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Hydrochemistry and hydrology of forest riparian wetlands
Forest Ecology and Management 196 (2004) 187–197
Hydrochemistry and hydrology of forest riparian wetlands G. Jacks*, A-C. Norrstro¨m Land and Water Resources Engineering, Royal Institute of Technology, SE-100 44, Stockholm, Sweden Received 30 July 2001; received in revised form 29 August 2003; accepted 6 January 2004
Abstract Forest stream riparian wetlands have a number of important features. This investigation treats one aspect, the nitrogen retention after upland clear-cutting which leads to elevated nitrate leaching, and the importance of the flow pathways in this connection. The runoff occurs mainly via the upper permeable section of the peat while the lower peat act as an aquitard, restricting the flow from the underlying till. The till groundwater is progressively artesian towards the discharging stream. Water analyses from piezometers show that the water chemistry in the peat is rather variable, indicating the presence of channelling. Channelling is also indicated by spring discharges from the peat that have elevated nitrate contents pointing to bypass flow. Redox bars indicating sulphate reduction display the same picture of irregular distribution. However, a general observation is that volumes with sulphate reduction increase towards the stream and that sulphate reduction occurs preferably in the surface peat, indicating the importance of a degradable substrate for the sulphate reducers. Nitrate reduction during the growth season occurs predominantly close to the upland till areas. The riparian tree stand is dominated by spruces which are likely to be disfavoured by the rising groundwater level after clear-cutting. The riparian tree stand does not extend far enough towards the upland to be benefited by the elevated nitrate flux. Buffer stands should be broader, extending into the till upland where they can utilise the leached nitrate and, more important, thanks to their deeper rooting depth protect the peatland trees against wind felling. # 2004 Elsevier B.V. All rights reserved. Keywords: Clear-cutting; Groundwater; Nitrate; Denitrification; Sulphate reduction; Riparian wetland
1. Introduction The marine waters between Sweden and Denmark have, since the mid-70s, experienced algal blooming caused by an elevated inflow of nitrogen from the surrounding land (Rosenberg et al., 1990). It is believed that most of the nitrogen comes from agricultural land, but from the Swedish side the transport from forest areas may actually be higher than that from agricultural areas and may be increasing (Fig. 1). The *
Corresponding author. Tel.: þ46-98-790-7380; fax: þ46-8-411-0775. E-mail address: [email protected] (G. Jacks).
nitrogen from forest areas contains a considerable portion of organic nitrogen which, however, is also available to marine algae (Carlsson, 1993). The forest in SW Sweden receives a nitrogen deposition of about 20 kg N ha1 per year. The removal during forest harvesting is of the order of 5–10 kg N ha1 per year and the runoff losses are about 2–5 kg N ha1 per year. Thus, about 10 kg N ha1 per year are accumulated in the soil each year. While the sulphur deposition has decreased since the 70s, there is little downward trend in nitrogen deposition (Naturva˚rdsverket/SLU, 1999). Elevated nitrogen deposition has resulted in nitrogen saturation in spruce forests in Central Europe (Katzensteiner et al., 1992). In Germany, secondary
0378-1127/$ – see front matter # 2004 Elsevier B.V. All rights reserved. doi:10.1016/j.foreco.2004.01.055
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Fig. 1. Transport of nitrogen by streams in the Lagan catchment, SW Sweden. The transport figures are normalised with respect to water flow (Joelsson and Fleischer, County Board of Holland).
magnesium deficiency has been observed (Ulrich, 1992). In SW Sweden stands exposed to high deposition have shown nitrogen saturation in the sense that nitrate leaching occurs below the root zone even during the vegetation season in intact spruce stands (Nohrstedt et al., 1996). Thelin et al. (1998) have observed sub-optimum contents of potassium and copper in spruce needles in South Sweden and a declining trend over time, indicating an approaching nutrient imbalance which may decrease the utilisation of soil nitrogen. Nitrification is promoted by a lowering of the C/N ratio in the organic horizon due to the gradual accumulation of nitrogen in the soil. A ratio below about 25 is critical according to Kriebitzsch (1978) and Gundersen et al. (1998). An additional criterion for nitrification is that the nitrogen deposition should be elevated (Dise et al., 1998). One-third of the spruce stands in SW Sweden have a ratio below that value (Fig. 2) in contrast to conditions in central Sweden (Fo¨ lster, 2000). Even though a low C/N ratio and a high deposition would indicate that nitrification should occur, nitrification may take time to induce according to Rudebeck and Persson (1998). So far only limited areas with an intact forest stand seem to be nitrogen-saturated, but large amounts of nitrogen are mineralised and mobilised from slash and soil after clear-cutting. As much as 700 kg N ha1 during 5 years after clear-cutting has been recorded as being lost below the root zone after harvest in SW ¨ rlander et al., 1997). This amount may not, Sweden (O however, reach draining water courses if there are good conditions for nitrogen retention in riparian areas. Nitrogen retention may be due to denitrification, dissimilitary nitrate reduction and plant uptake.
Uptake by residual stands has been studied in Central and Northern Sweden by Lundell and Albrektson (1997). They found an increase in growth in two stands out of five, indicating a fertilising effect of nitrate leached from a clear-cut upland. Shelter stands have been found to retain nitrogen to a ¨ rlander and Karlsson, 2000). The suclarge extent (O cession of the field layer and the mosses is much more gentle in a shelter stand than after clear-cutting (Hannerz, 1996). Unfortunately, shelter stands with spruce are not found in SW Sweden due to the risk of wind felling, so that clear felling is the general practice. This investigation has been aimed at studying nitrogen retention in riparian wetlands in SW Sweden after clear-cutting. The reasons are: High and increasing transport from forests (Fig. 1); Persisting high nitrogen deposition to forest land;
Fig. 2. C/N ratio in spruce stands in Holland (from Swedish Forest Inventory).
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Critically low C/N ratios in the humus layer in about 1/3 of the sites (Fig. 2); Large leaching losses from clear-cut areas. A further aim is to provide a basis for the management of the riparian wetlands in connection to forest harvesting.
2. Materials and methods 2.1. Field sites Two field sites were chosen for the studies, Luntoma in which the riparian wetlands were essentially natural and Ho¨ ga˚ sen in which drainage was undertaken. The areas studied were 6.4 and 2.0 ha, respectively. The field site with the natural wetland was clear-cut in 1998 to 73%, leaving untouched wetland area of 0.6 ha together with an upland spruce stand of 1.1 ha in the westernmost part of the site (Fig. 3). The wetland forest stand was 10–30 m wide. The ground layer in the riparian areas was mostly Spaghnum mosses with stands of Juncus effu´ sus. Trees were dominantly spruce with a few small birches and a few alders. The wetlands were covered with peat up to 0.8–1 m depth. The drained field site was completely clear-cut in 1998 except for a beech stand covering
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0.2 ha. The clear-cut areas had carried closed stands of Norway spruce with a ground layer of Holycomium splendens and some Polytrichum mosses. Due to the closed canopy there was virtually no field layer and most of the mosses died due to light impact after the clear-cutting. The soil was a sandy-loamy till. After clear-cutting, stems were removed and after 1 year branches and tops also as firewood. The ground layer partly disappeared due to the exposure to light and coverage by slash. During the first year, only a few herbs and grass appeared, with a coverage of only 1– 2%. Descha´ mpsia fiexuo´sa and Sene´cio vulgaris were common pioneering species after the clear-cutting. In the second year, scarification was carried out leaving 0.3 m broad flirrows and 0.3 m beds every 2.5 m. Thus, about 25% of the area was disturbed. Spruce seedlings were planted during the second year. Groundwater piezometers were installed in transects from the upland margin through the riparian wetland to the stream (Fig. 4). In the wetland area, piezometers were installed in the peat and in the underlying till. Each area was supplied with two reference piezometers in undisturbed upland areas (Fig. 4). The Ho¨ ga˚ sen area was supplied with a total of 10 piezometers while Luntorna had a total of 30 piezometers. In Ho¨ ga˚ sen two channels in the peat discharging into the ditch were also used as sampling points. Runoff was registered in V-notch dams at the
Fig. 3. Sketch map of the Luntoma catchment with installations marked.
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Fig. 4. Cross-section of the riparian wetland in the Luntoma catchment with installation of piezometers.
exit from the studied sites. The investigation is only halfway through the post-clear-cut period with elevated N-leaching. A budget concerning nitrogen for the third year after clear-cutting is presented here to further illustrate the findings regarding hydrology and hydrochemistry in the riparian areas. 2.2. Sampling and analysis Samples were taken from the piezometers every 5–6 weeks and the runoff was sampled at the same intervals in the summertime and more often during periods with high flow, notably in the autumn. Before sampling, the piezometers were pumped with a volume equal to twice the volume standing in the tube to ensure representative samples. Samples for N-species, major anions and cations and dissolved organic carbon were filtered in the field through 0.2 mm filters. The pH was also measured in the field. Groundwater levels were recorded on the sampling occasions. Redox measurements using platinum electrodes are relatively easy to carry out but they are not feasible due to the spatial variability of the redox conditions (Norrstro¨ m, 1994). Alternative methods of registering redox conditions were used in the form of iron and copper bars that were pushed into the peat. Some bars were polished while some were painted with lead oxide paint. The bars were kept in place for about 6 weeks at a time. The polished iron bars showed areas with oxygen characterised by rust and areas charac-
terised by grey deposits of magnetite (Tor-Gunnar Vinka, personal communication, Swed. Corrosion Institute). The redox potential at which magnetite is formed at a pH of about 6 is of the order of 150 mV (Garrels and Christ, 1965). Copper bars were oxidised in the top portion while they were still unchanged lower down indicating a redox-cline at about the same level (Garrels and Christ, 1965). The red-orange lead oxide-painted-bars turned black where hydrogen sulphide was formed. The patterns were copied onto sheets of plastic, cut out with scissors and weighed to assess the respective areas. These were taken as representing volumes in the peat. The redox bars were inserted in duplicate or triplicate at each site. As the formation of lead sulphide is irreversible, the bars indicate the maximum extent of sulphate reduction during the period of insertion. In Luntorna, seepage rates were measured in the stream bed using seepage meters (Norrstro¨ m and Jacks, 1996). In Ho¨ ga˚ sen, the temperature was measured on one occasion during the summer to detect channelled inflow of groundwater. Hydraulic conductivity was assessed in piezometers by ‘‘slug tests’’ (Mandel and Shiftan, 1981). These were performed by pouring water into the piezometers and measuring the recovery of the water level. This gives a measure of the hydraulic conductivity in the close neighbourhood of the peizometers. Composite samples from stands of Juncus effu´ sus were collected in October 2000 and subject to nitrogen
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isotope analysis at the Department of Forest Ecology at the Swedish University of Agricultural Sciences in Umea˚ .The nitrogen species were determined in the laboratory by a Tecator Aquatech instrument. Tot-N was determined by oxidation with K-peroxidisulphate to NO3-N. Major anions (Cl, NO3, SO42) and cations (Ca2þ, Mg2þ, Naþ, Kþ) were measured with a Dionex DX 120 ion chromatograph.
3. Results 3.1. Hydrology The hydraulic conductivity as measured by ‘‘slug tests’’ (Fig. 5) was low in the lower part of the peat (0.6–1 m depth), and intermediate in the till and at intermediate depths in the peat (0.3–0.4 m depth). It was very high in the uppermost section of the peat which was observed as the recovery after pumping. This means that the lower portion of the peat acted as an aquitard separating the till groundwater from the peat groundwater. This was manifested as artesian conditions in the till groundwater. The pressure level in the till groundwater was 4–16 cm above the peat groundwater level, with the difference increasing towards the draining stream. The transmissivity as measured by ‘‘slug tests’’ (Mandel and Shiftan, 1981) varied over several orders of magnitude in each of the compartments but the deeper peat was clearly the most impermeable (Fig. 5). The surface peat at about 0.1–0.2 m depth was so permeable, of the order of 0.001 m/s that it was impossible to observe the
Fig. 6. Temperature measurement in the drainage ditch at Ho¨ ga˚ sen indicating channelled inflow into the ditch.
recovery. This leaves two main drainage pathways, in the upper peat and via the till. The till is most permeable in its uppermost parts as shown by Lind and Lundin (1990). This is due to several factors, frost heaving, root penetration and the presence of surface till more loosely deposited on a bottom till compressed below the inland ice in glacial times. Thus, deeper pathways probably contribute only minor amounts to the runoff. The peat is channelled, manifested in a large range of hydraulic conductivities in the peat piezometers and a variable chemistry. Seepage rates were measured in the stream in the Luntoma site in November 2000 and they varied from 0.02 to 16.6 l m2 per hour ðn ¼ 9Þ. Further evidence of the channelling of the peat was the presence of two channels emptying in the slope of the ditch in the Ho¨ ga˚ sen site with a discharge varying in time from 0.1 to 2 l min1. Temperature measurements during the summer revealed an inflow of colder groundwater at two other points (Fig. 6). Piezometers installed close to each other (within a distance of 0.5 m) at depths of 0.2, 0.4 and 0.8 m in the peat showed differences in groundwater heads of up to 6 cm. This indicates a flow in closed channels at a depth within the peat. 3.2. Redox conditions
Fig. 5. Hydraulic conductivity in till, deep peat and intermediate peat obtained by ‘‘slug tests’’.
The lead-painted redox bars indicated increasing sulphate reduction volumes towards the stream in the Luntoma catchment (Fig. 7). Close to the upland, the sulphate reduction was, predominantly present in the lower section of the profile, while it was shifting
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Fig. 7. Sulphate reduction as monitored by redox bars in the riparian wetland in Luntoma catchment. LT-12 to LT-14 refer to the sites in Fig. 3.
upwards in the profile closer to the stream. The sulphate reduction at a depth close to the upland is due to the influence of more oxidising conditions close to the surface, e.g. the presence of nitrate as an electron acceptor. Closer to the stream, the availability of digestible substrate near the surface moves the sulphate reduction closer to the surface. The lower degree of sulphate reduction close to the stream in May and October profiles in Fig. 7 may be caused by the hyphoreic zone which exhibits water exchange with the stream. The redox bars in the lower parts of the wetland exhibit a spotted appearance indicating a
channelled nature of the flow in the peat, which supports the hydrological findings. The large volumes of peat exhibiting sulphate reduction does not correspond to the sulphate reduction as observed in the piezometers and runoff, about 20% in February and 50% in June (Fig. 8). The explanation is probably to be that the sulphate reduction takes place in less permeable sections of the peat and that affects only a part of the total water flow. Due to the lower dry deposition after the forest harvest there were gradually declining levels of sulphate and chloride with time in all the samples, levelling out after about 2 years. This indicates that the maximum turnover time of the upland groundwater was of that order of magnitude. In Ho¨ ga˚ sen, the redox conditions seemed on the whole to be less reducing (Fig. 9). This may be due to drainage, which has lowered the groundwater level at some distance from the stream, into the older peat layers where the organic substrate is usually more refractory (Koops et al., 1996). The cutting down of the riparian forest has left a lot of slash which does not permit any ground vegetation to develop and thus lowers the input of fresh organic matter. In the spruce needles in the slash, little degradable matter is left 2 years after the harvest (Berg et al., 1996). The uncoated iron bars showed a zone of red rust close to the groundwater level and a zone without
Fig. 8. Sulphate contents in piezometers in the February and June sampling in Luntoma catchment. Shaded areas are peat.
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Fig. 9. Sulphate reduction as monitored by redox bars in the riparian wetland in Ho¨ ga˚ sen catchment.
corrosion products below that level. This indicates an aeration cell in which liberated iron is deposited as red rust. Below these two zones, at about 15–20 cm depth there was a grey zone with a deposit of magnetite (TorGunnar Vinka, personal communication, Swed. Corrosion Institute). As already mentioned, this zone should be characterised by a redox potential of the order of 150 mV. The copper bars, used in a few numbers, were oxidised only in the top 15 cm giving a similar redox level at that depth as the iron bars.
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Thus, nitrate reduction during the growth season must occur above this zone and close to the upland margin, as is evident from the decreasing nitrate concentrations in the peat piezometers (Fig. 10) as well as the lower redox level downslope leading to sulphate reduction. The nitrate reduction during the growth season may be due to plant uptake and in this case uptake in the field layer as the tree strands were largely situated below the zone of nitrate reduction. Other possible mechanisms of nitrate reduction are denitrification and dissimilitary nitrate reduction to ammonium (DNRA). Matheson et al. (2002) have in a microcosm experiment studied the partitioning between these two pathways. They found that a microcosm with plants exhibited a dominance of denitrification while the unplanted one showed largely DNRA. The organic carbon to nitrate ratio is also important (Matheson et al., 2002), or rather the degradeability of the organic matter. A highly degradable organic matter would favour more reducing conditions leading to DNRA (Buresh and Patrick, 1981). In this case the organic matter is rather refractive, largely consisting of Spaghnum peat which would indicate that denitrification is a major pathway. If DNRA would be a major pathway there could be some loss of ammonium in
Fig. 10. Nitrate–nitrogen contents in the profile LT-12 to LT-14. Shaded areas are peat. Observation points marked with circles in the December profile are new piezometers. The arrow into the stream in the December profile is a sample from a seepage meter.
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runoff especially in winter time. However, this is not observed. The combined effect of denitrification and plant uptake in June is in the order of 85% of the input into the profile, as recorded in piezometers LT-12 to LT-14 (Fig. 10). Very little of this is uptake by trees, as the buffer stand is on the whole situated below LT-13 (Figs. 3 and 4) where the nitrate content has already decreased to a background level (Fig. 9). A sizeable fraction of the nitrate reduction is denitrification which is evident from the d15N values in composite samples of Juncus effu´ sus sampled in strips from the clear-cut to the stream (Table 1). The LT-7 to LT-9 profile of pizometers is parallel to the LT-12 to LT-14 profile, about 20 m to the southwest. It is also evident from the d15N values that the denitrification occurs near the upland margin of the wetland. During the cold season, there is a nitrate breakthrough in the upper layers of the peat all the way down to the stream in Luntoma (Fig. 10) as well as via channelled water in Ho¨ ga˚ sen, which carry water with elevated nitrate concentrations directly into the stream (Fig. 11). This is presumably a bypass flow in the peat, more or less directly from the upstream till area, as was observed in another investigation (Norrstro¨ m and Jacks, 1996). The nitrate reduction during the coldest months, January and February, was of the order of 30–50% in Luntoma (Table 1). On the whole, the nitrate reduction
Table 1 Transport of Tot-N from the clear-cut area, from intact forest and wetland and export from the catchment in the year 2000 Month January February March April May June July August September October November December
Clear-cut Tot-N (kg ha1) 24.3 18.0 12.5 9.81 5.44 9.32 6.17 7.12 10.2 22.5 36.7 29.3
Total loss/loss 191.2/41 kg ha1
Forest Tot-N (kg ha1)
Runoff Tot-N (kg ha1)
0.42 0.20 0.35 0.34 0.22 0.27 0.05 0.16 0.36 0.18 0.21 0.11
16.9 10.6 4.90 2.43 0.69 1.31 0.83 1.11 1.78 3.20 10.8 13.2
2.87/1.7
67.7/11
The lowermost row presents total loss from the respective compartments and loss per hectare.
seems to be lesser in the drained catchment at Ho¨ ga˚ sen (Table 2; Fig. 11). Anaerobicity and availability of substrate are the key factors that determine where denitrification occurs (Ettema et al., 1999). The drainage in Ho¨ ga˚ sen has decreased the contact between the groundwater and the surface soil with its more
Fig. 11. Nitrate–nitrogen contents in the profile HA-1 to HA-2. Shaded areas are peat. Observation point with arrow is a spring discharge into the ditch.
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Fig. 12. Redox-conditions during the growth season in the profile LT-12 to LT-14 inferred from redox studies and chemistry in piezometers.
degradable substrate (Koops et al., 1996). The buffer zone which remains more intact in Luntoma produces a litter which in combination with a high groundwater level promotes denitrification as well as sulphate reduction. The N-budget is so far only calculated for the Luntoma catchment and for the year 2000. The runoff is measured at the exit from Luntoma. The total runoff was 720 mm out of a total rainfall of 1200 mm. To distribute the runoff on the intact forest stand, the wetland and the clear-cut we have assumed that the runoff from the clear-cut is having 200 mm more yearly runoff than the intact stand and the wetland in conformity with the findings by Lindroth and Grip (1987) and Brandt et al. (1988). As concentration of
Table 2 d15N in Juncus effu´ sus along the section in Fig. 3 (clear-cut, LT-12 to LT-14, stream side) and a parallel section (LT-7 to L-9) and mean nitrate concentration in soil water and groundwater during the vegetation season (mean of four samplings from May to September) Site
Mg NO3-N/L
d15N
Clear-cut LT-12 LT-13 LT-14 Stream-side LT-7 LT-8 LT-9
5.1 4.4 0.17 0.14 0.23 3.57 1.16 0.36
1.03 5.46 8.18 7.61 6.60 6.36 7.19 7.31
the runoff from the clear-cut is taken the mean value of five piezometers situated at the transition between the clear-cut and the wetland. As concentration of the runoff from the intact forest stand is taken the mean value of the two reference piezometers. As the concentration in water that originates from the wetland is taken the values from samples from a small brook coming from the north in the center of the wetland. With this calculation the clear-cut at the Luntoma catchment lost 41 kg N ha1 during the third year after harvesting. Nitrogen loss from the intact stand was 1.7 kg N ha1 the same year. The losses via the stream were 11 kg N ha1 for the whole catchment. The retention was 73% over the whole year, decreasing to about half in the winter months (Table 1). The nitrogen budget for the year 2000, the third year after clear-cutting does not represent the peak leaching which is expected only in the fourth to fifth year as per previous experience (Wiklander et al., 1991; Jacks et al, 1997).
4. Conclusions The runoff was primarily formed by passage through the surface peat while smaller amounts were discharged via the underlying till. During the growth season, the reduction of nitrate was in the order of 85% (Table 1) and occurred close to the upland margin before it reached the buffer stand excluded from harvesting. Both nitrate concentrations in piezometers
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as well as redox bars of polished iron indicated that the volumes of peat where the nitrate reduction occurred was close to the upland margin and above a depth in the peat of 15–20 cm. Sulphate reduction occurred further downslope and preferably in the top portion of the peat favoured by a degradable substrate. Sulphate reduction varied between 20 and 50% during winter and during the summer, respectively. The large volumes of peat exhibiting sulphate reduction as observed by redox indicators do not correspond to the sulphate reduction in the water, which means that it is probable that sulphate reduction takes place in less permeable sections of the peat, affecting only part of the water flow while the rest of the water is subject to bypass flow. The residual forest was too narrow to be able to effectively utilise the nitrate released from the clear-cut (Fig. 12). Trees were left only in the wetland riparian area and, as most of them were spruce, they were disfavoured by the higher groundwater level after the clear-cutting. In contrast to alder, spruces do not have aerenchyma for the exchange of the root zone atmosphere (Rusch and Rennenberg, 1998). The buffer stand should be young at the time of clear-cutting in the upland if it is to have good nutrient uptake (Lundell and Albrektson, 1997). Storm felling of the loosely rooted wetland stand exposed to winds further accentuated the loss of tree uptake, which is a common observation (Jacks et al., 1997). Thus, a buffer stand extending into the till area is advisable to protect the wetland stand for winds. While buffer stands today often consists of trees that have little value or may not be reached by the harvesters without destruction of the peatland soil it seems necessary in the future to have a more planned management strategy, especially as broader buffer stands will occupy larger fractions of the catchments. While a buffer stand 10 m wide occupies about 6% of the catchment area, extending it to 20 m will increase this area to 12% (Bren, 1995).
Acknowledgements This project is part of the VASTRA programme and was partly funded by the MISTRA foundation. We are thankful to Ulf Johansson, To¨ nnersjo¨ Experimental Forest, Swedish University of Agricultural Sciences for prompt assistance and help in connection with the field work.
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G. Jacks, A.-C. Norrstro¨ m / Forest Ecology and Management 196 (2004) 187–197 Lind, B., Lundin, L., 1990. Saturated hydraulic conductivity of Scandinavian tills. Nordic Hydrol. 21, 107–118. Mandel, S., Shiftan, Z.L., 1981. Groundwater Resources Investigation and Development. Academic Press, New York, 269 pp. Matheson, F.E., Nguyen, M.L., Cooper, A.B., Burt, T.P., Bull, D.C., 2002. Fate of 15N-nitrate in unplanted, planted and harvested riparian wetland soil microcosms. Ecol. Eng. 19, 249–264. Naturva˚ rdsverket/SLU, 1999. Environmental Quality in Swedish Forests. Naturva˚ rdsverket, Stockholm, 100 pp. Nohrstedt, H.-O., Sikstro¨ m, U., Ring, E., Na¨ sholm, T., Ho¨ gberg, P., Persson, T., 1996. Nitrate in soil water in three Norway spruce stands in southwest Sweden as related to N-deposition and soil, stand and foliage properties. Can. J. For. Res. 26, 836–848. Norrstro¨ m, A.-C., 1994. Field-measured redox potentials in soils at the groundwater-surface water interface. Eur. J. Soil Sci. 45, 31–36. Norrstro¨ m, A.-C., Jacks, G., 1996. Water pathways and chemistry at the groundwater/surface water interface to Lake Skjervatjern Norway. Water Resour. Res. 32, 2221–2229. ¨ rlander, G., Karlsson, C., 2000. Influence of shelterwood density O on survival and height increment of Picea abies advance growth. Scand. J. For. Res. 15, 20–29.
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¨ rlander, G., Langvall, O., Petersson, P., Westling, O., 1997. O Nutrient Losses After Slash Removal and Soil Preparation on South Swedish Clear-cuts. Department of South Swedish Forestry, Sweden University of Agriculture Science Report 15, 15 pp. Rosenberg, R., Elmgren, R., Fleischer, S., Jonsson, P., Persson, G., Dahlin, H., 1990. Marine eutrophication studies in Sweden. Ambio 19, 102–108. Rudebeck, A., Persson, T., 1998. Nitrification in organic and mineral soil layers in coniferous forests in response to acidity. Environ. Pollut. 102 (Suppl. 1), 377–383. Rusch, H., Rennenberg, H., 1998. Black alder trees mediate methane and nitrous oxide emission from the soil to the atmosphere. Plant Soil 201, 1–7. Thelin, G., Rosengren-Brinck, U., Nihlga˚ rd, B., Barkman, A., 1998. Trends in needle and soil chemistry of Norway spruce and Scots pine stands in South Sweden 1985–1994. Environ. Pollut. 99, 149–158. Ulrich, B., 1992. Effects of acidic precipitation on forest ecosystems in Europe. In: Acidic Precipitation, vol. 2: Biological and ecological effects. Advances in Environmental Science. Springer Verlag, New York, pp. 189–272. Wiklander, G., Norlander, G., Andersson, R., 1991. Leaching of nitrogen from a forest catchment at So¨ dera˚ sen in Southern Sweden. Water Air Soil Pollut. 55, 263–282.
Hydrochemistry and hydrology of forest riparian wetlands G. Jacks*, A-C. Norrstro¨m Land and Water Resources Engineering, Royal Institute of Technology, SE-100 44, Stockholm, Sweden Received 30 July 2001; received in revised form 29 August 2003; accepted 6 January 2004
Abstract Forest stream riparian wetlands have a number of important features. This investigation treats one aspect, the nitrogen retention after upland clear-cutting which leads to elevated nitrate leaching, and the importance of the flow pathways in this connection. The runoff occurs mainly via the upper permeable section of the peat while the lower peat act as an aquitard, restricting the flow from the underlying till. The till groundwater is progressively artesian towards the discharging stream. Water analyses from piezometers show that the water chemistry in the peat is rather variable, indicating the presence of channelling. Channelling is also indicated by spring discharges from the peat that have elevated nitrate contents pointing to bypass flow. Redox bars indicating sulphate reduction display the same picture of irregular distribution. However, a general observation is that volumes with sulphate reduction increase towards the stream and that sulphate reduction occurs preferably in the surface peat, indicating the importance of a degradable substrate for the sulphate reducers. Nitrate reduction during the growth season occurs predominantly close to the upland till areas. The riparian tree stand is dominated by spruces which are likely to be disfavoured by the rising groundwater level after clear-cutting. The riparian tree stand does not extend far enough towards the upland to be benefited by the elevated nitrate flux. Buffer stands should be broader, extending into the till upland where they can utilise the leached nitrate and, more important, thanks to their deeper rooting depth protect the peatland trees against wind felling. # 2004 Elsevier B.V. All rights reserved. Keywords: Clear-cutting; Groundwater; Nitrate; Denitrification; Sulphate reduction; Riparian wetland
1. Introduction The marine waters between Sweden and Denmark have, since the mid-70s, experienced algal blooming caused by an elevated inflow of nitrogen from the surrounding land (Rosenberg et al., 1990). It is believed that most of the nitrogen comes from agricultural land, but from the Swedish side the transport from forest areas may actually be higher than that from agricultural areas and may be increasing (Fig. 1). The *
Corresponding author. Tel.: þ46-98-790-7380; fax: þ46-8-411-0775. E-mail address: [email protected] (G. Jacks).
nitrogen from forest areas contains a considerable portion of organic nitrogen which, however, is also available to marine algae (Carlsson, 1993). The forest in SW Sweden receives a nitrogen deposition of about 20 kg N ha1 per year. The removal during forest harvesting is of the order of 5–10 kg N ha1 per year and the runoff losses are about 2–5 kg N ha1 per year. Thus, about 10 kg N ha1 per year are accumulated in the soil each year. While the sulphur deposition has decreased since the 70s, there is little downward trend in nitrogen deposition (Naturva˚rdsverket/SLU, 1999). Elevated nitrogen deposition has resulted in nitrogen saturation in spruce forests in Central Europe (Katzensteiner et al., 1992). In Germany, secondary
0378-1127/$ – see front matter # 2004 Elsevier B.V. All rights reserved. doi:10.1016/j.foreco.2004.01.055
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Fig. 1. Transport of nitrogen by streams in the Lagan catchment, SW Sweden. The transport figures are normalised with respect to water flow (Joelsson and Fleischer, County Board of Holland).
magnesium deficiency has been observed (Ulrich, 1992). In SW Sweden stands exposed to high deposition have shown nitrogen saturation in the sense that nitrate leaching occurs below the root zone even during the vegetation season in intact spruce stands (Nohrstedt et al., 1996). Thelin et al. (1998) have observed sub-optimum contents of potassium and copper in spruce needles in South Sweden and a declining trend over time, indicating an approaching nutrient imbalance which may decrease the utilisation of soil nitrogen. Nitrification is promoted by a lowering of the C/N ratio in the organic horizon due to the gradual accumulation of nitrogen in the soil. A ratio below about 25 is critical according to Kriebitzsch (1978) and Gundersen et al. (1998). An additional criterion for nitrification is that the nitrogen deposition should be elevated (Dise et al., 1998). One-third of the spruce stands in SW Sweden have a ratio below that value (Fig. 2) in contrast to conditions in central Sweden (Fo¨ lster, 2000). Even though a low C/N ratio and a high deposition would indicate that nitrification should occur, nitrification may take time to induce according to Rudebeck and Persson (1998). So far only limited areas with an intact forest stand seem to be nitrogen-saturated, but large amounts of nitrogen are mineralised and mobilised from slash and soil after clear-cutting. As much as 700 kg N ha1 during 5 years after clear-cutting has been recorded as being lost below the root zone after harvest in SW ¨ rlander et al., 1997). This amount may not, Sweden (O however, reach draining water courses if there are good conditions for nitrogen retention in riparian areas. Nitrogen retention may be due to denitrification, dissimilitary nitrate reduction and plant uptake.
Uptake by residual stands has been studied in Central and Northern Sweden by Lundell and Albrektson (1997). They found an increase in growth in two stands out of five, indicating a fertilising effect of nitrate leached from a clear-cut upland. Shelter stands have been found to retain nitrogen to a ¨ rlander and Karlsson, 2000). The suclarge extent (O cession of the field layer and the mosses is much more gentle in a shelter stand than after clear-cutting (Hannerz, 1996). Unfortunately, shelter stands with spruce are not found in SW Sweden due to the risk of wind felling, so that clear felling is the general practice. This investigation has been aimed at studying nitrogen retention in riparian wetlands in SW Sweden after clear-cutting. The reasons are: High and increasing transport from forests (Fig. 1); Persisting high nitrogen deposition to forest land;
Fig. 2. C/N ratio in spruce stands in Holland (from Swedish Forest Inventory).
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Critically low C/N ratios in the humus layer in about 1/3 of the sites (Fig. 2); Large leaching losses from clear-cut areas. A further aim is to provide a basis for the management of the riparian wetlands in connection to forest harvesting.
2. Materials and methods 2.1. Field sites Two field sites were chosen for the studies, Luntoma in which the riparian wetlands were essentially natural and Ho¨ ga˚ sen in which drainage was undertaken. The areas studied were 6.4 and 2.0 ha, respectively. The field site with the natural wetland was clear-cut in 1998 to 73%, leaving untouched wetland area of 0.6 ha together with an upland spruce stand of 1.1 ha in the westernmost part of the site (Fig. 3). The wetland forest stand was 10–30 m wide. The ground layer in the riparian areas was mostly Spaghnum mosses with stands of Juncus effu´ sus. Trees were dominantly spruce with a few small birches and a few alders. The wetlands were covered with peat up to 0.8–1 m depth. The drained field site was completely clear-cut in 1998 except for a beech stand covering
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0.2 ha. The clear-cut areas had carried closed stands of Norway spruce with a ground layer of Holycomium splendens and some Polytrichum mosses. Due to the closed canopy there was virtually no field layer and most of the mosses died due to light impact after the clear-cutting. The soil was a sandy-loamy till. After clear-cutting, stems were removed and after 1 year branches and tops also as firewood. The ground layer partly disappeared due to the exposure to light and coverage by slash. During the first year, only a few herbs and grass appeared, with a coverage of only 1– 2%. Descha´ mpsia fiexuo´sa and Sene´cio vulgaris were common pioneering species after the clear-cutting. In the second year, scarification was carried out leaving 0.3 m broad flirrows and 0.3 m beds every 2.5 m. Thus, about 25% of the area was disturbed. Spruce seedlings were planted during the second year. Groundwater piezometers were installed in transects from the upland margin through the riparian wetland to the stream (Fig. 4). In the wetland area, piezometers were installed in the peat and in the underlying till. Each area was supplied with two reference piezometers in undisturbed upland areas (Fig. 4). The Ho¨ ga˚ sen area was supplied with a total of 10 piezometers while Luntorna had a total of 30 piezometers. In Ho¨ ga˚ sen two channels in the peat discharging into the ditch were also used as sampling points. Runoff was registered in V-notch dams at the
Fig. 3. Sketch map of the Luntoma catchment with installations marked.
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Fig. 4. Cross-section of the riparian wetland in the Luntoma catchment with installation of piezometers.
exit from the studied sites. The investigation is only halfway through the post-clear-cut period with elevated N-leaching. A budget concerning nitrogen for the third year after clear-cutting is presented here to further illustrate the findings regarding hydrology and hydrochemistry in the riparian areas. 2.2. Sampling and analysis Samples were taken from the piezometers every 5–6 weeks and the runoff was sampled at the same intervals in the summertime and more often during periods with high flow, notably in the autumn. Before sampling, the piezometers were pumped with a volume equal to twice the volume standing in the tube to ensure representative samples. Samples for N-species, major anions and cations and dissolved organic carbon were filtered in the field through 0.2 mm filters. The pH was also measured in the field. Groundwater levels were recorded on the sampling occasions. Redox measurements using platinum electrodes are relatively easy to carry out but they are not feasible due to the spatial variability of the redox conditions (Norrstro¨ m, 1994). Alternative methods of registering redox conditions were used in the form of iron and copper bars that were pushed into the peat. Some bars were polished while some were painted with lead oxide paint. The bars were kept in place for about 6 weeks at a time. The polished iron bars showed areas with oxygen characterised by rust and areas charac-
terised by grey deposits of magnetite (Tor-Gunnar Vinka, personal communication, Swed. Corrosion Institute). The redox potential at which magnetite is formed at a pH of about 6 is of the order of 150 mV (Garrels and Christ, 1965). Copper bars were oxidised in the top portion while they were still unchanged lower down indicating a redox-cline at about the same level (Garrels and Christ, 1965). The red-orange lead oxide-painted-bars turned black where hydrogen sulphide was formed. The patterns were copied onto sheets of plastic, cut out with scissors and weighed to assess the respective areas. These were taken as representing volumes in the peat. The redox bars were inserted in duplicate or triplicate at each site. As the formation of lead sulphide is irreversible, the bars indicate the maximum extent of sulphate reduction during the period of insertion. In Luntorna, seepage rates were measured in the stream bed using seepage meters (Norrstro¨ m and Jacks, 1996). In Ho¨ ga˚ sen, the temperature was measured on one occasion during the summer to detect channelled inflow of groundwater. Hydraulic conductivity was assessed in piezometers by ‘‘slug tests’’ (Mandel and Shiftan, 1981). These were performed by pouring water into the piezometers and measuring the recovery of the water level. This gives a measure of the hydraulic conductivity in the close neighbourhood of the peizometers. Composite samples from stands of Juncus effu´ sus were collected in October 2000 and subject to nitrogen
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isotope analysis at the Department of Forest Ecology at the Swedish University of Agricultural Sciences in Umea˚ .The nitrogen species were determined in the laboratory by a Tecator Aquatech instrument. Tot-N was determined by oxidation with K-peroxidisulphate to NO3-N. Major anions (Cl, NO3, SO42) and cations (Ca2þ, Mg2þ, Naþ, Kþ) were measured with a Dionex DX 120 ion chromatograph.
3. Results 3.1. Hydrology The hydraulic conductivity as measured by ‘‘slug tests’’ (Fig. 5) was low in the lower part of the peat (0.6–1 m depth), and intermediate in the till and at intermediate depths in the peat (0.3–0.4 m depth). It was very high in the uppermost section of the peat which was observed as the recovery after pumping. This means that the lower portion of the peat acted as an aquitard separating the till groundwater from the peat groundwater. This was manifested as artesian conditions in the till groundwater. The pressure level in the till groundwater was 4–16 cm above the peat groundwater level, with the difference increasing towards the draining stream. The transmissivity as measured by ‘‘slug tests’’ (Mandel and Shiftan, 1981) varied over several orders of magnitude in each of the compartments but the deeper peat was clearly the most impermeable (Fig. 5). The surface peat at about 0.1–0.2 m depth was so permeable, of the order of 0.001 m/s that it was impossible to observe the
Fig. 6. Temperature measurement in the drainage ditch at Ho¨ ga˚ sen indicating channelled inflow into the ditch.
recovery. This leaves two main drainage pathways, in the upper peat and via the till. The till is most permeable in its uppermost parts as shown by Lind and Lundin (1990). This is due to several factors, frost heaving, root penetration and the presence of surface till more loosely deposited on a bottom till compressed below the inland ice in glacial times. Thus, deeper pathways probably contribute only minor amounts to the runoff. The peat is channelled, manifested in a large range of hydraulic conductivities in the peat piezometers and a variable chemistry. Seepage rates were measured in the stream in the Luntoma site in November 2000 and they varied from 0.02 to 16.6 l m2 per hour ðn ¼ 9Þ. Further evidence of the channelling of the peat was the presence of two channels emptying in the slope of the ditch in the Ho¨ ga˚ sen site with a discharge varying in time from 0.1 to 2 l min1. Temperature measurements during the summer revealed an inflow of colder groundwater at two other points (Fig. 6). Piezometers installed close to each other (within a distance of 0.5 m) at depths of 0.2, 0.4 and 0.8 m in the peat showed differences in groundwater heads of up to 6 cm. This indicates a flow in closed channels at a depth within the peat. 3.2. Redox conditions
Fig. 5. Hydraulic conductivity in till, deep peat and intermediate peat obtained by ‘‘slug tests’’.
The lead-painted redox bars indicated increasing sulphate reduction volumes towards the stream in the Luntoma catchment (Fig. 7). Close to the upland, the sulphate reduction was, predominantly present in the lower section of the profile, while it was shifting
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Fig. 7. Sulphate reduction as monitored by redox bars in the riparian wetland in Luntoma catchment. LT-12 to LT-14 refer to the sites in Fig. 3.
upwards in the profile closer to the stream. The sulphate reduction at a depth close to the upland is due to the influence of more oxidising conditions close to the surface, e.g. the presence of nitrate as an electron acceptor. Closer to the stream, the availability of digestible substrate near the surface moves the sulphate reduction closer to the surface. The lower degree of sulphate reduction close to the stream in May and October profiles in Fig. 7 may be caused by the hyphoreic zone which exhibits water exchange with the stream. The redox bars in the lower parts of the wetland exhibit a spotted appearance indicating a
channelled nature of the flow in the peat, which supports the hydrological findings. The large volumes of peat exhibiting sulphate reduction does not correspond to the sulphate reduction as observed in the piezometers and runoff, about 20% in February and 50% in June (Fig. 8). The explanation is probably to be that the sulphate reduction takes place in less permeable sections of the peat and that affects only a part of the total water flow. Due to the lower dry deposition after the forest harvest there were gradually declining levels of sulphate and chloride with time in all the samples, levelling out after about 2 years. This indicates that the maximum turnover time of the upland groundwater was of that order of magnitude. In Ho¨ ga˚ sen, the redox conditions seemed on the whole to be less reducing (Fig. 9). This may be due to drainage, which has lowered the groundwater level at some distance from the stream, into the older peat layers where the organic substrate is usually more refractory (Koops et al., 1996). The cutting down of the riparian forest has left a lot of slash which does not permit any ground vegetation to develop and thus lowers the input of fresh organic matter. In the spruce needles in the slash, little degradable matter is left 2 years after the harvest (Berg et al., 1996). The uncoated iron bars showed a zone of red rust close to the groundwater level and a zone without
Fig. 8. Sulphate contents in piezometers in the February and June sampling in Luntoma catchment. Shaded areas are peat.
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Fig. 9. Sulphate reduction as monitored by redox bars in the riparian wetland in Ho¨ ga˚ sen catchment.
corrosion products below that level. This indicates an aeration cell in which liberated iron is deposited as red rust. Below these two zones, at about 15–20 cm depth there was a grey zone with a deposit of magnetite (TorGunnar Vinka, personal communication, Swed. Corrosion Institute). As already mentioned, this zone should be characterised by a redox potential of the order of 150 mV. The copper bars, used in a few numbers, were oxidised only in the top 15 cm giving a similar redox level at that depth as the iron bars.
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Thus, nitrate reduction during the growth season must occur above this zone and close to the upland margin, as is evident from the decreasing nitrate concentrations in the peat piezometers (Fig. 10) as well as the lower redox level downslope leading to sulphate reduction. The nitrate reduction during the growth season may be due to plant uptake and in this case uptake in the field layer as the tree strands were largely situated below the zone of nitrate reduction. Other possible mechanisms of nitrate reduction are denitrification and dissimilitary nitrate reduction to ammonium (DNRA). Matheson et al. (2002) have in a microcosm experiment studied the partitioning between these two pathways. They found that a microcosm with plants exhibited a dominance of denitrification while the unplanted one showed largely DNRA. The organic carbon to nitrate ratio is also important (Matheson et al., 2002), or rather the degradeability of the organic matter. A highly degradable organic matter would favour more reducing conditions leading to DNRA (Buresh and Patrick, 1981). In this case the organic matter is rather refractive, largely consisting of Spaghnum peat which would indicate that denitrification is a major pathway. If DNRA would be a major pathway there could be some loss of ammonium in
Fig. 10. Nitrate–nitrogen contents in the profile LT-12 to LT-14. Shaded areas are peat. Observation points marked with circles in the December profile are new piezometers. The arrow into the stream in the December profile is a sample from a seepage meter.
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runoff especially in winter time. However, this is not observed. The combined effect of denitrification and plant uptake in June is in the order of 85% of the input into the profile, as recorded in piezometers LT-12 to LT-14 (Fig. 10). Very little of this is uptake by trees, as the buffer stand is on the whole situated below LT-13 (Figs. 3 and 4) where the nitrate content has already decreased to a background level (Fig. 9). A sizeable fraction of the nitrate reduction is denitrification which is evident from the d15N values in composite samples of Juncus effu´ sus sampled in strips from the clear-cut to the stream (Table 1). The LT-7 to LT-9 profile of pizometers is parallel to the LT-12 to LT-14 profile, about 20 m to the southwest. It is also evident from the d15N values that the denitrification occurs near the upland margin of the wetland. During the cold season, there is a nitrate breakthrough in the upper layers of the peat all the way down to the stream in Luntoma (Fig. 10) as well as via channelled water in Ho¨ ga˚ sen, which carry water with elevated nitrate concentrations directly into the stream (Fig. 11). This is presumably a bypass flow in the peat, more or less directly from the upstream till area, as was observed in another investigation (Norrstro¨ m and Jacks, 1996). The nitrate reduction during the coldest months, January and February, was of the order of 30–50% in Luntoma (Table 1). On the whole, the nitrate reduction
Table 1 Transport of Tot-N from the clear-cut area, from intact forest and wetland and export from the catchment in the year 2000 Month January February March April May June July August September October November December
Clear-cut Tot-N (kg ha1) 24.3 18.0 12.5 9.81 5.44 9.32 6.17 7.12 10.2 22.5 36.7 29.3
Total loss/loss 191.2/41 kg ha1
Forest Tot-N (kg ha1)
Runoff Tot-N (kg ha1)
0.42 0.20 0.35 0.34 0.22 0.27 0.05 0.16 0.36 0.18 0.21 0.11
16.9 10.6 4.90 2.43 0.69 1.31 0.83 1.11 1.78 3.20 10.8 13.2
2.87/1.7
67.7/11
The lowermost row presents total loss from the respective compartments and loss per hectare.
seems to be lesser in the drained catchment at Ho¨ ga˚ sen (Table 2; Fig. 11). Anaerobicity and availability of substrate are the key factors that determine where denitrification occurs (Ettema et al., 1999). The drainage in Ho¨ ga˚ sen has decreased the contact between the groundwater and the surface soil with its more
Fig. 11. Nitrate–nitrogen contents in the profile HA-1 to HA-2. Shaded areas are peat. Observation point with arrow is a spring discharge into the ditch.
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Fig. 12. Redox-conditions during the growth season in the profile LT-12 to LT-14 inferred from redox studies and chemistry in piezometers.
degradable substrate (Koops et al., 1996). The buffer zone which remains more intact in Luntoma produces a litter which in combination with a high groundwater level promotes denitrification as well as sulphate reduction. The N-budget is so far only calculated for the Luntoma catchment and for the year 2000. The runoff is measured at the exit from Luntoma. The total runoff was 720 mm out of a total rainfall of 1200 mm. To distribute the runoff on the intact forest stand, the wetland and the clear-cut we have assumed that the runoff from the clear-cut is having 200 mm more yearly runoff than the intact stand and the wetland in conformity with the findings by Lindroth and Grip (1987) and Brandt et al. (1988). As concentration of
Table 2 d15N in Juncus effu´ sus along the section in Fig. 3 (clear-cut, LT-12 to LT-14, stream side) and a parallel section (LT-7 to L-9) and mean nitrate concentration in soil water and groundwater during the vegetation season (mean of four samplings from May to September) Site
Mg NO3-N/L
d15N
Clear-cut LT-12 LT-13 LT-14 Stream-side LT-7 LT-8 LT-9
5.1 4.4 0.17 0.14 0.23 3.57 1.16 0.36
1.03 5.46 8.18 7.61 6.60 6.36 7.19 7.31
the runoff from the clear-cut is taken the mean value of five piezometers situated at the transition between the clear-cut and the wetland. As concentration of the runoff from the intact forest stand is taken the mean value of the two reference piezometers. As the concentration in water that originates from the wetland is taken the values from samples from a small brook coming from the north in the center of the wetland. With this calculation the clear-cut at the Luntoma catchment lost 41 kg N ha1 during the third year after harvesting. Nitrogen loss from the intact stand was 1.7 kg N ha1 the same year. The losses via the stream were 11 kg N ha1 for the whole catchment. The retention was 73% over the whole year, decreasing to about half in the winter months (Table 1). The nitrogen budget for the year 2000, the third year after clear-cutting does not represent the peak leaching which is expected only in the fourth to fifth year as per previous experience (Wiklander et al., 1991; Jacks et al, 1997).
4. Conclusions The runoff was primarily formed by passage through the surface peat while smaller amounts were discharged via the underlying till. During the growth season, the reduction of nitrate was in the order of 85% (Table 1) and occurred close to the upland margin before it reached the buffer stand excluded from harvesting. Both nitrate concentrations in piezometers
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as well as redox bars of polished iron indicated that the volumes of peat where the nitrate reduction occurred was close to the upland margin and above a depth in the peat of 15–20 cm. Sulphate reduction occurred further downslope and preferably in the top portion of the peat favoured by a degradable substrate. Sulphate reduction varied between 20 and 50% during winter and during the summer, respectively. The large volumes of peat exhibiting sulphate reduction as observed by redox indicators do not correspond to the sulphate reduction in the water, which means that it is probable that sulphate reduction takes place in less permeable sections of the peat, affecting only part of the water flow while the rest of the water is subject to bypass flow. The residual forest was too narrow to be able to effectively utilise the nitrate released from the clear-cut (Fig. 12). Trees were left only in the wetland riparian area and, as most of them were spruce, they were disfavoured by the higher groundwater level after the clear-cutting. In contrast to alder, spruces do not have aerenchyma for the exchange of the root zone atmosphere (Rusch and Rennenberg, 1998). The buffer stand should be young at the time of clear-cutting in the upland if it is to have good nutrient uptake (Lundell and Albrektson, 1997). Storm felling of the loosely rooted wetland stand exposed to winds further accentuated the loss of tree uptake, which is a common observation (Jacks et al., 1997). Thus, a buffer stand extending into the till area is advisable to protect the wetland stand for winds. While buffer stands today often consists of trees that have little value or may not be reached by the harvesters without destruction of the peatland soil it seems necessary in the future to have a more planned management strategy, especially as broader buffer stands will occupy larger fractions of the catchments. While a buffer stand 10 m wide occupies about 6% of the catchment area, extending it to 20 m will increase this area to 12% (Bren, 1995).
Acknowledgements This project is part of the VASTRA programme and was partly funded by the MISTRA foundation. We are thankful to Ulf Johansson, To¨ nnersjo¨ Experimental Forest, Swedish University of Agricultural Sciences for prompt assistance and help in connection with the field work.
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