Ecological Engineering 25 (2005) 528–541 Emission of N2O, N2, CH4, and CO2 from constructed wetlands for wastewater treatment and from riparian buffer zones Sille Teiter, Ülo Mander ∗ Institute of Geography, University of Tartu, 46 Vanemuise St., 51014 Tartu, Estonia Received 1 March 2005; accepted 11 July 2005 Abstract We measured nitrous oxide (N2 O), dinitrogen (N2 ), methane (CH4 ), and carbon dioxide (CO2 ) fluxes in horizontal and vertical flow constructed wetlands (CW) and in a riparian alder stand in southern Estonia using the closed chamber method in the period from October 2001 to November 2003. The replicates’ average values of N2 O, N2 , CH4 and CO2 fluxes from the riparian gray alder stand varied from −0.4 to 58 ␮g N2 O-N m−2 h−1 , 0.02–17.4 mg N2 -N m−2 h−1 , 0.1–265 ␮g CH4 -C m−2 h−1 and 55–61 mg CO2 -C m−2 h−1 , respectively. In horizontal subsurface flow (HSSF) beds of CWs, the average N2 emission varied from 0.17 to 130 and from 0.33 to 119 mg N2 -N m−2 h−1 in the vertical subsurface flow (VSSF) beds. The average N2 O-N emission from the microsites above the inflow pipes of the HSSF CWs was 6.4–31 ␮g N2 O-N m−2 h−1 , whereas the outflow microsites emitted 2.4–8 ␮g N2 O-N m−2 h−1 . In VSSF beds, the same value was 35.6–44.7 ␮g N2 O-N m−2 h−1 . The average CH4 emission from the inflow and outflow microsites in the HSSF CWs differed significantly, ranging from 640 to 9715 and from 30 to 770 ␮g CH4 -C m−2 h−1 , respectively. The average CO2 emission was somewhat higher in VSSF beds (140–291 mg CO2 C m−2 h−1 ) and at the inflow microsites of HSSF beds (61–140 mg CO2 -C m−2 h−1 ). The global warming potential (GWP) from N2 O and CH4 was comparatively high in both types of CWs (4.8 ± 9.8 and 6.8 ± 16.2 t CO2 eq ha−1 a−1 in the HSSF CW 6.5 ± 13.0 and 5.3 ± 24.7 t CO2 eq ha−1 a−1 in the hybrid CW, respectively). The GWP of the riparian alder forest from both N2 O and CH4 was relatively low (0.4 ± 1.0 and 0.1 ± 0.30 t CO2 eq ha−1 a−1 , respectively), whereas the CO2 -C flux was remarkable (3.5 ± 3.7 t ha−1 a−1 ). The global influence of CWs is not significant. Even if all global domestic wastewater were treated by wetlands, their share of the trace gas emission budget would be less than 1%. © 2005 Elsevier B.V. All rights reserved. Keywords: Carbon dioxide; Constructed wetland; Dinitrogen; Global warming potential; Methane; Nitrous oxide 1. Introduction Constructed wetlands (CW) for wastewater treatment and riparian buffer zones are important ecotech∗ Corresponding author. Tel.: +372 7 375819; fax: +372 7 375825. E-mail address: [email protected] (Ü. Mander). 0925-8574/$ – see front matter © 2005 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2005.07.011 nological measures for controlling water quality in agricultural catchments (Peterjohn and Correll, 1984; Kadlec and Knight, 1996; Kuusemets and Mander, 1999). Denitrification, which is generally referred to as the microbial reduction of NO3 − -N to NO2 − -N and further to gaseous forms NO, N2 O and N2 (Knowles, 1982), has been found in numerous studies to be a sig- S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 nificant process in nitrogen removal in riparian buffer zones (Cooper, 1990; Ambus and Lowrance, 1991; Groffman et al., 1991; Lowrance, 1992; Ambus and Christensen, 1993; Haycock and Pinay, 1993; Schipper et al., 1993; Hanson et al., 1994; Weller et al., 1994; Nelson et al., 1995; Hill, 1996; Schnabel et al., 1997; Gold et al., 1998; Hefting and de Klein, 1998; Mogge et al., 1998; Groffman et al., 2000). In the majority of these studies, nitrous oxide (N2 O) fluxes have been measured, and only a few studies devote attention to dinitrogen (molecular nitrogen; N2 ) emission (Blicher-Mathiesen et al., 1998; Watts and Seitzinger, 2000). N2 O, as one of the greenhouse gases, is increasing in the atmosphere at a rate of about 0.3% year−1 (Mosier, 1998). It has an atmospheric lifetime of about 120 years, a global warming potential of 296 relative to CO2 over a 100 year time horizon, and is responsible for about 6% of anticipated warming (IPCC, 2001). Riparian zones have the potential to be hotspots of N2 O production in the landscape (Groffman et al., 2000). Numerous studies consider emissions and sequestration of carbon dioxide (CO2 ) in wetlands (Mitsch and Gosselink, 1993; Funk et al., 1994; Hamilton et al., 1995; Lafleur et al., 1997; Joiner et al., 1999; Griffis et al., 2000; Christensen et al., 2003). Depending on meteorological and hydrological conditions, wetlands can be sources or sinks of carbon (Clark et al., 1999; Waddington and Roulet, 2000; Whiting and Chanton, 2001; Arneth et al., 2002). Likewise, riparian wetlands and wet riparian forests can be sources of CO2 (Jones and Mulholland, 1998a; Scott et al., 2004) and methane (CH4 ) (Jones and Mulholland, 1998b; Rusch and Rennenberg, 1998; Gulledge and Schimel, 2000; Rask et al., 2002), which is another greenhouse gas increasing in the atmosphere at the rate of about 0.8% year−1 (Mosier, 1998). Methane in the atmosphere has a lifetime of 8.4 years. On a 100 year time horizon, CH4 has a global warming potential of 23 relative to CO2 , and is responsible for about 20% of anticipated warming (IPCC, 2001). Both denitrification and methane formation depend on the oxygen status of the soil or sediment. In this relation, the spatial and temporal variability of fluxes of both N2 O (Robertson and Tiedje, 1984; Struwe and Kjøller, 1990; Ambus and Christensen, 1993; Pinay et al., 1993; Brooks et al., 1997; Augustin et al., 1998b; Gold et al., 1998; Jacinthe et al., 1998) and CH4 (Saarnio et al., 1997; Willison et al., 1998; Reay et al., 2001) is extremely high. In addition to oxy- 529 gen status, denitrification rates in soils are influenced by carbon availability, nitrate availability, temperature, and pH (Erich et al., 1984). Biological methane oxidation is an important environmental process preventing the release into the atmosphere of much of the CH4 produced in anoxic soils and sediments. Well-drained soil acts as a sink for atmospheric CH4 due to methane oxidation (negative emission), through either ammonia oxidizers or methanotrophs (Hanson et al., 1994). In contrast to riparian buffer zones and natural wetlands, far fewer studies have been carried out on N2 O and CH4 fluxes from CWs for wastewater treatment. Most of the data are available on the contribution of free water surface constructed wetlands to N2 O (Lund, 1999; Xue et al., 1999; Bachand and Horne, 2000; Lund et al., 2000; Spieles and Mitsch, 2000; Wild et al., 2002; Johansson et al., 2003) and CH4 (Tanner et al., 1997; Tai et al., 2002; Wild et al., 2002) emissions. Only two works (Fey et al., 1999; Tanner et al., 2002) considered the nitrous oxide fluxes from subsurface flow constructed wetlands, and only one paper considers dinitrogen emission from a CW (Mander et al., 2003). Surprisingly, we could find no published materials on CO2 emissions from constructed wetlands. The main objectives of this research were: (1) to quantify N2 O, N2 , CH4 and CO2 emission rates from two subsurface flow CWs for municipal wastewater treatment and in a grey alder stand in Estonia using the closed chamber method and (2) to compare N2 O, N2 , CH4 , and CO2 fluxes and their global warming potential (GWP) from riparian buffer zones and constructed wetlands. 2. Methods 2.1. Site description A description of the Kodijärve horizontal subsurface flow (HSSF) planted sand filter (constructed in October 1996, purifies the wastewater from a hospital for about 40 population equivalents (PE); Fig. 1A) is given by Mander et al. (2001, 2003). The hybrid treatment wetland system in Kõo, Viljandi County, Estonia, consists of a two-bed vertical subsurface flow (VSSF) filter (2 m × 64 m, filled with 5–10 mm crushed limestone, planted with Phragmites australis), a HSSF filter (365 m2 , filled with 15–20 mm 530 S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 Fig. 1. Diagrams of study sites: (A) horizontal subsurface flow planted sand filter (HSSF) in Kodijärve, (B) hybrid wetland system in Kõo and (C) riparian study area in Porijõgi. In part A: plant names in italics indicate the present dominant species in the beds, (M) automatic weather station. In part B: (1) pumping station; (2) septic tank; (3) vertical subsurface flow filter (VSSF; 2 × 64 m2 ), (a) right part, (b) left part; (4) HSSF (365 m2 ), (a) left inflow, (b) right inflow, (c) outflow; (5) 1st free-water surface wetland (FWSW; 3600 m2 ), (6) 2nd FWSW (5500 m2 ), (7) polishing pond (500 m2 ). S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 crushed limestone, planted with Typha latifolia and P. australis), and two free water surface wetland (FWSW) beds (3600 and 5500 m2 , planted with T. latifolia; Fig. 1B). The system was constructed in 2000 for the purification of the raw municipal wastewater generated by about 300 PE. This wetland system showed a good purification, being for BOD7, total-N and total-P 88, 65 and 72%, respectively. Regarding the very high average loading rates (96, 34 and 4.7 g m−2 for BOD7 , total N and total P, respectively), however, the purification efficiency of the VSSF and HSSF part was only 57 and 31% for BOD7 , 16 and 21% for total N, and 19 and 21% for total P, correspondingly. The Porijõgi riparian buffer zone site is a grey alder stand situated in the moraine plain of southeast Estonia (Tartu County, Sirvaku; 58◦ 13′ N, 26◦ 47′ E) on the right bank of a small river, the Porijõgi, which flows in a primeval valley where agricultural activities ceased in 1992. The landscape study transect in this valley crosses several plant communities: an abandoned field (last cultivated in 1992) on planosols and podzoluvisols; an abandoned cultivated grassland (last mowed in 1993) on colluvial podzoluvisol (dominated by Dactylis glomerata and Alopecurus pratensis); an 11m-wide wet grassland on gleysol (two parallel communities, one dominated by Filipendula ulmaria, another by Aegopodium podagraria); and a 20-m-wide grey alder stand (Alnus incana) on gleysol (Fig. 1C). For a more detailed description, see Kuusemets et al. (2001). The mean annual air temperature at the study sites varied from 5.0 to 5.5 ◦ C. In winter the lowest daily mean temperatures reach −20 ◦ C. The variation in long-term annual precipitation is 500–700 mm. 2.2. Sampling and laboratory analysis For the measurement of N2 O, N2 , CH4 and CO2 , two emission methods—the “closed chamber” (closed soil cover box) method (Denmead and Raupach, 1993; Hutchinson and Livingston, 1993) and the helium–oxygen (He–O) method (Butterbach-Bahl et al., 1997; Scholefield et al., 1997; Mander et al., 2003) were used. The latter was used especially for the measurement of N2 fluxes. Gas samplers (closed chambers; cover made from PVC, height 50 cm, Ø 50 cm, volume 65 l, sealed with a water-filled ring on the soil surface, painted white to avoid heating during application) were installed in five replicates in various parts 531 of the studied systems: (1) on the inlet and outlet pipes of both beds, in Kodijärve (Fig. 1A) and (2) in three different microsites (EDGE, WET and DRY; range of water table depth 45–95, 0–50 and 45–95 cm, respectively) in the Porijõgi riparian buffer zone (Fig. 1C). In the hybrid wetland system in Kõo, 8 gas samplers were installed in the vertical flow filter (4 in each bed), and 15 in the horizontal flow filter (5 on two inlet pipes and 5 on the outlet pipe; Fig. 1B). At the end of the 1 h measuring time, gas samples were taken from the enclosures of the samplers, using previously evacuated gas bottles (100 ml; see Augustin et al., 1998b). Gas sampling was carried out 15 times on the following time schedule: once a month in October and November 2001, and in March, May to December 2002, January to March, July and November 2003. Simultaneously, the soil temperature and water depth in the sampling wells was measured, and the NH4 -N and NO3 -N concentration in soil samples was analysed using the Kjeldahl method (APHA, 1989). The trace gas concentration in the collected air was determined using the gas chromatography system (electron capture detector and flame ionization detector; Loftfield et al., 1997) in the lab of the Institute of Primary Production and Microbial Ecology, Centre for Agricultural Landscape and Land Use Research (ZALF), Germany. The trace gas flux rates were calculated according to Hutchinson and Livingston (1993) from a linear change in trace gas concentration over time with reference to the internal volume of the chamber and the soil area covered. Soil temperature and groundwater tables were measured simultaneously (Augustin et al., 1998a). Intact soil cores (diameter 6.8 cm, height 6 cm) for use with the He–O method were sampled from the topsoil (0–10 cm) at gas sampler (closed chamber) sites, after gas sampling was completed, in the following order: in October and November 2001, March, June to August, and October 2002, and January to March 2003. In following text, months from November to April have been considered as “winter”, and months from May to October as “summer”. Soil samples were weighted, kept at low temperature (4 ◦ C), and transported to the ZALF laboratory. At the lab, intact soil cores were introduced into special gas-tight incubation vessels. In these vessels, N2 was removed by three subsequent slight evacuation/flushing cycles with an artificial gas mixture (21.3% O2 , 78.6% He, 337 ppm CO2 , 374 ppb N2 O, 1882 ppb CH4 and approximately 5 ppm 532 S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 N2 ). This was followed by the establishing of a new flow equilibrium by continuously flushing the vessel headspace with the artificial gas mixture at 10 ml/min for 12 h. For the start value, N2 and the greenhouse gas concentration in the continuous gas flow was measured. The measuring of the gas concentrations in the incubation headspace (final value) followed after closing the incubation headspace for one hour to accumulate the emission of N2 and the greenhouse gases. The final accumulation value minus the start continuous flow value served as the basis for the calculation of the emission rates. The procedures used for the determination of the actual gas emission rate are described by Mander et al. (2003). In Kodijärve, we also measured the water table once a month and took samples from 18 sampling wells and the inflows and outflows of both beds for further analyses for BOD7 , NH4 -N, NO2 -N, NO3 -N, total N, PO4 -P and total P in the lab of Tartu Environmental Research Ltd. (APHA, 1989). Water discharge was measured using tipping buckets installed in the inlet and outlet wells. Air and soil temperature, wind velocity, solar radiation and precipitation evapotranspiration were measured using a DAVIS Groweather automatic weather station installed close to the CW (Fig. 1A). 2.3. Statistical analysis The normality of variables was checked using the Kolmogorov-Smirnov, Lilliefors, and Shapiro-Wilk tests. In most cases of gas analyses the distribution differed from the normal, and hence non-parametric tests were performed. We used the Duncan Test, Wilcoxon Matched Pairs Test and the Mann–Whitney U-test to check the significance (α = 0.05 was accepted in all cases) of differences between the gas emission rates at different times and sites. The Spearman Rank Order Correlation was performed to analyse correlations between gaseous fluxes and environmental parameters. 3. Results and discussion 3.1. Temporal variation of gas emissions According to the Duncan test, a significantly higher release of all gases from CWs was observed during the warmer period (Fig. 2A–D), although the N2 O flux showed no significant correlation with air and water temperature. However, mean topsoil temperatures (varied from 0.1 to 20.5 ◦ C) correlated significantly with the emission rates of all analysed gases (R2 values for N2 O-N, N2 -N, CH4 -C and CO2 -C are 0.32, 0.56, 0.50 and 0.54, respectively. In the HSSF wetlands, the season-dependence of CH4 emission was extremely remarkable. It resulted in significant differences in average values of CH4 fluxes from both the HSSF CW in Kodijärve and the hybrid CW in Kõo in summer (5000–21900 and 1700–14400 ␮g CH4 -C m−2 h−1 in Kodijärve and Kõo, respectively) and winter (24–300 and 16–2000 ␮g CH4 -C m−2 h−1 , respectively; Fig. 2B). The very cold winter of 2002/2003 with air temperatures from −15 to −25 ◦ C for almost 2 months apparently influenced both water purification efficiency (Noorvee et al., 2005) and gas emissions. As with purification performance, gaseous emission was significantly lower in spring and early summer than in autumn. In the riparian grey alder stand, only the CO2 emission varied in accordance to variations of water and air temperature. The average CO2 emission varied from 13.6 ± 11.3 mg CO2 -C m−2 h−1 in January to 187.8 ± 56.3 mg CO2 -C m−2 h−1 in August. The emission of N2 O from the riparian zone showed the highest values in January and March 2003 (up to 180 ␮g N2 ON m−2 h−1 from the WET microsite), remaining relatively low during the rest of the study period (from −3.3 to 24 ␮g N2 O-N m−2 h−1 ; Fig. 2A). Likewise, the results of some other investigations demonstrate that N2 O emission does not clearly depend on soil temperature, and the release of this gas from the soil in cold periods can be as high or even higher in winter as in summer (Augustin et al., 1996; Fey et al., 1999). For instance, N2 O-N fluxes through the snowpacks in winter reached 112 ␮g N2 O-N m−2 d−1 (Brooks et al., 1997), which is comparable with the lower emission values from our study sites. In the warm and dry summer of 2002, the N2 O emission from the riparian zone increased significantly with the lowering water table level (Spearman R = 0.38). This effect has been noted in several studies on natural wetlands (Martikainen et al., 1993; Dowrick et al., 1999). The average CH4 emission from the riparian alder stand varied from 0.1–29 to 1.2–265 ␮g CH4 -C m−2 h−1 in winter and summer, respectively (Fig. 2B). In our riparian study area, the emission of CH4 in the snow-covered period is signifi- S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 Fig. 2. Temporal variation of emission rates of nitrous oxide (A), dinitrogen (B), methane (C) and carbon dioxide (D; average ± S.D.) from the Kodijärve HSSF CW, Kõo hybrid wetland system and the Porijõgi riparian grey alder stand, averaged over all sampling sites. For better visualization, polynomial curves are added. Hidden values in part C: (1) 21890 ± 43570; (2) 18110; (3) 27425; (4) 14020 ± 17920; (5) 1030; (6) 14410 ± 14290; (7) 17570. 533 534 S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 cantly less than that reported by Wickland et al. (1999) for subalpine wetland sites in the Rocky Mountains (23–73% of the annual fluxes). The average dinitrogen flux from the microsites in Kodijärve was 2–3 magnitudes higher than the N2 O flux, ranging from 19.5 to 33.3 mg N2 -N m−2 h−1 . In the Kõo hybrid CW and in the Porijõgi riparian alder stand, the difference between the N2 and N2 O fluxes was 20–750 and 150–700 times, respectively. In Kõo, the variation of N2 emission was from 0.3 to 99, and in Kodijärve, from 0.6 to 17 mg N2 -N m−2 h−1 (Fig. 2A and B). In our study, CO2 emission is not connected with fluxes related to plant photosynthesis. Therefore, only data for cold periods can be considered as losses to the atmosphere. For calculating the net ecosystem CO2 exchange, a more advanced measurement technique is required. For instance, the eddy covariation technique allows the analysis of full C balance in ecosystems (Shurpali et al., 1993; Kormann et al., 2001). However, some studies on C sequestration in wetlands and forest ecosystems (Butnor et al., 2003) allow one to estimate that about 50% of CO2 , released during soil respiration in the vegetation period cycles back to the atmosphere. It is important to take this into consideration when calculating the GWP of CWs and riparian buffer ecosystems. 3.2. Spatial variation of gas emissions The average flux of nitrous oxide from the microsites in the Kodijärve HSSF CW and Kõo hybrid CW ranged from 27 to 370 and from 72 to 500 ␮g N2 O-N m−2 h−1 , respectively (Fig. 3A). In Kodijärve, according to the Wilcoxon Matched Pairs Test, significant differences were found in average N2 O fluxes between the microsites: 325–350 ␮g N2 ON m−2 h−1 from chambers installed above the inflow pipes and 30–40 ␮g N2 O-N m−2 h−1 from chambers above the outflow pipes. In Kõo, the VSSF beds emitted more nitrous oxide than the HSSF bed (405–510 and 70–165 ␮g N2 O-N m−2 h−1 , respectively), although Fig. 3. Emission rates of nitrous oxide (A), methane (B), carbon dioxide (C) and dinitrogen (D; average ± S.D.) from sampling sites in the Kodijärve HSSF CW, Kõo hybrid wetland system and Porijõgi riparian grey alder forest. *: Significantly differing value (p < 0.05) with at least two other microsites according to the Wilcoxon Matched Pairs Test. For the locations of sampling sites, see Fig. 1. S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 the differences were not significant. This kind of difference is probably due to higher initial load in the VSSF bed (see Section 2.1). In Kodijärve, differences in individual values of N2 emission from replicate soil cores varied greatly, from 170 to 130,000 ␮g N2 -N m−2 h−1 , but the variations were statistically non-significant (Fig. 3B). In contrast to the N2 emission, we found significant differences in average N2 O fluxes between the microsites: about 320–350 ␮g N2 O-N m−2 h−1 from chambers installed above the inflow pipes and <50 ␮g N2 O-N m−2 h−1 from chambers above the outflow pipes (Fig. 3A). In Kõo, a significantly lower N2 emission from the microsite above the outlow pipe of the HSSF bed was found (Fig. 3B). In the Porijõgi riparian forest, the WET microsite emitted significantly more N2 O than the EDGE and DRY microsites (average values 30, 9 and 8 ␮g N2 ON m−2 h−1 , respectively; Fig. 3A). On the other hand, the decreasing water table level in the WET microsite in summer 2002 caused a significant increase in N2 O emission. CH4 fluxes showed great variability in both the Kodijärve and Porijõgi study areas: significantly lower emission was observed from all the microsites in the Porijõgi riparian alder forest and from the chambers installed above the outflow pipes in the Kodijärve HSSF filter. The average methane emission from the microsites in the Kodijärve HSSF CW and the Kõo hybrid CW ranged from 30 to 9715 and from 770 to 5540 ␮g CH4 -C m−2 h−1 , respectively (Fig. 3C). These values are about 2 magnitudes higher than found in natural boreal wetlands (MacDonald et al., 1998), two to three times higher than reported on re-flooded fens (Augustin et al., 1996) or constructed cat-tail wetlands (Wild et al., 2002), but 2.53 times lower than on fertilised wet grasslands on peat soils in the Netherlands (Van den Pol-Van Dasselaar et al., 1999), up to four times lower than on floodplain wetlands (Boon et al., 1997), and up to five times lower than observed in FWSW-s for wastewater treatment (Tai et al., 2002). With respect to the differences in CH4 flux between the microsites, we found quite a similar pattern with nitrous oxide fluxes. According to the Wilcoxon Matched Pairs Test, significantly more methane was released from the microsites situated above the inflow pipes of both HSSF CWs (up to 9720 ␮g CH4 -C m−2 h−1 in Kodijärve and 4630–5540 ␮g CH4 -C m−2 h−1 in Kõo) than from the 535 microsites above the outlet pipes (30–125 ␮g CH4 C m−2 h−1 ; Fig. 3C). This is consistent with the significant rank correlation between water table depth and CH4 flux (Spearman R = −0.37; Mander et al., 2003). Likewise, this relationship has on many occasions been mentioned in other studies on wetlands (Boon et al., 1997; Cao et al., 1998). Also, methane release is positively influenced by suspended solids, NH4 -N, total N, PO4 -P and total P, and BOD7 levels in wastewater. At the same time, a significant negative rank correlation was found between CH4 flux and the NO2 -N and NO3 -N concentrations in water (Mander et al., 2003). A certain part of methane could be oxidized by transporting oxygen by plants to water saturated filter bed (Kettunen, 2003) or by methanotrophs living on green leaves (see Berestovskaya et al., 2002); however, this influence is probably non-significant (see Brix et al., 2001). In the riparian alder stand, CH4 flux was significantly lower than from the VSSF beds and the inflow microsites of HSSF CWs, varying from 14 to 144 ␮g CH4 -C m−2 h−1 (Fig. 3D). The WET microsite showed significantly higher CH4 emission values than the EDGE and DRY sites (Fig. 1C). In comparison with other gases measured, the CO2 flux from soils showed the lowest spatial variation. Slightly higher CO2 release was found from the microsites in the VSSF beds and above the inflow pipes of HSSF CWs (140–290 and 61–130 mg CO2 C m−2 h−1 , respectively), although these differences were not significant (Fig. 3D). In CWs, a clear relation was observed between the BOD7 value of wastewater and the average CO2 release from the filter material. In both the HSSF and VSSF beds of CWs, the fluxes of measured gases were significantly positively correlated (Spearman R values ranged from 0.38 to 0.58 in Kodijärve, 0.45 to 0.61 in the VSSF beds in Kõo, and from 0.20 to 0.31 in the HSSF beds in Kõo), whereas no significant rank correlation between gas emissions was found within the riparian alder stand (Spearman R = 0.05–0.09). The last finding is probably related to the relatively high carbon storage in this riparian soil (4–5.3%; Mander et al., 1997). In filter beds of CWs, carbon can become limited due to intensive mineralization, which is reflected in the correlation between the gaseous N and C fluxes. Similar trends have been reported by Paludan and Blicher-Mathiesen (1996) for a Danish freshwater wetland where high 536 K-1 K-2 K-3 K-4 K-v-1 K-v-2 K-v-3 K-v-4 K-h-1 K-h-2 K-h-3 K-h-4 P-1 P-2 P-3 P-4 K-1 K-2 K-3 K-4 K-v-1 K-v-2 K-v-3 K-v-4 K-h-1 K-h-2 K-h-3 K-h-4 P-1 1.00 0.58** 1.00 0.38** 0.50** 1.00 0.45** 0.02 0.08 1.00 0.49** 0.30* 0.01 −0.01 1.00 0.20* 0.14 0.08 0.01 0.46** 1.00 0.28* 0.16 −0.08 −0.01 0.45** 0.61** 1.00 0.54** 0.22 0.15 0.23 0.43** 0.28 0.08 1.00 0.59** 0.27** 0.22* −0.03 0.54** 0.20* 0.28* 0.38 1.00 0.24** 0.17* 0.09 −0.01 0.07 0.23* 0.16 0.15 0.20* 1.00 0.34** 0.18* −0.03 −0.06 0.26* 0.31** 0.28** 0.04 0.31** 0.28* 1.00 0.27 −0.02 0.44** 0.23 0.02 −0.23 −0.32 0.47** 0.50** 0.12 0.23 1.00 0.59** 0.02 0.25** −0.04 0.09 −0.18* −0.12 0.05 0.33** 0.09 0.12 0.13 0.20* 0.00 0.13 0.05 0.62** 0.13 0.09 0.11 0.25** 0.09 0.27 −0.07 1.00 0.09 1.00 P-2 P-3 P-4 0.02 0.33* 0.05 0.11 0.02 0.13 −0.07 −0.18 0.26* 0.31 0.22* 0.50** 0.09 0.34* −0.27 0.55** 0.23** 0.41** 0.07 0.15 0.09 0.44** −0.01 0.06 0.05 0.14 0.09 0.09 1.00 0.14 1.00 K: Kodijärve HSSF CW, Kv: VSSF part of the Kõo hybrid CW, Kh: HSSF part of the Kõo hybrid CW, P: Porijõgi riparian grey alder stand, 1: CO2 -C; 2: N2 O-N; 3: CH4 -C; 4: N2 -N. Bold numbers represent relations within the same system. * p < 0.05. ** p < 0.001. S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 Table 1 Spearman rank correlation of gas emissions from various study sites S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 NO3 − loading resulted in an accelerated loss of gaseous C. In the Kodijärve HSSF filter and in both the VSSF and HSSF parts of the Kõo hybrid CW, we found a significant (p < 0.05) Spearman rank correlation between the fluxes of CO2 -C and emissions of N2 O, N2 and CH4 , whereas in most cases the correlation was strong (p < 0.001; Table 1). Likewise, N2 O and CH4 flux was strongly correlated in these systems. On the other hand, no significant correlation was found between the gas emissions within the riparian alder forest. The strong relationship found between the gas fluxes from different study sites indicates the temperature dependence of emissions (e.g. the strong Spearman rank correlation between the CO2 fluxes from all the study sites) or has an occasional character. 3.3. Global warming potential of systems analysed The cumulated emission of all studied gases from CWs varied from 16.3 to 21.9, from 230 to 295 and from 9100 to 9700 kg ha−1 year−1 for N2 O-N, CH4 -C and CO2 -C, respectively. Considering average annual organic material load of 2750 kg BOD7 in Kodijärve and about 3500 kg BOD7 in Kõo, the CO2 emission measured is significantly higher than the potential carbon source entered into the CWs with wastewater. It shows that the CO2 emission can be overestimated and a significant part of CO2 emitted is fixed from the atmosphere by plants. The emission level in Kõo always exceeded the corresponding values in Kodijärve, which is probably due to the relatively high loading of the vertical flow system (only two beds of 64 m2 for about 300 PE). In Kodijärve the nominal loading is only 20–40 PE per 312.5 m−2 . When properly functioning, however, the vertical flow system can have a relatively small area, although this seems to enhance N2 O emission. Regarding CH4 flux, it is crucial to avoid clogging both vertical flow and horizontal flow filters: this might help in the case of a higher N2 flux and correspondingly lower N2 O flux; however, it significantly increases methane emissions. Sometimes such clogging took place in both CWs studied, which probably led to high CH4 emission values (Fig. 3C). The GWP of the studied systems were calculated by converting the fluxes of N2 O and CH4 into CO2 equivalents (eq; IPCC, 2001). In Kodijärve, the average N2 O flux from both beds was quite similar: 537 3.85 ± 7.10 t CO2 eq ha−1 year−1 in the right bed and 3.85 ± 3.85 t CO2 eq ha−1 year−1 in the left bed. Methane flux rates, however, showed significant differences, ranging from 0.69 ± 2.05 t CO2 eq ha−1 year−1 in the right bed to 12.4 ± 26.3 t CO2 eq ha−1 year−1 in the left bed. The differences are caused by higher water table kept in the left (wet) bed (see Mander et al., 2000, 2003). In Kõo, the highest GWP of nitrous oxide was found in the vertical flow beds (12.4 ± 19.5 t CO2 eq ha−1 year−1 ), while the horizontal flow bed showed a high methane flux (9.13 ± 17.9 t CO2 eq ha−1 year−1 ). Both Kodijärve HSSF CW and Kõo hybrid CW emit remarkable amounts of CO2 -C, N2 O-N, and CH4 -C: 6.3 ± 6.4, 4.8 ± 9.8 and 6.8 ± 16.2 t CO2 eq ha−1 year−1 in Kodijärve, and 6.8 ± 20.2, 6.5 ± 13.0, and 5.3 ± 24.7 t CO2 eq ha−1 year−1 in Kõo, respectively (Fig. 4). The cumulated emission of N2 O and CH4 in riparian alder forest in Porijõgi was significantly lower than from the CWs (0.4 ± 1.0 and 0.1 ± 0.30 t CO2 eq ha−1 year−1 , respectively), whereas the CO2 -C flux was remarkable (3.5 ± 3.7 t ha−1 year−1 ). When comparing the greenhouse potential of CH4 and N2 O over a long time scale (100–500 years), one can speculate that due to the short adjustment time for CH4 in the atmosphere (8.4 years; IPCC, 2001), the radiative forcing of CH4 will fall relative to CO2 Fig. 4. Cumulated flux rates of major greenhouse gases from Kodijärve and Kõo constructed wetlands and the Porijõgi riparian study site from October 2001 to November 2003, presented as CO2 equivalent values (kg CO2 -C ha−1 year−1 ; mean ± S.D.). The conversion of the flux rates into CO2 equivalents is given with 296 for N2 O and 23 for CH4 , over a time horizon of 100 years (IPCC, 2001). 538 S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 (Whiting and Chanton, 2001). N2 O, with its atmospheric lifespan of about 120 years and GWP value of 296, however, has less expectable impact. Therefore, further investigations should concentrate on the factors that regulate N2 O and N2 emission rates from constructed wetlands. 4. Conclusions The emission of nitrous oxide and methane from CWs showed higher values than those from the riparian gray alder forest, whereas carbon dioxide fluxes did not differ significantly. In CWs we found a remarkable variability in the average flux rates of N2 O-N, CH4 -C and CO2 -C, ranging from 1 to 2600 ␮g m−2 h−1 , −1.7 to 87200 ␮g m−2 h−1 and −6.1 to 1050 mg m−2 h−1 , respectively. In the riparian grey alder forest these values were −3.3 to 190, −5.9 to 805 and −3.9 to 290 mg m−2 h−1 , respectively. Dinitrogen emission was higher from HSSF than from VSSF beds, whereas in the riparian buffer zone the average N2 emission was lower than in CWs. Although the CO2 emission being measured high, we can assume that vegetation cover can reduce emissions of CO2 by photosynthesis substantially. The release of all gases studied was significantly higher during the warmer period, although no significant correlation was found between the N2 O flux and soil/water temperature. Similar to the purification performance, gaseous emissions in spring and early summer were significantly lower than in autumn. Apparently the very cold winter of 2002/2003, with air temperature from −15 to −25 ◦ C for almost 2 months, did influence both water purification efficiency and gas emission. The most intensive flux of N2 O and CH4 was observed in chambers installed above the inflow pipes of horizontal flow beds. The vertical flow wetland did emit significantly more N2 O than the horizontal flow beds. Water table increase in the horizontal flow systems may not significantly influence the efficiency of water purification, although it will increase methane emissions by a few magnitudes. Thus it is very important to avoid the clogging of pipes, which is normally guaranteed by the regular cleaning of sediments from the septic tank. Clogging of the filter material may have the same influence on gas emissions as does the clogging of pipes. Therefore, the careful selection of the optimal grain size of the filter material in the construction phase plays a crucial role in the performance of both horizontal and vertical flow wetlands and also in the regulation of trace gas emission from the planted soil/sand filters. Although the cumulated nitrous oxide emission from studied CWs was relatively high (4–43 kg N ha−1 year−1 as averaged over all microsites), the most important release of gaseous nitrogen from these CWs was observed in the form of N2 -N, varying between 300 and 1300 kg N ha−1 year−1 in different microsites being 650 kg N ha−1 year−1 on average. For comparison: in the Kodijärve HSSF filter in 1997–2001, the average initial total N loading was 2410 kg N ha−1 year−1 (Mander et al., 2003). The emission of N2 O and CH4 from constructed wetlands was found to be relatively high, although their global influence is not significant. Even if all global domestic wastewater were treated by wetlands, their share in the trace gas emission budget would be less than 1%. Acknowledgements This study was supported by EU 5 FP RTD project EVK1-2000-00728 “PRocess Based Integrated Management of Constructed and Riverine Wetlands for Optimal Control of Wastewater at Catchment ScalE” (PRIMROSE), the Estonian Science Foundation project No. 5247, and the Target Funding Project No. 0182534s03 of the Ministry of Education and Science, Estonia. We would like to thank Dr. Jürgen Augustin from the Institute of Primary Production and Microbial Ecology of the Centre for Agricultural Landscape and Land Use Research (ZALF), Müncheberg, Germany, for his assistance in gas analysis. References Ambus, P., Christensen, S., 1993. Denitrification variability and control in a riparian fen irrigated with agricultural drainage water. Soil Biol. Biochem. 25 (7), 915–923. Ambus, P., Lowrance, R., 1991. Comparison of denitrification in two riparian soils. Soil Sci. Soc. Am. J. 55 (4), 994–997. APHA, 1989. Standard Methods for the Examination of Water and Waste Water, 17th ed., Washington. S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 Arneth, A., Kurbatova, J., Kolle, O., Shibistova, O.B., Lloyd, J., Vygodskaya, N.N., Schulze, E.-D., 2002. Comparative ecosystem-atmosphere exchange of energy and mass in a European Russian and a central Siberian bog. II. Interseasonal and interannual variability of CO2 fluxes. Tellus B 54 (5), 696– 712. Augustin, J., Merbach, W., Schmidt, W., Reining, E., 1996. Effect of changing temperature and water table on trace gas emission from minerotrophic mires. J. Appl. Bot. 70, 45–51. Augustin, J., Merbach, W., Steffens, L., Snelinski, B., 1998a. Nitrous oxide fluxes of disturbed minerotrophic peatlands. Agribiol. Res. 51 (1), 47–57. Augustin, J., Merbach, W., Rogasik, J., 1998b. Factors influencing nitrous oxide and methane emissions from minerotrophic fens in Northeast Germany. Biol. Fertil. Soils 28 (1), 1–4. Bachand, P.A.M., Horne, A.J., 2000. Denitrification in constructed free-water surface wetlands. I. Very high nitrate removal rates in a macrocosm study. Ecol. Eng. 14 (1–2), 9–15. Berestovskaya, Y.Y., Vasil’eva, L.V., Chestnykh, O.V., Zavarzin, G.A., 2002. Methanotrophs of the psychrophilic microbial community of the Russian arctic tundra. Microbiology 71 (4), 460–466. Blicher-Mathiesen, G., McCarty, G.W., Nielsen, L.P., 1998. Denitrification and degassing in groundwater estimated from dissolved dinitrogen and argon. J. Hydrol. 208 (1–2), 16–24. Boon, P.I., Mitchell, A., Lee, K., 1997. Effects of wetting and drying on methane emissions from ephemeral floodplain wetlands in south-eastern Australia. Hydrobiologia 357, 73–87. Brix, H., Sorrell, B.K., Lorenzen, B., 2001. Are Phragmitesdominated wetlands a net source or net sink of greenhouse gases? Aquat. Bot. 69 (2–4), 313–324. Brooks, P.D., Schmidt, S.K., Williams, M.W., 1997. Winter production of CO2 and N2 O from alpine tundra: environmental controls and relationship to inter-system C and N fluxes. Oecologia 110 (3), 403–413. Butnor, J.R., Johnsen, K.H., Oren, R., Katul, G.G., 2003. Reduction of forest floor respiration by fertilization on both carbon dioxideenriched and reference 17-year-old loblolly pine stands. Global Change Biol. 9 (6), 849–861. Butterbach-Bahl, K., Willibald, G., Papen, H., 1997. A new method for simultaneous measurements of N2 and N2 O-emissions from intact soil cores. In: Van Cleemput, O., Haneklaus, S., Hofman, G., Schnug, E., Vermoesen, A. (Eds.), Fertilization for Sustainable Plant Production and Soil Fertility. Proceedings of 11th World Fertilizer Congress of CIEC, vol. 2, pp. 618– 624. Cao, M., Gregson, K., Marshall, S., 1998. Global methane emission from wetlands and its sensitivity to climate change. Atmos. Environ. 32 (19), 3293–3299. Christensen, T.R., Panikov, N., Mastepanov, M., Joabsson, A., Stewart, A., Öquist, M., Sommerkorn, M., Reynaud, S., Svensson, B., 2003. Biotic controls on CO2 and CH4 exchange in wetlands—a closed environmental study. Biogeochemistry 64 (3), 337–354. Clark, K.L., Gholz, H.L., Moncrieff, J.B., Cropley, F., Loescher, H.W., 1999. Environmental controls over net exchanges of carbon dioxide from contrasting Florida ecosystems. Ecol. Appl. 9 (3), 936–948. 539 Cooper, A.B., 1990. Nitrate depletion in the riparian zone and stream channel of a small headwater catchment. Hydrobiologia 202 (1–2), 13–26. Denmead, O.T., Raupach, M.R., 1993. Agricultural Ecosystem Effects on Trace Gases and Global Climate Change. American Society of Agronomy, Special Publication No. 55, pp. 19–43. Dowrick, D.J., Hughes, S., Freeman, C., Lock, M.A., Reynolds, B., Hudson, J.A., 1999. Nitrous oxide emissions from a gully mire in mid-Wales, UK, under simulated summer draught. Biogeochemistry 44 (2), 151–162. Erich, M.S., Bekerie, A., Duxbury, J.-M., 1984. Activities of denitrifying enzymes in freshly sampled soils. Soil Sci. 138 (1), 25–32. Fey, A., Benckiser, G., Ottow, J.C.G., 1999. Emissions of nitrous oxide from a constructed wetland using a groundfilter and macrophytes in waste-water purification of a dairy farm. Biol. Fertil. Soils 29 (4), 354–359. Funk, D.W., Pullman, E.R., Peterson, K.M., Crill, P.M., Billings, W.D., 1994. Influence of water-table on carbon dioxide, carbon monoxide, and methane fluxes from taiga bog microcosms. Global Biogeochem. Cycles 8 (3), 271–278. Gold, A.J., Jacinthe, P.A., Groffman, P.M., Wright, W.R., Puffer, R.H., 1998. Patchiness in groundwater nitrate removal in a riparian forest. J. Environ. Qual. 27 (1), 146–155. Griffis, T.J., Rouse, W.R., Waddington, J.M., 2000. Interannual variability of net ecosystem CO2 exchange at a subarctic fen. Global Biogeochem. Cycles 14 (4), 1109–1121. Groffman, P.M., Axelrod, E.A., Lemunyon, J.L., Sullivan, W.M., 1991. Denitrification in grass and forest vegetated filter strips. J. Environ. Qual. 20 (3), 671–674. Groffman, P.M., Gold, A., Addy, K., 2000. Nitrous oxide production in riparian zones and its importance to national emission inventories. Chemosphere—Global Change Sci. 2 (3–4), 291– 299. Gulledge, J., Schimel, J.P., 2000. Controls on soil carbon dioxide and methane fluxes in a variety of taiga forest stands in interior Alaska. Ecosystems 3 (3), 269–282. Hamilton, S.K., Sippel, S.J., Melack, J.M., 1995. Oxygen depletion and carbon dioxide and methane production in waters of the Pantanal wetland in Brazil. Biogeochemistry 30 (2), 115–141. Hanson, G.C., Groffman, P.M., Gold, A.J., 1994. Denitrification in riparian wetlands receiving high and low groundwater nitrate inputs. J. Environ. Qual. 23 (5), 917–922. Haycock, N.E., Pinay, G., 1993. Groundwater nitrate dynamics in grass and poplar vegetated riparian buffer strips during the winter. J. Environ. Qual. 22 (2), 273–278. Hefting, M.M., de Klein, J.J.M., 1998. Nitrogen removal in buffer strips along a lowland stream in the Netherlands: a pilot study. Environ. Pollut. 102 (S1), 521–526. Hill, A.R., 1996. Nitrate removal in stream riparian zones. J. Environ. Qual. 25, 743–755. Hutchinson, G.L., Livingston, G.P., 1993. Use of chamber systems to measure trace gas fluxes. In: Duxbury, J.M., et al. (Eds.), Agricultural Ecosystems Effects on Trace Gases and Global Climate Change. American Society of Agronomy, Madison, MI, pp. 1–55, ASA Special Publication No. 55. IPCC, 2001. Atmospheric chemistry and greenhouse gases. In: Houghton, J.T., et al. (Eds.), Climate Change (2001). The Scien- 540 S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 tific Basis. Cambridge University Press, Cambridge, New York, pp. 239–287 (Chapter 4). Jacinthe, P.A., Groffman, P.M., Gold, A.J., Mosier, A., 1998. Patchiness in microbial nitrogen transformations in groundwater in a riparian forest. J. Environ. Qual. 27 (1), 156–164. Johansson, A.E., Klemedtsson, A.K., Klemedtsson, L., Svensson, B.H., 2003. Nitrous oxide exchanges with the atmosphere of a constructed wetland treating wastewater – Parameters and implications for emission factors. Tellus B 55 (3), 737– 750. Joiner, D.W., Lafleur, P.M., McCaughey, J.H., Bartlett, P.A., 1999. Interannual variability in carbon dioxide exchanges at a boreal wetland in the BOREAS northern study area. J. Geophys. Res. Atmos. 104 (D22), 27663–27672. Jones, J.B., Mulholland, P.J., 1998a. Carbon dioxide variation in a hardwood forest stream: an integrative measure of whole catchment soil respiration. Ecosystems 1 (2), 183–196. Jones, J.B., Mulholland, P.J., 1998b. Methane input and evasion in a hardwood forest stream: Effects of subsurface flow from shallow and deep pathways. Limnol. Oceanogr. 43 (6), 1243–1250. Kadlec, R.H., Knight, R.L., 1996. Treatment Wetlands. Lewis Publishers, Boca Raton, 893 pp. Kettunen, A., 2003. Connecting methane fluxes to vegetation cover and water table fluctuations at microsite level: a modelling study. Global Biogeochem. Cycles 17 (2), 1051, doi:10.1029/2002GB001958. Knowles, R., 1982. Denitrification. Microbiol. Rev. 46 (1), 43–70. Kormann, R., Muller, H., Werle, P., 2001. Eddy flux measurements of methane over the fen “Murnauer Moos”, 11◦ 11′ E, 47◦ 39′ N, using a fast tunable diode laser spectrometer. Atmos. Environ. 35 (14), 2533–2544. Kuusemets, V., Mander, Ü., 1999. Ecotechnological measures to control nutrient losses from catchments. Water Sci. Techol. 40 (10), 195–202. Kuusemets, V., Mander, Ü., Lõhmus, K., Ivask, M., 2001. Nitrogen and phosphorus variation in shallow groundwater and assimilation in plants in complex riparian buffer zones. Water Sci. Techol. 44 (11–12), 615–622. Lafleur, P.M., McCaughey, J.H., Joiner, D.W., Bartlett, P.A., Jelinski, D.E., 1997. Seasonal trends in energy, water, and carbon dioxide fluxes at a northern boreal wetland. J. Geophys. Res. Atmos. 102 (D24), 29009–29020. Loftfield, N., Flessa, H., Augustin, J., Beese, F., 1997. Automated gas chromatographic system for rapid analysis of the atmospheric trace gases methane, carbon dioxide, and nitrous oxide. J. Environ. Qual. 26 (2), 560–564. Lowrance, R., 1992. Groundwater nitrate and denitrification in a coastal plain riparian forest. J. Environ. Qual. 21 (3), 401–405. Lund, L.J., 1999. Nitrogen balance in a pond system receiving tertiary effluent. J. Environ. Qual. 28 (4), 1258–1263. Lund, L.J., Horne, A.J., Williams, A.E., 2000. Estimating denitrification in a large constructed wetland using stable nitrogen isotope ratios. Ecol. Eng. 14 (1–2), 67–76. MacDonald, J.A., Fowler, D., Hargreaves, K.J., Skiba, U., Leith, I.D., Murray, M.B., 1998. Methane emission rates from a northern wetland; response to temperature, water table and transport. Atmos. Environ. 32 (19), 3219–3227. Mander, Ü., Kuusemets, V., Öövel, M., Ihme, R., Sevola, P., Pieterse, A., 2000. Experimentally constructed wetlands for wastewater treatment in Estonia. J. Environ. Sci. Heal. A 35 (8), 1389–1401. Mander, Ü., Kuusemets, V., Öövel, M., Mauring, T., Ihme, R., Sevola, P., Pieterse, A., 2001. Wastewater purification efficiency in experimental treatment wetlands in Estonia. In: Vymazal, J. (Ed.), Nutrient Transformations in Natural and Constructed Wetlands. Backhuys Publishers, Leiden, pp. 201–224. Mander, Ü., Kuusemets, V., Lõhmus, K., Mauring, T., Teiter, S., Augustin, J., 2003. Nitrous oxide, dinitrogen, and methane emission in a subsurface flow constructed wetland. Water Sci. Technol. 48 (5), 135–142. Mander, Ü., Lõhmus, K., Kuusemets, V., Ivask, M., 1997. The potential role of wet meadows and grey alder forests as buffer zones. In: Haycock, N.E., Burt, T.P., Goulding, K.W.T., Pinay, G. (Eds.), Buffer Zones, Their Processes and Potential in Water Protection. Quest Environmental, Oxford, pp. 35–46. Martikainen, P.J., Nykänen, H., Crill, P., Silvola, J., 1993. Effect of a lowered water table on nitrous oxide fluxes from northern peatlands. Nature 366 (6450), 51–53. Mitsch, W.J., Gosselink, J.G., 1993. Wetlands. Van Nostrand Reinhold, New York, 722 pp. Mosier, A.R., 1998. Soil processes and global changes. Biol. Fertil. Soils 27 (3), 221–229. Mogge, B., Kaiser, E.-A., Munch, J.-C., 1998. Nitrous oxide emissions and denitrification N-losses from forest soils in the Bornhöved Lake Region (Northern Germany). Soil Biol. Biochem. 30 (6), 703–710. Nelson, W.M., Gold, A.J., Groffman, P.M., 1995. Spatial and temporal variation in groundwater nitrate removal in a riparian forest. J. Environ. Qual. 24 (4), 691–699. Noorvee, A., Repp, K., Põldvere, E., Mander, Ü., 2005. Aeration effects and the application of the k-C* model in a subsurface flow constructed wetland. J. Environ. Sci. Heal. A. 40 (6–7), 1445–1456. Paludan, C., Blicher-Mathiesen, G., 1996. Losses of inorganic carbon and nitrous oxide from a temperate freshwater wetland in relation to nitrate loading. Biogeochemistry 35 (2), 305–326. Pinay, G., Roques, L., Fabre, A., 1993. Spatial and temporal patterns of denitrification in a riparian forest. J. Appl. Ecol. 30 (4), 581–591. Peterjohn, W.T., Correll, D.L., 1984. Nutrient dynamics in an agricultural watershed: observations on the role of a riparian forest. Ecology 65 (5), 1466–1475. Rask, H., Schoenau, J., Anderson, D., 2002. Factors influencing methane flux from a boreal forest etland in Saskatchewan, Canada. Soil Biol. Biochem. 34 (4), 435–443. Reay, D.S., Radajewski, S., Murrell, J.C., McNamara, N., Nedwell, D.B., 2001. Effects of land-use on the activity of methane oxidizing bacteria in forest soils. Soil Biol. Biochem. 33 (12–13), 1613–1623. Robertson, G.P., Tiedje, J.M., 1984. Denitrification and nitrous oxide production in successional and old-growth Michigan forest. Soil Sci. Soc. Am. J. 48 (2), 383–389. Rusch, H., Rennenberg, H., 1998. Black alder (Alnus glutinosa (L.) Gaertn.) trees mediate methane and nitrous oxide emission from the soil to the atmosphere. Plant Soil 201 (1), 1–7. S. Teiter, Ü. Mander / Ecological Engineering 25 (2005) 528–541 Saarnio, S., Alm, J., Silvola, J., Lohila, A., Nykänen, H., Martikainen, P.J., 1997. Seasonal variation in CH4 emissions and production and oxidation potentials at microsites on an oligotrophic pine fen. Oecologia 110 (3), 414–422. Schipper, L.A., Cooper, A.B., Harfoot, C.G., Dyck, W.J., 1993. Regulators of denitrification in an organic riparian soil. Soil Biol. Biochem. 25 (7), 925–933. Schnabel, R.R., Shaffer, J.A., Stout, W.L., Cornish, L.F., 1997. Denitrification distributions in four valley and ridge riparian ecosystems. Environ. Manage. 21 (2), 283–290. Scholefield, D., Hawkins, J.M.B., Jackson, S.M., 1997. Development of a helium atmosphere soil incubation technique for direct measurement of nitrous oxide and dinitrogen fluxes during denitrification. Soil Biol. Biochem. 29 (9–10), 1345– 1352. Scott, R.L., Edwards, E.A., Shuttleworth, W.J., Huxman, T.E., Watts, C., Goodrich, D.C., 2004. Interannual and seasonal variation in fluxes of water and carbon dioxide from a riparian woodland ecosystem. Agr. Forest Meteorol. 122 (1–2), 65–84. Shurpali, N.J., Verma, S.B., Clement, R.J., Billesbach, D.P., 1993. Seasonal distribution of methane flux in a Minnesota peatland measured by eddy-correlation. J. Geophys. Res. Atmos. 98 (D11), 20649–20655. Spieles, D.J., Mitsch, W.J., 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 (1–2), 77–91. Struwe, S., Kjøller, A., 1990. Seasonality of denitrification in waterlogged alder stands. Plant Soil 128 (1), 109–113. Tai, P.-D., Li, P.-D., Sun, T.-H., He, Y.-W., Zhou, Q.-Z., Gong, Z.Q., Mizuochi, M., Inamori, Y., 2002. Greenhouse gas emissions from a constructed wetland for municipal sewage treatment. J. Environ. Sci. (China) 14 (1), 27–33. Tanner, C.C., Adams, D.D., Downes, M.T., 1997. Methane emissions from constructed wetlands treating agricultural wastewaters. J. Environ. Qual. 26 (4), 1056–1062. 541 Tanner, C.C., Kadlec, R.H., Gibbs, M.M., Sukias, J.P.S., Nguyen, M.L., 2002. Nitrogen processing gradients in subsurface-flow wetlands—influence of wastewater characteristics. Ecol. Eng. 18, 499–520. Van den Pol-Van Dasselaar, A., Van Beusichem, M.L., Oenema, O., 1999. Methane emissions from wet grasslands on peat soil in a nature preserve. Biogeochemistry 44 (2), 205–220. Waddington, J.M., Roulet, N.T., 2000. Carbon balance of a boreal patterned peatland. Global Change Biol. 6 (1), 87–97. Watts, S., Seitzinger, S.P., 2000. Denitrification rates in organic and mineral soils from riparian sites: a comparison of N2 flux and acetylene inhibition methods. Soil Biol. Biochem. 32 (10), 1383–1392. Weller, D.E., Correll, D.L., Jordan, T.E., 1994. Denitrification in riparian forests receiving agricultural discharges. In: Mitsch, W.J. (Ed.), Global Wetlands: Old World and New. Elsevier, New York, pp. 117–131. Whiting, G.J., Chanton, J.P., 2001. Greenhouse carbon balance of wetlands: methane emission versus carbon sequestration. Tellus B 53 (5), 521–528. Wickland, K.P., Striegl, R.G., Schmidt, S.K., Mast, M.A., 1999. Methane flux in subalpine wetland and unsaturated soils in the southern Rocky Mountains. Global Biogeochem. Cycles 13 (1), 101–113. Wild, U., Lenz, A., Kamp, T., Heinz, S., Pfadenhauer, J., 2002. Vegetation development, nutrient removal and trace gas fluxes in constructed Typha wetlands. In: Mander, Ü., Jenssen, P.D. (Eds.), Natural Wetlands for Wastewater Treatment in Cold Climates, vol. 12. WIT Press, Southampton and Boston, pp. 101–125, Adv. Ecol. Sci. Willison, T.W., Baker, J.C., Murphy, D.V., 1998. Methane fluxes and nitrogen dynamics from a drained fenland peat. Biol. Fertil. Soils 27 (3), 279–283. Xue, Y., Kovacic, D.A., David, M.B., Gentry, L.E., Mulvaney, R.L., Lindau, C.W., 1999. In situ measurements of denitrification in constructed wetlands. J. Environ. Qual. 28 (1), 263–269.