HOME HELP FEEDBACK SUBSCRIPTIONS ARCHIVE SEARCH TABLE OF CONTENTS		QUICK SEARCH:   [advanced] Author: Keyword(s): Year:  Vol:  Page:

This Article Abstract Figures Only Full Text (PDF) Free Alert me when this article is cited Alert me if a correction is posted Services Similar articles in this journal Similar articles in ISI Web of Science Similar articles in PubMed Alert me to new issues of the journal Download to citation manager Citing Articles Citing Articles via HighWire Citing Articles via ISI Web of Science (13) Citing Articles via Google Scholar Google Scholar Articles by Liikanen, A. Articles by Martikainen, P. Search for Related Content PubMed PubMed Citation Articles by Liikanen, A. Articles by Martikainen, P. Agricola Articles by Liikanen, A. Articles by Martikainen, P. Related Collections Watershed and Landscape Processes Wetlands and Aquatic Processes Surface Water Quality Air Pollution Water Pollution

TECHNICAL REPORT
Wetlands and Aquatic Processes

Gas Dynamics in Eutrophic Lake Sediments Affected by Oxygen, Nitrate, and Sulfate

Department of Environmental Sciences, Univ. of Kuopio, Research and Development Unit of Environmental Health, P.O. Box 1627, FIN-70211 Kuopio, Finland

* Corresponding author (Anu.Liikanen{at}uku.fi)

Received for publication February 26, 2001.

	ABSTRACT

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
In many freshwater ecosystems, the contents of NO^-₃ and SO^2-₄ have increased, whereas O₂ has been depleted due to the increased acid and nutrient loads. These changes may affect carbon turnover and the dynamics of the major greenhouse gases CO₂, CH₄, and N₂O. We studied the effects of O₂, NO^-₃, and SO^2-₄ availability on carbon mineralization, and fluxes of CO₂, CH₄, and N₂O in the sediments of hyper-eutrophic Lake Kevätön, Finland. Undisturbed sediment cores from the deep (9 m) and shallow (4 m) profundal were incubated in a laboratory microcosm with oxic and anoxic water flows with NO^-₃ or SO^2-₄ concentrations of 0, 30, 100, 300, and 2000 µM. The carbon mineralization rate (i.e., the sum of released CO₂–C and CH₄–C) was not affected by the oxidants. However, the oxidants did change the pathways of carbon degradation and the release of CH₄. All of the oxidants depressed CH₄ fluxes in the shallow profundal sediments, which had low organic matter content. In the deep profundal sediments rich in organic matter, the CH₄ release was reduced by O₂ but was not affected by SO^2-₄ (the effect of NO^-₃ was not studied). There was an increase in N₂O release as the overlying water NO^-₃ concentration increased. Anoxia and highly elevated NO^-₃ concentrations, associated with eutrophication, increased drastically the global warming potential (GWP) of the sedimentary gases in contrast to the SO^2-₄ load, which had only minor effects on the GWP.

Abbreviations: GWP, global warming potential • MPP, methane production potential • RQ, respiratory quotient • SOC, sediment oxygen consumption

	INTRODUCTION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
GREENHOUSE gases, carbon dioxide (CO₂), methane (CH₄), and nitrous oxide (N₂O) are produced in decomposition processes in sediments. The production of these gases is associated with the availability of oxygen (O₂), nitrate (NO^-₃), and sulfate (SO^2-₄), which are terminal electron acceptors in microbial energy production. Organic matter is degraded predominantly through oxidation reactions, where organic carbon is oxidized by O₂ (aerobic respiration), NO^-₃ (denitrification), or SO^2-₄ (sulfate reduction) to CO₂. In the absence of these oxidants, organic carbon is degraded in various fermentative processes and methanogenesis to CH₄ and CO₂. Nitrous oxide is produced in denitrification, where NO^-₃ is reduced in anaerobic conditions to N₂O and molecular nitrogen (N₂). Also, nitrification produces some N₂O when ammonium (NH⁺₄) is oxidized to NO^-₃ in the presence of O₂. In natural, unpolluted freshwater lakes, methanogenesis is generally an important anaerobic process in carbon mineralization. In freshwater environments, methanogenesis is regulated by O₂ content, because availabilities of SO^2-₄ and NO^-₃ are low and therefore not considered as major terminal electron acceptors (Capone and Kiene, 1988). However, the availability of SO^2-₄ and NO^-₃ can increase as a result of atmospheric acid deposition and nitrogen leaching (Henriksen et al., 1997). Acid deposition contains SO^2-₄ and NO^-₃, and NO^-₃ can be leached from agricultural and forest soils into surface waters. Freshwater ecosystems also receive NO^-₃ from wastewater discharge. The general trends in freshwaters presently are that NO^-₃ concentrations are increasing while SO^2-₄ concentrations are decreasing (Stoddard et al., 1999) and O₂ content is decreasing associated with eutrophication.

It is important to understand how any change in the availability of the electron acceptors O₂, NO^-₃, and SO^2-₄ would change decomposition processes and associated greenhouse gas production. Decrease in the availability of any electron acceptors may enhance the production of CH₄, since more organic carbon is available for methanogenic bacteria. On the contrary, increasing electron acceptor content depresses CH₄ production, since organic carbon is degraded through energetically more favorable oxidation reactions (Capone and Kiene, 1988). However, increasing availability of electron acceptors may accelerate total carbon mineralization in freshwaters (Wieder et al., 1990; Paludan and Blicher-Mathiesen, 1996; Clavero et al., 1997) and thus CO₂ production is increased. In addition, increased NO^-₃ content stimulates denitrification in anoxic conditions and can also increase the production of N₂O, which is a more efficient greenhouse gas than CH₄ and CO₂. Thus, the increase in NO^-₃ content has both positive and negative effects on sediment gas dynamics, and the production of CH₄ may decrease in contrast to CO₂ and N₂O production. Changes in O₂, NO^-₃, and SO^2-₄ contents have diverse effects on sediment greenhouse gas production and the total change in gas production is unknown. The gases produced in sediments can be emitted to the atmosphere. Therefore, the availability of electron acceptors can affect the impact of sediment microbial processes upon the atmosphere.

The biogeochemical processes in sediments, which are regulated by the competitive electron acceptors, O₂, NO^-₃, and SO^2-₄, are well understood (Capone and Kiene, 1988) but little is known about the changes in the total greenhouse gas balance of freshwaters with changing contents of O₂, NO^-₃, and SO^2-₄. This study shows how the availability of O₂, NO^-₃, and SO^2-₄ affects redox conditions, carbon mineralization, and the dynamics of CO₂, CH₄, and N₂O in freshwater sediments. The studies were conducted in a laboratory microcosm, which permitted good control of the environmental parameters.

	MATERIALS AND METHODS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Study Lake and Sampling
Sediments were collected from mid-boreal, hyper-eutrophic Lake Kevätön (63°6' N, 27°37' E, Finland; Huttunen et al., 2001) in summer 1999. Intact sediment cores were taken directly into the incubation cores (acrylic tube, i.d. = 94 mm, core height = 650 mm, sediment height = ca. 400 mm) with a Limnos (Turku, Finland) sediment sampler. Twenty cores were collected from a deep profundal (9-m depth) in June 1999 to study the effects of O₂ and SO^2-₄, and 18 cores were collected from a shallow profundal (4-m depth) in July 1999 to study the effects of O₂, SO^2-₄, and NO^-₃ (Table 1). Natural SO^2-₄ concentrations in a water column of Lake Kevätön varied in 1999 from 6.2 to 76 µM, and NO^-₃ concentrations from 0.14 to 39 µM.

View larger version (32K): [in this window] [in a new window]   Fig. 1. The continuous water flow system (A–H) with a gas and water tight lid (I–O). Components: reservoir (A), sampling outlet for reservoir water (B), peristaltic pump (C), sediment core (D), bottom cap (E), outflow to drain (F), water sampling with a syringe (G), water sampling with a flask (H), inflow (I), gas trap (J), outflow (K), magnetic stirrer (L), rotating magnet (M), port for electrodes (N), removable clamp (O).

Fluxes of Gases and Ions
The fluxes (µmol or mmol m^-2 d^-1) of dissolved CO₂, CH₄, N₂O, H₂S, NO^-₂ + NO^-₃, NH⁺₄, SO^2-₄, and sediment oxygen consumption (SOC) were determined from concentration differences between in- and outflowing water using flow rates and sediment surface area (69 cm²). The ebullited gas fluxes (µmol or mmol m^-2 d^-1) were calculated by dividing the amount of gas released with the incubation time and with the sediment surface area.

Bubble gases were collected with syringes from the gas traps when 3 to 10 mL of bubble gas had accumulated. The volumes and concentrations of CH₄, CO₂, N₂O, and H₂S in the bubble gases were then determined within 24 h after sampling. Water samples (20–30 mL) for dissolved gas analysis (CH₄, CO₂, N₂O, and H₂S) were taken weekly three to five times using 50-mL syringes (Terumo Europe, Leuven, Belgium) equipped with three-way stopcocks (Codan Steritex, Hoejvangen, Denmark) from in- and outflowing water immediately before and after the cores. The concentrations of dissolved gases were determined with a headspace equilibration technique within 24 h after sampling (McAuliffe, 1971; Jones and Simon, 1980). Water samples for gas analysis were preserved with sulfuric acid (1 mL H₂SO₄, 20% v/v) immediately after the sampling. Water was equilibrated in syringes with 30 to 40 mL of added nitrogen by shaking for 3 min. The concentrations of CO₂, CH₄, N₂O, and H₂S were determined from the headspace gas, and the dissolved gas concentrations in water were calculated according to Henry's law using the values from Lide and Fredrikse (1995).

Concentrations of CH₄, CO₂, and N₂O were determined with two gas chromatographs (Hewlett Packard [Palo Alto, CA] 5890 Series II) equipped with a flame ionization detector for CH₄, a thermal conductivity detector for CO₂ and CH₄ (>1000 µL L^-1 CH₄), and an electron capture detector for N₂O (Nykänen et al., 1995). Concentrations of H₂S were determined with a gas chromatograph (511-19; Analytical Instrument Development, Flemington, NJ) with a flame photometric detector (100 AT; Meloy Laboratories, Springfield, VA) using a Carboback B 60-80 (i.d. 3 mm, length 1.5 m, teflon) column (Supelco, Bellefonte, PA).

Concentrations of dissolved NO^-₂ + NO^-₃, NH⁺₄, and SO^2-₄ were measured from the samples collected from the water reservoirs (i.e., inflowing water) and from the outflows at the end of each incubation period. Samples were stored at -20°C prior to analysis. Sulfate and combined nitrite and nitrate concentrations were determined with an ion chromatograph (Dionex [Sunnyvale, CA] DX-120 with an AS 9-HC 4-mm anion column and an ASRS-ULTRA 4-mm suppressor). Ammonium was determined photometrically according to Fawcett and Scott (1960). To determine SOC, O₂ concentrations in the water reservoirs and in the outflows were measured with an oxygen electrode (Oxi 330 dissolved oxygen meter with CellOx 325 dissolved oxygen probe; WTW, Weilheim, Germany) four to seven times during the oxic incubations. Water flow rates from the cores were measured during each sampling.

Conditions in the Sediment Microcosms
At the end of each incubation period, redox potentials in the surface sediments (0.5 cm below sediment surface), as well as pH and O₂ concentration in the overlying water (1 cm above sediment surface), were measured with a Microprocessor pH320 pH meter (WTW, Germany) with a Hamilton (Bonaduz, Switerland) pH electrode and with an InLab 501 redox electrode (Mettler Toledo, Schwerzenbach, Switzerland). The O₂ electrode is described above.

Profiles of redox potential and dissolved O₂ concentration across the sediment–water interface were measured with mini-size electrodes at the end of the incubation period. Redox potentials were determined using a combination redox-potential needle electrode (1.3-mm o.d.; Diamond General Development Corp., Ann Arbor, MI) connected to a pH/mV meter (MP 120 pH meter, Mettler Toledo, Switzerland) from 2 to 6 mm above the sediment surface to a depth between 40 and 50 mm using 2- to 5-mm depth intervals. The concentration profiles of dissolved O₂ were measured only in the shallow profundal sediments after the oxic incubation with a Clark-type needle oxygen electrode (0.9-mm o.d., Model 768-20R; Diamond General Development Corp.) using a chemical microsensor (Product no. 1231; Diamond General Development Corp.). Dissolved O₂ concentrations were measured from 2 to 4 mm above the sediment surface to the depth where the dissolved O₂ was depleted using 1-mm depth intervals. The electrodes were positioned with a minimanipulator, which permitted 1-mm vertical resolution.

Properties of the Sediments
At the end of the experiments, the sediment cores were characterized visually. The deep profundal sediments were divided into layers of 0 to 2, 2 to 5, and 5 to 10 cm, and the shallow profundal sediments into layers of 0 to 1, 1 to 2, and 2 to 5 cm according to the observed color layers in the sediments. The sediments were sliced to the 1- to 5-cm layers by pushing them stepwise upward in the cores with the help of the moveable bottom cap (Fig. 1E).

Sediment methane production potentials (MPPs) were determined from one of the replicate cores from the uppermost sediment layers of each treatment and from all layers of the sediments treated with 0, 300, and 2000 µM SO^2-₄ and NO^-₃. The MPPs of the sediments were determined in triplicate 120-mL flasks with 15 mL of sediment and anaerobic headspace (modified from Saarnio et al., 1997). The flasks were sealed with rubber septa, and anoxic conditions were created by flushing them with nitrogen (99.5% N₂) for 1 min through two needles in the septa. The flasks were incubated with 42 kPa N₂ overpressure (=50 mL N₂) in the dark for 7 d at 22°C. Gas samples of 25 mL from the headspace were taken for CH₄ analysis at the beginning and at the end of the incubation. Before the sampling, the flasks were shaken vigorously to release entrapped gases from the sediment. The sediment MPP was calculated by dividing the amount of CH₄ produced in the sediments with the incubation time.

Contents of carbon, hydrogen, nitrogen, sulfur, and water in the sediments and bulk densities of the sediments were determined from the 0- to 5-cm layer from one of the replicate cores. The sulfur contents of the sediments were determined also for the SO^2-₄ treated sediments from the same layers where MPPs were analyzed (see above). Sediment carbon, hydrogen, and nitrogen contents were determined with a LECO (St. Joseph, MI) CHN-600 element analyzer (Pajunen et al., 2000) and sediment sulfur contents were determined with a LECO SC-32 element analyzer (Hall and Vaive, 1989). Analyses were done from sediment samples stored at -20°C.

Data Processing
The rate of carbon mineralization in the sediments was calculated as the sum of released CO₂–C and CH₄–C (den Hayer and Kalff, 1998). Respiratory quotients (RQ), the molar ratios of CO₂ release to O₂ consumption, were calculated for the sediments incubated with oxic water (Rich, 1975). The oxygen equivalents (O₂–e) of the various treatments were calculated. We assumed that NO^-₃ was reduced completely to N₂, and SO^2-₄ to S^2-. Thus, according to the changes in valences of N and S atoms, 1 mol of NO^-₃ = 5/4 mol of O₂ and 1 mol of SO^2-₄ = 2 (i.e., 8/4) mol of O₂. Global warming potentials (GWP, time horizon 100 yr) for CO₂, CH₄, and N₂O release from the sediments were calculated in CO₂ mass equivalents (CO₂–e, Eq. [1]) using the coefficients of 21 for CH₄ and 310 for N₂O (Intergovernmental Panel on Climate Change, 1996).

[1]

where GWP = global warming potential, mg CO₂–e m^-2 d^-1; CO₂ flux = CO₂ flux from sediment, mg CO₂ m^-2 d^-1; CH₄ flux = CH₄ flux from sediment, mg CH₄ m^-2 d^-1; and N₂O flux = N₂O flux from sediment, mg N₂O m^-2 d^-1.

The SPSS statistical package (SPSS, 1999) was used in the statistical analyses of the gas fluxes (O₂, CH₄, CO₂, and N₂O). The normal distribution of the gas fluxes was tested using the Kolmogorov–Smirnov Test. Independent sample t tests were used to compare gas fluxes in the untreated sediments (0 µM SO^2-₄ or NO^-₃) between the oxic and anoxic flow, and between the shallow and deep profundal sediments. The effects of SO^2-₄ and NO^-₃ on the gas fluxes with the oxic and anoxic flow were tested with one-way analysis of variance (ANOVA) using the Tukey-B post hoc test. If O₂ and SO^2-₄ or NO^-₃ had a co-effect on gas fluxes (P 0.05, tested with the analysis of variance) then the effects of SO^2-₄ or NO^-₃ on the gas fluxes were tested separately with the oxic and anoxic flows, otherwise results with the oxic and anoxic flows were processed together.

	RESULTS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Sediment Characteristics
Carbon, hydrogen, nitrogen, and sulfur contents were greater in the deep profundal sediments than in the shallow profundal sediments (Table 1). Water content was greater and bulk density was lower in the deep profundal sediments than in the shallow profundal sediments (Table 1).

Experimental Conditions
The properties of test waters and conditions in the sediment cores with the different treatments are presented in Table 2. The SO^2-₄ and NO^-₃ concentrations of the test water were within 13% of the treatment concentrations. The overlying water O₂ concentration was <12 µM O₂ with the anoxic flow, and >110 µM O₂ with the oxic flow. Additions of SO^2-₄ and NO^-₃ did not affect the O₂ concentrations in the overlying water. However, the highest addition of NO^-₃ increased the penetration of O₂ into the sediment by 4 mm. The average sediment O₂ penetration depths were 6, 7, 8, 9, and 10 mm in the shallow profundal sediments treated with 0, 30, 100, 300, and 2000 µM NO^-₃, respectively. Redox potentials were similar in the deep and shallow profundal sediments with the anoxic flow (Table 2; Fig. 2a,c) . Oxygen increased the redox potentials in the sediments (Table 2). Sediment redox potentials were higher with the oxic flow than with the anoxic flow down to the depth of 10 to 15 mm (Fig. 2a,b,d,e). In the shallow profundal sediments, redox potentials were higher in the NO^-₃ treated sediments than in the sediments without NO^-₃ addition (Table 2) down to a depth of 20 to 30 mm with anoxic flow (Fig. 2d), and to a depth of 8 to 45 mm with the oxic flow (Fig. 1e). Generally, SO^2-₄ did not affect the sediment redox potentials (Table 2). Only in the uppermost (0-5 mm) shallow profundal sediments were redox potentials higher in the sediments treated with 2000 µM SO^2-₄ than in the sediments without SO^2-₄ addition (Fig. 2c).

View larger version (36K): [in this window] [in a new window]   Fig. 2. Redox potential profiles in the deep profundal sediments with (a) the anoxic flow containing SO2-4, (b) the oxic flow containing SO2-4, and in the shallow profundal sediments with (c) the anoxic flow containing SO2-4, (d) the anoxic flow containing NO-3, and (e) the oxic flow containing NO-3.

View larger version (27K): [in this window] [in a new window]   Fig. 3. Gas fluxes (sum of diffusion and ebullition) from the sediments. (a) CH4 from the deep profundal, (b) CH4 from the shallow profundal, (c) CO2 from the deep profundal, (d) CO2 from the shallow profundal, (e) N2O from the deep profundal, and (f) N2O from the shallow profundal sediments. The averages and standard error of the means are shown. Asterisks denote that increase in NO-3 or SO2-4 concentration affects gas fluxes in a statistically significant manner (***, P < 0.001).

The release of CO₂ was greater in the deep profundal sediments than in the shallow profundal sediments (P < 0.001) (Fig. 3c,d) and was equal with the oxic and anoxic flows in both sediments. The addition of NO^-₃ did not affect the CO₂ fluxes in the shallow profundal sediments. The addition of SO^2-₄ had no effect on the CO₂ fluxes in the deep profundal sediments, but slightly increased the CO₂ release in the shallow profundal sediments (P < 0.001). Ebullition was not the source of CO₂.

The N₂O fluxes from the shallow profundal sediments were much greater than those from the deep profundal sediments (P < 0.001) (Fig. 3e,f). In the shallow profundal sediments, N₂O fluxes increased with increasing NO^-₃ concentration at the anoxic flow (P < 0.001). Also with the oxic flow, NO^-₃ increased the N₂O fluxes from sediments (P < 0.001), and the highest N₂O flux rates were measured with 300 µM NO^-₃. Sulfate did not affect the release of N₂O from the sediments. Ebullition was not the source of N₂O. There was no H₂S release from the sediments.

Carbon mineralization rates (i.e., the sum of released CO₂–C and CH₄–C) were greater in the deep profundal sediments than in the shallow profundal sediments (P < 0.001) (Table 3). The carbon mineralization rates did not differ between the oxic and anoxic sediments, and furthermore were not affected by NO^-₃ addition. In the deep profundal sediments, SO^2-₄ did not affect the carbon turnover, but in the shallow profundal sediments, there was a slightly higher mineralization rate with increasing SO^2-₄ concentrations (P = 0.001).

Global Warming Potentials
The GWPs of gases produced in the sediments with the different treatments are presented in Table 3. Without addition of NO^-₃ or SO^2-₄, the gas release resulted in greater GWP in the deep profundal sediments than in the shallow profundal sediments (P < 0.001, anoxic and oxic flow). The GWP of the sedimentary gases was greater with the anoxic than with the oxic flow (P < 0.001) due to the difference in CH₄ fluxes between the sediments. When more than 30 µM NO^-₃ was added, the GWP of the released gases increased drastically in the shallow profundal sediments with the anoxic flow (P < 0.001) due to increased N₂O release. In contrast, an increase in the SO^2-₄ concentration reduced slightly the GWP of gases with the anoxic flow in the deep profundal (P < 0.05) and in the shallow profundal sediments (P = 0.001) due to a decrease in CH₄ fluxes.

Fluxes of Nitrate, Sulfate, and Ammonium
The consumption of NO^-₂ + NO^-₃ and SO^2-₄ in the anoxic sediments increased with increasing NO^-₃ and SO^2-₄ concentrations in the inflowing water (Fig. 4a–d) . Consumption of SO^2-₄ was similar in both sediments. With the oxic water flow, there was some release of NO^-₂ + NO^-₃ and SO^2-₄ from the sediments.

View larger version (26K): [in this window] [in a new window]   Fig. 4. Fluxes of (a) SO2-4 from the deep profundal, (b) SO2-4 from the shallow profundal, (c) NO-2 + NO-3 from the deep profundal, (d) NO-2 + NO-3 from the shallow profundal, (e) NH+4 from the deep profundal, and (f) NH+4 from the shallow profundal sediments. The averages and standard error of the means are shown.

Properties of the Sediments After the Incubations
In the deep profundal sediments, methane production potentials (MPPs) were variable, 3.7 to 27 µmol CH₄ L^-1 h^-1, and no trends were found between the different SO^2-₄ treatments and sediment layers (Fig. 5a–d) . Only in the deep profundal sediments treated with oxic 2000 µM SO^2-₄ was MPP reduced by SO^2-₄ in the top layer of 0 to 2 cm (P < 0.001). In the shallow profundal sediments, MPP was reduced with increasing SO^2-₄ concentration in the top layer of 0 to 1 cm (P < 0.001). The MPP was greatest with 30 µM SO^2-₄ and lowest with 2000 µM SO^2-₄. In the shallow profundal sediments, MPPs were completely inhibited by all NO^-₃ concentrations in the top layer of 0 to 1 cm (P < 0.001). Also in the layers of 1 to 2 (P < 0.001) and 2 to 5 cm (P = 0.002), MPPs were considerably reduced with 300 and 2000 µM NO^-₃ treatments.

View larger version (47K): [in this window] [in a new window]   Fig. 5. Methane production potential (MPP) in the deep profundal sediments (a) after oxic SO2-4 and (b) after anoxic SO2-4 treatments, and in the shallow profundal sediments (c) after anoxic SO2-4 and (d) after oxic NO-3 treatments. The averages and standard error of the means are shown. Statistical differences between the treatments in each layers are marked with a, b, and c (Tukey-B post hoc test).

View larger version (32K): [in this window] [in a new window]   Fig. 6. Sulfur content in the sediments treated with SO2-4.

	DISCUSSION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Effects of Oxygen, Nitrate, and Sulfate in the Sediments
The electron acceptors O₂, NO^-₃, and SO^2-₄ affected sediment redox conditions and microbiology in the uppermost sediments. Oxygen was consumed within 6 to 10 mm from the sediment surface and thus was able to increase sediment redox potentials only in the uppermost sediment layers. However, NO^-₃ addition resulted in higher redox potentials with depth in the sediment profile as compared with O₂. Nitrate is also a strong oxidant, being consumed by denitrifying bacteria. In submerged paddy soils, both O₂ and NO^-₃ have been reported to disappear at redox potentials of 200 to 400 mV (Kimura, 2000). In our study, sediment redox potentials of approximately 200 mV observed to depths of 0.5 to 3 cm in the shallow profundal sediments treated with the O₂ and NO^-₃ indicated the presence of O₂ or NO^-₃. In the same layers, the sediment color changed from brown-gray to yellow-brown indicating the presence of oxidized iron. A similar color change with increasing redox potentials has been observed also by Hansen and Blackburn (1992). When both O₂ and NO^-₃ were available in the overlying water, increased redox potentials extended deeper into the sediments. Organic matter degradation in the uppermost sediments was not able to consume all of the O₂ and NO^-₃, allowing the diffusion and eventual consumption of O₂ and NO^-₃ in the deeper sediment layers. Nitrate addition increased O₂ penetration into the sediments, but it did not affect the rate of sediment oxygen consumption (SOC).

Sulfate, which is a weaker oxidant than O₂ and NO^-₃, could not increase the sediment redox potentials. Sulfate reduction has been reported to begin at -50 mV (Kimura, 2000). In the present sediments, redox potentials were sufficiently low to allow SO^2-₄ reduction to occur immediately at the sediment–water interface with the anoxic flow, and at a depth of 0.5 to 1.5 cm in the oxic, deep profundal sediments. Based on the sediment sulfur contents and black sulfide layers, SO^2-₄ is believed to have penetrated down to a depth of 5 cm in the deep profundal sediments and to 2 cm in the shallow profundal sediments. The diffusion of SO^2-₄ to lower depths was probably limited in the shallow profundal sediments, due to their higher bulk density.

Mineralization and Gas Production in the Sediments
In river sediments (Clavero et al., 1997), wetlands (Paludan and Blicher-Mathiesen, 1996), and peatlands (Wieder et al., 1990), carbon degradation has been found to occur more rapidly in oxic conditions or with increased NO^-₃ or SO^2-₄ loadings. In our study, organic carbon was mineralized in the sediments at the same rate despite the presence of the oxidants. However, pathways of carbon mineralization changed in the deep profundal sediments incubated with O₂, and in the shallow profundal sediments incubated with O₂, NO^-₃, and SO^2-₄.

Aerobic respiration was able to proceed only in the uppermost sediments, above a depth of 1 cm, where O₂ was available. The molar ratios of CO₂ release to O₂ consumption (i.e., RQs) were greater than one in the deep profundal sediments, highlighting the importance of anaerobic processes in carbon mineralization in the carbon-rich sediments (Rich, 1975). In the shallow profundal sediments, the RQ values were less than one. This different pattern in the RQ values between the deep and shallow profundal sediments is due to the difference in their CO₂ release rates, since the O₂ consumption was similar in both sediments. The aerobic degradation of plankton results in an RQ value of approximately 0.85 and organic matter of low quality yields even lower RQ values (Wetzel, 1975). In the shallow profundal sediments with low RQ values, O₂ might have been used in CH₄ oxidation, where 1.6 moles of O₂ are consumed in the production of 1 mole of CO₂ (i.e., the RQ of CH₄ oxidation is 0.625) (Joergensen and Degn, 1983). In addition, part of the produced CO₂ might have been consumed in methanogenesis or homoacetogenesis (Lay et al., 1998). Jones and Simon (1985) found high acetogenesis from CO₂ and H₂, when methanogenesis in sediments was limited by the input of organic carbon. Lay et al. (1998) reported that in freshwater sediments, homoacetogenesis was more than three times higher than methanogenesis from CO₂ and H₂. Thus, homoacetogenesis might have represented a sink for CO₂, since no significant CH₄ release in the oxic sediments was observed. Rich (1975) proposed that low RQ values represent conditions where O₂ is consumed in chemolithotrophic microbial processes, where CO₂ is not produced (e.g., in the oxidation of NH⁺₄ or other inorganic compounds). However, in the shallow profundal sediments, significant NH⁺₄ oxidation did not occur, since the NH⁺₄ fluxes were similar with the oxic and anoxic flows, and no release of NO^-₃ was observed.

Denitrification was an important sink for NO^-₃ and source of N₂O in the shallow profundal sediments incubated with the anoxic NO^-₃ flow. Only less than 3% of the NO^-₃ consumed was released from the sediments as N₂O indicating that N₂ was the main gaseous product in denitrification. An increase in overlying water NO^-₃ concentrations has been reported to increase denitrification (Andersen, 1977b), N₂O production (García-Ruiz et al., 1999), and NO^-₃ uptake (e.g., Hordijk et al., 1987) in freshwater sediments. In our study, NO^-₃ reduction and associated N₂O release increased linearly in the anoxic sediments with increasing water NO^-₃ concentration. It is believed that the penetration of NO^-₃ increased with increasing overlying water NO^-₃ concentration, allowing denitrification to extend deeper in the sediments. Thus, although the supply of organic carbon was sufficient, denitrification in sediments was limited by NO^-₃. Denitrifying bacteria were able to outcompete methanogens for organic substrates and suppress the methanogenic activity of the sediments, even at low NO^-₃ concentrations. In freshwaters, denitrification proceeds readily at 1 µM NO^-₃ (Rudd et al., 1990) and will significantly contribute to the degradation of organic matter at 140 µM NO^-₃ (Andersen, 1977a). Freshwater NO^-₃ concentrations are generally low, below 10 µM (Henriksen et al., 1997), but can occur up to 700 µM (Wetzel, 1975). With oxic flow, bacteria able to denitrify used mostly O₂ as an electron acceptor, since NO^-₃ was not consumed from overlying water, and the N₂O release rate with the oxic flow was lower than with the anoxic flow. In sediments, denitrifying bacteria have been reported to consume NO^-₃ only after all of the O₂ has been depleted (Hordijk et al., 1987). Part of the consumed NO^-₃ was probably assimilated; about 10% of the added NO^-₃ has been reported to be immobilized into microbial biomass within 48 h in lake sediments (Keeney, 1973). In addition to denitrification, NO^-₃ may be reduced to NH⁺₄ (Drake et al., 1996). In these sediments, NH⁺₄ reduction was probably not the sink for NO^-₃ since the NH⁺₄ flux from the sediments did not increase in parallel with the overlying water NO^-₃ concentrations.

The extent to which sulfate reducers were able to outcompete methanogens depended on the availabilities of SO^2-₄ and organic carbon. Sulfate-reducing bacteria were able to partly outcompete methanogens in the shallow profundal sediments with low organic carbon content. However, in the deep profundal sediments with high organic carbon contents, CH₄ release was not inhibited by SO^2-₄. The high availability of organic substrates evidently allowed activity of methanogenic bacteria even though part of their potential substrates was consumed during sulfate reduction. Methanogenesis has been reported to be inhibited by SO^2-₄ concentrations from 60 to 500 µM SO^2-₄ (Winfrey and Zeikus, 1977; Jones et al., 1982; Lovley and Klug, 1983). In freshwaters, SO^2-₄ concentration varies from 0 to 200 µM (Capone and Kiene, 1988). Based on our results, methanogenesis tolerates higher SO^2-₄ concentrations in sediments, where there is an abundant supply of organic carbon.

Gas fluxes obtained in the microcosm study do not necessarily represent absolute fluxes in situ, but indicate relative changes in the sediment gas dynamics with various availability of electron acceptor. The continuous water flow may have increased the diffusion of gases from the sediments to continuously changing water, because there is no development of stagnation and buildup of gas gradient, a normal situation in lakes, which decreases the gas diffusion rate from sediment. On the other hand, the ebullition may have been underestimated, since some gas, probably containing mainly CH₄, was entrapped in the deep profundal sediment cores and this gas production was neglected. Random release of the entrapped gas probably could cause the great variation in CH₄ fluxes from the deep profundal sediments. In the shallow profundal sediments loaded with NO^-₃, entrapped gas bubbles, probably with high concentrations of N₂ and some N₂O, were observed below the yellow-brown sediment layers.

	CONCLUSIONS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Acidification and eutrophication have increased SO^2-₄ and NO^-₃ concentrations in freshwater ecosystems in northern Europe (Henriksen et al., 1997). Presently, the NO^-₃ concentrations are still increasing, whereas SO^2-₄ concentrations are declining (Stoddard et al., 1999). In addition, eutrophication increases O₂ depletion in freshwaters. According to our study, all these environmental changes, increase in NO^-₃, and decrease in SO^2-₄ and O₂ availability, will increase the atmospheric load of sedimentary gases. Depletion of O₂ associated with eutrophication enhances emissions of CH₄. High loads of NO^-₃, more than 100 µM NO^-₃ (e.g., from agriculture or wastewaters), drastically increase global warming potential of the greenhouse gases produced in sediments. This is a result from highly increased release of N₂O, which is a radiatively 310 and 15 times more efficient greenhouse gas than CO₂ and CH₄, respectively (Intergovernmental Panel on Climate Change, 1996). With high NO^-₃ load, the atmospheric effect of increased N₂O release would not be compensated by the decrease in CH₄ release. However, with slightly increased NO^-₃ availability, here below 30 µM (e.g., from atmospheric deposition), the global warming potential of sedimentary gases decreases, since decreased CH₄ release compensates increased N₂O release. Present decrease in SO^2-₄ concentration in freshwaters, from 0 to 60 µM L^-1 during the last 20 years (Stoddard et al., 1999), can increase the CH₄ release and the atmospheric impact of sediments, especially if the sediments have low organic matter content. However, in highly organic sediments, the changes in water SO^2-₄ concentrations probably have no effect on sediment CH₄ release. The availabilities of O₂, NO^-₃, and SO^2-₄ have only minor effects on the total carbon mineralization rates in sediments, although these electron acceptors regulate the degradation pathways and associated greenhouse gas production.

	ACKNOWLEDGMENTS

 
The study was supported by the Academy of Finland. We thank North Savo Regional Environment Center for research facilities and Päivi Noponen and Henna Harju for gas and nutrient analysis.

	REFERENCES

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 

• Andersen, J.M. 1977a. Importance of the denitrification process for the rate of degradation of organic matter in lake sediments. p. 357–362. In H.L. Golterman (ed.) Proc. Int. Symp.: Interactions between Sediments and Freshwater, Amsterdam. 6 Sept. 1976. W. Junk. Center for Agric. Publ. and Documentation, Wageningen, the Netherlands.
• Andersen, J.M. 1977b. Rates of denitrification of undisturbed sediment from six lakes as a function of nitrate concentration, oxygen and temperature. Arch. Hydrobiol. 50:147–159.
• Capone, D.G., and R.P. Kiene. 1988. Comparison of microbial dynamics in marine and freshwater sediments: Contrast in anaerobic carbon catabolism. Limnol. Oceanogr. 33:725–749.
• Clavero, V., J. García-Sanchez, F.X. Niell, and J.A. Fernández. 1997. Influence of sulfate enrichment on the carbon dioxide and phosphate fluxes across the sediment–water interface. Hydrobiologia 345:59–65.
• Den Hayer, C., and J. Kalff. 1998. Organic matter mineralization rates in sediments: A within- and among-lake study. Limnol. Oceanogr. 43:695–705.
• Drake, L., N.G. Aumen, C. Kuhner, C. Wagner, A. Grießhammer, and M. Schmittroth. 1996. Anaerobic microflora of Everglades sediments: Effects of nutrients on population profiles and activities. Appl. Environ. Microbiol. 62:486–493.[Abstract]
• Fawcett, J.K., and J.E. Scott. 1960. A rapid and precise method for the determination of urea. J. Clin. Pathol. 13:156–159.
• García-Ruiz, R., S.N. Pattinson, and B.A. Whitton. 1999. Nitrous oxide production in the river Swale-Ouse, North-East England. Water Res. 33:1231–1237.
• Hall, G.E.M., and J. Vaive. 1989. Comparison of the determination of sulfur in geological materials by pyrolysis and ion chromatography with other production oriented methods. Paper 89IF:17–21. Geol. Survey of Canada, Ottawa.
• Hansen, L.S., and T.H. Blackburn. 1992. Mineralization budgets in sediment microcosms: Effect of the infauna and anoxic conditions. FEMS Microb. Ecol. 102:33–43.
• Henriksen, A., J. Mannio, A. Wilander, T. Moiseenko, R. Harriman, and J.P. Jensen. 1997. Results of national lake surveys (1995) in Finland, Norway, Sweden, Denmark, Russian Kola, Russian Karelia, Scotland and Wales. NIVA Rep. SNO 3645-97. Norwegian Inst. for Water Res., Tromso.
• Hordijk, C.A., M. Snieder, J.J.M. van Engel, and T.E. Cappenberg. 1987. Estimation of bacterial nitrate reduction rates at in situ concentrations in freshwater sediments. Appl. Environ. Microbiol. 53: 217–223.[Abstract/Free Full Text]
• Huttunen, J.T., T. Hammar, J. Alm, J. Silvola, and P.J. Martikainen. 2001. Greenhouse gases in non-oxygenated and artificially oxygenated eutrophied lakes during winter stratification. J. Environ. Qual. 30:387–394.[Abstract/Free Full Text]
• Intergovernmental Panel on Climate Change. 1996. Climate change 1995: The science of climate change. Contribution of Working Group I to the Second Assessment Report of the IPCC. Cambridge Univ. Press, Cambridge.
• Joergensen, L., and H. Degn. 1983. Mass spectrometric measurements of methane and oxygen utilization by methanotrophic bacteria. FEMS Microbiol. Lett. 20:331–335.
• Jones, J.G., and B.M. Simon. 1980. Decomposition processes in the profundal region of Belham Tarn and Lund tubes. J. Ecol. 68:493–512.
• Jones, J.G., and B.M. Simon. 1985. Interaction of acetogens and methanogens in anaerobic freshwater sediments. Appl. Environ. Microbiol. 49:944–948.[Abstract/Free Full Text]
• Jones, J.G., B.M. Simon, and S. Gardener. 1982. Factors affecting methanogenesis and associated anerobic processes in the sediments of a stratified eutrophic lake. J. General Microbiol. 128:1–11.
• Keeney, D.R. 1973. The nitrogen cycle in sediment–water systems. J. Environ. Qual. 2:15–29.
• Kimura, M. 2000. Anaerobic microbiology in waterlogged rice fields. p. 35–138. In J.M. Bollag and G. Stotzky (ed.) Soil biochemistry. Vol. 10. Marcel Dekker, New York.
• Lay, J.J., Y.Y. Li, and T. Noike. 1998. Interactions between homoacetogens and methanogens in lake sediments. J. Fermen. Bioeng. 86: 467–471.
• Lide, D.R., and H.P.R. Fredrikse (ed.) 1995. CRC handbook of chemistry and physics. 76th ed. CRC Press, Boca Raton, FL.
• Lovley, D.R., and M.J. Klug. 1983. Sulfate reducers can outcompete methanogens at freshwater sulfate concentrations. Appl. Environ. Microbiol. 45:187–192.[Abstract/Free Full Text]
• McAuliffe, C. 1971. GC determination of solutes by multiple phase equilibration. ChemTechnology 1:46–51.
• Nykänen, H., J. Alm, K. Lång, J. Silvola, and P.J. Martikainen. 1995. Emissions of CH₄, N₂O and CO₂ from a virgin fen and a fen drained for grassland in Finland. J. Biogeogr. 22:351–357.
• Pajunen, H., L. Hämäläinen, T. Kankainen, M. Korpijaakko, P. Kortelainen, O. Lehto, T. Saarinen, M. Saarnisto, J. Vuorenmaa, and T. Väänänen. 2000. Methods. p. 7–17. In H. Pajunen (ed.) Carbon in Finnish lake sediments. Special Paper 29. Geol. Survey of Finland, Kuopio.
• Paludan, C., and G. Blicher-Mathiesen. 1996. Losses of inorganic carbon and nitrous oxide from a temperate freshwater wetland in relation to nitrate loading. Biogeochemistry 35:305–326.
• Rich, P.H. 1975. Benthic metabolism of a soft-water lake. Verh. Int. Verein. Limnol. 19:1023–1028.
• Rudd, J.W.M., C.A. Kelly, D.W. Schindler, and M.A. Turner. 1990. A comparison of the acidification efficiencies of nitric and sulfuric acids by two whole-lake addition experiments. Limnol. Oceanogr. 35:663–679.
• Saarnio, S., J. Alm, J. Silvola, A. Lohila, H. Nykänen, and P.J. Martikainen. 1997. Seasonal variation in CH₄ emissions and production and oxidation potentials at microsites on an oligotrophic pine fen. Oecol. 110:414–422.
• SPSS. 1999. SPSS for Windows 9.0.1. SPSS, Chicago.
• Stoddard, J.L., D.S. Jeffries, A. Lükewille, T.A. Clair, P.J. Dillon, C.T. Driscoll, M. Forsius, M. Johannessen, J.S. Kahl, J.H. Kellogg, A. Kemp, J. Mannio, D.T. Monteith, P.S. Murdoch, S. Patrick, A. Rebsdorf, B.L. Skjelkvåle, M.P. Stainton, T. Traaen, H. van Dam, K.E. Webster, J. Wieting, and A. Wilander. 1999. Regional trends in aquatic recovery from acidification in North America and Europe. Nature 401:575–578.
• Wetzel, R.G. 1975. Limnology. 1st ed. W.B. Saunders Company, Philadelphia, PA.
• Wieder, R.K., J.B. Yavitt, and G.E. Land. 1990. Methane production and sulfate reduction in two Appalachian peatlands. Biogeochemistry 10:81–104.
• Winfrey, M.R., and J.G. Zeikus. 1977. Effect of sulfate on carbon and electron flow during microbial methanogenesis in freshwater sediments. Appl. Environ. Micorbiol. 33:275–281.[Abstract/Free Full Text]

This article has been cited by other articles:

				A. Liikanen, M. Puustinen, J. Koskiaho, T. Vaisanen, P. Martikainen, and H. Hartikainen Phosphorus Removal in a Wetland Constructed on Former Arable Land J. Environ. Qual., May 1, 2004; 33(3): 1124 - 1132. [Abstract] [Full Text] [PDF]

				R. Uusitalo and E. Turtola Determination of Redox-Sensitive Phosphorus in Field Runoff without Sediment Preconcentration J. Environ. Qual., January 1, 2003; 32(1): 70 - 77. [Abstract] [Full Text] [PDF]

This Article Abstract Figures Only Full Text (PDF) Free Alert me when this article is cited Alert me if a correction is posted Services Similar articles in this journal Similar articles in ISI Web of Science Similar articles in PubMed Alert me to new issues of the journal Download to citation manager Citing Articles Citing Articles via HighWire Citing Articles via ISI Web of Science (13) Citing Articles via Google Scholar Google Scholar Articles by Liikanen, A. Articles by Martikainen, P. Search for Related Content PubMed PubMed Citation Articles by Liikanen, A. Articles by Martikainen, P. Agricola Articles by Liikanen, A. Articles by Martikainen, P. Related Collections Watershed and Landscape Processes Wetlands and Aquatic Processes Surface Water Quality Air Pollution Water Pollution