m-2 (n = 24, max. value: 21 g dry wt. m-2) within the volume of a cylinder (50 cm i.d.) in summer 1999. Vegetation was harvested once during autumn in 1997, 1998, and 1999. During the harvest, macro-algae were removed, and submersed and emergent macrophytes were cut.
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Fig. 1. (A) The wetland of Hässleholm, comprising a common initial basin, (B–E) four parallel basins, and (F) a final common basin. The term x shows the sampling locations.
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Field and Laboratory Studies
Studies were conducted in the summer (29 June–23 July) and in autumn–winter (28 October–22 November) of 1999. Water discharge and water chemistry were monitored during the investigated periods. In connection with the measurements, laboratory experiments were performed to measure potential nitrification and denitrification rates in biofilms on various surfaces.
Water Chemistry and Water Discharge
Water samples were taken every second or third day during the investigated periods. The samples were collected in the middle of the water column (depth 0.3–0.6 cm) at the inlets of the two basins (B1 and D1), at the outlets of the first (B2 and D2) and second (B3 and D3) sections, and at the outlet of the third (last) section (B4 and D4). Water samples were filtered using Whatman GF/C filters, frozen, and later analyzed for content of nitrate, ammonium, and phosphate, whereas unfiltered water was frozen and later analyzed for total N and total P, using flow injection analysis (FIA, Tecator AB, Sweden). Dissolved organic carbon (DOC) was analyzed using a Shimadzu 5000 carbon analyzer. Water samples were collected for analysis of pH, alkalinity, and conductivity once during the sampling period in the summer and every second or third day during the sampling period in the autumn–winter. Dissolved oxygen and temperature were measured in the field using an oxygen probe (Oxygen guard Handy Mk III). The amount of wastewater flowing into Basin A was determined using an electromagnetic flow detector in the channel that connects the basin with the wastewater treatment plant. The water flow from Basin A to the Basins B–E was regulated as well as measured using rectangular weirs. Weirs were adjusted to relate water level to water discharge. Precipitation was measured at the treatment plant and evapotranspiration was calculated by the mass-transfer approach assuming free-water evaporation (Dingman, 1994).
Measurements of Potential Nitrification and Denitrification
Experiments were performed to ascertain whether potential nitrification and denitrification differed between surfaces within each basin and between the same surfaces from Basins B and D. Potential nitrification and denitrification rates were measured on green parts of Eurasian watermilfoil, on filamentous macro-algae, on twigs, and in sediment. Samples were collected in the middle section of Basins B (depth 1.2 m) and D (depth 0.6 m) (Fig. 1). Eurasian watermilfoil and macro-algae were present only in July. Scissors were used to cut Eurasian watermilfoil, macro-algae, and twigs, and the samples were collected and placed in plastic cylinders (i.d. 4.4 cm; length 25 cm). Algae attached to Eurasian watermilfoil and twigs were not removed. Intact sediment cores were collected using a sediment sampler and the cylinders (i.d. 4.4 cm; length 25 cm) were filled with equal parts of sediment and water. Water from the different basins was collected and filtered through a net (50-µm mesh size) to remove large particles. The samples were stored in darkness at 20°C overnight. The experiments were performed the following day. To avoid anaerobic conditions in the water of the cylinders during storage, the water was aerated using plastic tubes that were connected to air pumps. The aeration equipment was placed to avoid disturbing the sediment surface.
The samples of sediment, twigs, Eurasian watermilfoil, and macro-algae from the two basins were incubated in wastewater from the inlet basin. The incubation water was enriched with 5 mg NH4–N L-1, 0.7 mg PO4–P L-1, 2.5 mg acetate-C L-1, 2.5 mg glucose-C L-1, and about 5 mg 15NO3–N L-1 in summer and 25 mg 15NO3–N L-1 in autumn to obtain final concentrations of about 15 mg NO3–N L-1 and 40 mg NO3–N L-1 in summer and autumn, respectively. Essential nutrients were added in excess to the incubations to assure that potential nitrification and denitrification rates were a function of the size of the nitrifying and denitrifying populations, and not the result of nutrient limitation. Because of a miscalculation, the nitrate additions differ between summer and autumn–winter incubations. However, preliminary studies had shown that the additions of nitrate made both in summer and in winter were sufficient to maximize denitrification, i.e., give correct values of the potential denitrification activity.
The samples were measured consecutively for nitrification and denitrification. In all experiments, aerobic and anaerobic conditions were achieved through bubbling the water phase for 45 min with air or N2 gas, respectively. For measuring nitrification and denitrification, the samples were incubated for 4 h in darkness at 20°C. Water samples for analysis of nitrous oxide (16 mL) and nitrate (14 mL) were taken before and after the incubation. After the experiments the content of each cylinder was dried at 105°C and weighted.
Nitrification rates were measured under aerobic conditions using the isotope-dilution technique (Koike and Hattori, 1978). Nitrogen-15–nitrate was added to the aquatic phase of the samples. Nitrifying bacteria diluted the 15N-nitrate pool by oxidizing 14N-ammonium to 14N-nitrate, and this dilution was considered to be proportional to nitrification. The 14N- and 15N-nitrate were measured according to Davidsson et al. (1997) after converting nitrate to N2 gas by use of a denitrifying culture (Pseudomonas nautica); 28N2, 29N2 and 30N2 were analyzed using a mass spectrometer (Hewlett-Packard MS engine quadropole mass spectrometer).
Denitrification rates were measured under anaerobic conditions using the acetylene-inhibition technique (Balderston et al., 1976). This technique was used instead of only measuring nitrate decrease to avoid overestimation of denitrification because of nitrate consumption by assimilative uptake and dissimilative reduction of nitrate to ammonium. At the start of an experiment, after initial sampling, 10% of the water phase of the samples was replaced by water saturated with acetylene, and the cylinders were closed with stoppers. The production of nitrous oxide was measured using a gas chromatograph (Varian 3300) equipped with a Porapak Q column (oven temperature 50°C) and a 63Ni electron capture detector (310°C). The carrier gas consisted of 10% CH4 and 90% Ar. Henry's Law was used to calculate the N2O content in the cylinders.
Estimations of Nitrification and Denitrification per Biofilm Surface Area
Nitrification and denitrification were expressed per surface area of the twigs, submersed macrophytes, and sediment. The activities on the filamentous macro-algae were expressed per gram dry weight because it was not possible to estimate their surface area. The biofilm area on the twigs was calculated on the basis of length and diameter, assuming that the twigs were cylindrical. The biofilm area of Eurasian watermilfoil, i.e., surface area of the stems, was calculated from dry weight using the conversion factor 1205 cm2 g-1 dry wt. (Sher-Kaul et al., 1995). For comparison with the activities on the submersed macro-algae, the activities on the macrophytes were expressed per gram dry weight. For sediment, the biofilm area was considered equal to wetland area; surface area of sediment particles and pore spaces were not accounted for. Depth variations were not considered as the process rates were calculated on an areal basis.
Statistics
Data on potential nitrification and denitrification were log-transformed to obtain normal distributions and analyzed by two-way Analysis of Variance (ANOVA) (season and wetland basin were used as fixed variables) followed by Tukey's post-hoc test. Differences were accepted as significant at the P = 0.05 level. A correlation analysis between nitrous oxide production and the decrease in nitrate during the denitrification incubations was conducted.
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RESULTS
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Water Chemistry and Water Flow
The hydraulic loading to Basin B was 13 to 35% lower than that to Basin D during both sampling periods (Table 1). The volumes of the basins differed between July and November due to differential adjustment of the weirs. The N concentrations were higher in water entering Basin B than Basin D (Table 2). Basins B and D had similar total N transformation rates, calculated as the difference between the inlet and outlet mass of total N, although the values were lower in November than in July (Table 1). The water in Basins B and D had similar DOC concentrations, 11.7 and 13.0 mg L-1, respectively. The water flowing into Basin D had higher oxygen concentrations (Fig. 2)
and pH (Table 3), and the oxygen concentrations increased throughout the basins in summer and autumn.
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Table 1. Hydraulic load, total wetland area (Atot), total basin volume, ammonium and total N loads, and ammonium and total N transformation rates in Basins B and D based on inflow, precipitation, evapotranspiration, and total wetland area.
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Fig. 2. Oxygen concentrations in Basins B and D in July (black bars) and November (white bars). The data represent oxygen concentrations in the water at the inlets of the two basins (B1 and D1), at the outlets of the first (B2 and D2) and second (B3 and D3) sections, and at the outlet of the third section (B4 and D4).
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Nitrification and Denitrification
The potential nitrification per biofilm area was more than three times higher for twigs than for sediment samples and about 100 times higher for twigs than for shoots of Eurasian watermilfoil (Fig. 3A)
(n = 5, p < 0.0005). There was no significant difference between Eurasian watermilfoil and the macro-algae in regard to potential nitrification rates (Fig. 3B) (n = 5). All surfaces showed similar nitrification rates in July between Basins B and D. However, in November, nitrification rates in sediments were higher in Basin D than in Basin B (n = 5, p = 0.013).
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Fig. 3. Potential nitrification (black bars), potential denitrification (striped bars), and nitrate removal (white bars) on the surfaces of Eurasian watermilfoil, twigs, sediment, and macro-algae (biofilm area-1 in A; dry wt.-1 in B). The samples were collected in Basins B and D (n = 5). Error bars show SD.
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The potential denitrification rates were about 40 times higher in the sediment than on Eurasian watermilfoil and three times higher in the sediment than on the twigs (Fig. 3A) (n = 5, p < 0.0005). There was no significant difference between the potential denitrification rates for Eurasian watermilfoil and the macro-algae. Moreover, there was no significant difference between the two basins in regard to potential denitrification rates on the different types of surfaces. The potential nitrification and denitrification rates in July and November were similar for twigs and sediment samples (Fig. 3A). During incubation of the samples, nitrate loss was larger than the N2O-production. However, the correlation between the two measurements was strong (r2 = 0.76) and nitrate removal was significantly different among the surfaces (n = 5, p < 0.0005) (Fig. 3A).
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DISCUSSION
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Nitrification
Wetland surfaces other than the sediment proved to be important sites for colonization by nitrifying bacteria. Earlier studies have pointed out submersed macrophytes and their epiphytic algae as important sites for nitrification (Eriksson and Weisner, 1999; Körner, 1999). The nitrifying communities of submersed plants are situated in the water column, surrounded by flowing water supplying ammonium and oxygen, which favor the nitrifying bacteria. Competition by heterotrophic bacteria is also likely to be less on other surfaces compared with the sediment. The results of this study showed the highest potential nitrification in biofilms of twigs, with rates that were about 100 times greater than those noted for Eurasian watermilfoil, despite the similar chemical and physical environments of these two surfaces. Also, the sediment showed higher nitrifying capacity than Eurasian watermilfoil. Interestingly, Eriksson and Andersson (1999) have reported values representing nitrification in litter of emergent macrophytes that are also much higher than the corresponding values for the biofilms on the living submersed plants in the present study (comparison based on plant dry wt.). The nitrifying capacity of Eurasian watermilfoil shoots recorded in the current investigation was about 3 to 10 times lower than that seen in sago pondweed shoots in another study of a wastewater treatment wetland (Eriksson, 2001). The ammonium concentrations were higher in the other wetland (10–25 mg NH4–N L-1) than in the wetland of the present study, which probably favored the nitrifying bacteria. The varying nitrification capacity between twigs, litter, and living submersed plants is probably due to differences in the characteristics of the surfaces.
Denitrification
Denitrification was measured as the production of nitrous oxide, which, in general, was found to be lower than the decrease of nitrate in the incubations. Measurements of changes in nitrate concentrations include nitrate consumption by assimilative uptake and dissimilative reduction of nitrate to ammonium, and are therefore not directly comparable to measurements of nitrous oxide production, which only include denitrification. Differences between the nitrate loss and production of nitrous oxide could also occur if denitrification were underestimated by the acetylene inhibition technique, which has been shown in some earlier studies (Seitzinger et al., 1993). However, the nitrate loss and the nitrous oxide production showed the same relation among the surfaces, which strongly supports the results. The variation between the replicates in this study was large, probably due to the heterogeneous environment of the wetland that is affected by water patterns, vegetation, and fauna. In spite of these small-scale factors influencing the wetland conditions, the differences between the surfaces were large.
The sediment was the most important surface for denitrifying bacteria, and the measured rates were higher or similar to potential denitrification activities noted in other wastewater treatment wetlands (Gale et al., 1993; Kozub and Liehr, 1999; Thompson et al., 2000). Moreover, other investigators have shown that the sediment is of greater importance than submersed plants as a site for denitrifying bacteria (Körner, 1999; Eriksson, 2001). The population of denitrifying bacteria can be regulated by bioavailable organic C and supply of electron acceptors, that is, oxygen and nitrate concentrations (Tiedje, 1988). Since the largest amounts of organic material in a wetland are usually found in the sediment, that location will probably be the best place for growth of denitrifying bacteria. Potential denitrification rates did not differ between the sediments in Basins B and D, even though there were more submersed plants in Basin D. These plants could be assumed to contribute large amounts of organic material to the sediment, but the organic content was found to be similar in the sediments of both basins (loss on ignition 70 ± 10 SD % of dry wt.). This was probably due to regular harvesting of plants in Basin D, leaving only a minor part of the organic matter as a source of C and energy for the denitrifying bacteria.
The twigs showed higher potential denitrification than Eurasian watermilfoil or algae, which might also be the effect of the availability of organic material. After 3 yr, decaying twigs may constitute a readily available source of energy and C. In support of the latter hypothesis, Eriksson and Weisner (1996) found substantially higher denitrification activity on older shoots than on fresh ones, and they attributed this difference to greater leakage of available organic C from the older shoots as they were starting to decay.
The denitrifying bacteria can also be limited by the availability of electron acceptors. The denitrifiers are facultative anaerobic bacteria and both oxygen and nitrate could be used as electron acceptors and support a population of denitrifying bacteria. It has been reported that aquatic systems receiving lower nitrate concentrations show much lower denitrification potential (Sørensen et al., 1988; Eriksson and Weisner, 1996). In the wetland examined in the current investigation, both oxygen and nitrate concentrations during the study period were high, which would be able to support a large population of denitrifying bacteria. Furthermore, oxygen and nitrate concentrations varied between the basins and were likely to be lower at the sediment than at the surfaces in the water column. However, the denitrification potential of the surfaces of different sites in the wetland were not reflected by the field nitrate and oxygen concentrations of that site. This suggests that the population size of denitrifying bacteria on the different surfaces in this wetland was regulated primarily by the availability of organic material.
The Importance of Different Surfaces for the Wetland Nitrogen Removal
Knowledge about the type of surfaces in a wetland that are preferred by nitrifying and denitrifying bacteria is needed to improve these systems in regard to nitrogen removal. Such information can also be used to explain spatial differences in the nitrogen transformation processes within and between wetlands. In the present study, potential nitrification and denitrification rates were measured in relation to biofilm area of a surface. Additionally, to discuss the importance of these surfaces on a wetland scale, we estimated the potential process rates in relation to wetland area in the D basin.
Nitrification is often the limiting process in treatment wetlands (Hammer, 1992). In the present study, the field ammonium transformation rate (Table 1) was in the same range as the estimated potential nitrification in Basin D (Fig. 4)
. However, the field ammonium transformation rate also includes ammonium uptake by plants (Kadlec and Knight, 1996). The potential process rates per wetland surface area were estimated using biomass values of macro-algae and Eurasian watermilfoil, and considering activities per sediment surface area to be similar to activities per wetland surface area, as we collected intact sediment cores. As the nitrifying capacity of the twigs was high in relation to biofilm area, the field nitrification rates in wetland systems with high concentrations of nutrients and oxygen can be considerably increased by the presence of structures, such as twigs. The estimated potential rate of denitrification in relation to the wetland sediment area was much higher than the total N removal rate in the field. In November, this could be explained by lower temperature in the field, but not in July, when the temperature in the field was high. The low denitrification in the field during summer could be because the basins are shallow (about 1 m) and open, and thus contain high oxygen concentrations (Fig. 2) that restrict the denitrifying activity to deeper areas in the sediments (Nielsen et al., 1990). Similar to other studies, this makes nitrate diffusion critical for the active denitrifying zone (Jensen et al., 1994; Martin and Reddy, 1997).
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Fig. 4. Potential nitrification (black bars) and denitrification (white bars) on the surfaces in the D basin during July estimated per wetland surface area, using maximum and average biomass values for macro-algae and Eurasian watermilfoil, and surface area for sediment. Error bars show SD.
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CONCLUSIONS
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The nitrifying population was favored by physical structures, such as twigs, and by high nutrient and oxygen concentrations. As the nitrifying capacity of the twigs was high, the field N removal rates in wetland systems with high concentrations of nutrients and oxygen can be considerably increased by the presence of structures, such as twigs. The distribution of denitrifiers was most likely regulated by the availability of organic material, with higher denitrification rates in the sediment than on surfaces in the water column. In the field, denitrification may be restricted to the deeper sediments due to high oxygen concentrations.
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ACKNOWLEDGMENTS
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The authors acknowledge Per-Åke Nilsson, Hässleholm Municipality, Lena Karlsson, Mauricio Petrucio, and Jesper Persson for providing data and helpful information during this study. We thank David Bastviken and three anonymous referees for valuable comments on the manuscript. The study was conducted as part of the Swedish Water Management Research Program (VASTRA), which is financed by the Swedish Foundation for Strategic Environmental Research (MISTRA).
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ABSTRACT
INTRODUCTION
