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Published online 27 June 2007
Published in J Environ Qual 36:1163-1171 (2007)
DOI: 10.2134/jeq2006.0354
© 2007 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
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Natural Alkalinity Generation in Neutral Lakes Affected by Acid Mine Drainage

Matthias Koschorreck* and Jörg Tittel

Helmholtz Centre for Environmental Research-UFZ, Dep. of Lake Research, Brückstr. 3a, D-39114 Magdeburg, Germany

* Corresponding author (matthias.koschorreck{at}ufz.de)

Received for publication September 1, 2006.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 IN-LAKE GENERATION OF ALKALINITY
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Lakes in surface mining areas are often subject to continuous loads of acid mine drainage. The knowledge of internal alkalinity generation in a lake is necessary to predict if the lake will stay circumneutral or may acidify. The most important processes of alkalinity production in lakes are sulfate reduction, denitrification, and the burial of N in the sediment. By summarizing data from the literature, we present probable rates of these different processes in circumneutral mining lakes. The critical acidity load that can probably be compensated for by internal processes, is 5.09 mmol(–) m–2 d–1 in productive lakes and 0.50 mmol(–) m–2 d–1 in less productive lakes. Under the assumption that methanogenesis is inhibited by high sulfate concentrations, the highest probable acidity loads in such lakes are 6.85 mmol(–) m–2 d–1 and 1.06 mmol(–) m–2 d–1, respectively. Denitrification, sulfate reduction, and N burial contributed significantly to total alkalinity production. Sulfate reduction had the largest potential. However, existing models cannot predict alkalinity generation from sulfate concentrations alone because the long-term stability of reduced S compounds in the sediment is crucial for a sustainable biological alkalinity generation. The larger acid-neutralizing potential of higher trophic lakes is caused both by higher rates of microbial activity and by a greater stability of reduced reaction products in the sediment. The largest uncertainties in our knowledge with respect to the total alkalinity budget are related to microbial processes in sulfate-rich freshwater lakes and the long-term stability of reduced reaction products in the sediment.

Abbreviations: Alk, alkalinity • AMD, acid mine drainage • DIN, dissolved inorganic nitrogen • LP, less productive • NSFe(II), non-sulfidic Fe (II) • P, productive • PSR, potential sulfate reduction


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
IN-LAKE GENERATION OF ALKALINITY
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
MINING ACTIVITIES result in the aeration of anaerobic material and the oxidation of sulphidic minerals like pyrite or marcasite. Their weathering products, Fe2+ and sulfuric acid, cause the low pH and high acidity frequently observed in the drainage from these sites (Geller et al., 1998). Lakes in mining areas may receive continuous inputs of acid mine drainage (AMD) via surface and groundwater flow. Once in the lake, the oxidation and hydrolysis of Fe (II) result in the additional liberation of protons (Blodau 2006). Despite the AMD input, many lakes in mining regions are circumneutral, either because a former acidic lake was neutralized or the mine voids were flooded with neutral river water (Schultze et al., 2002). In Germany, about half of all lignite mining pit lakes are circumneutral (Klapper et al., 2001).

Lakes can compensate for acidity influx by internal alkalinity production (King et al., 1974). Even acidic mining lakes may undergo a process of "aging" and become neutral by themselves (Campbell and Lind, 1969; Peine and Peiffer, 1996), indicating the presence of alkalinity-generating processes. The amount of natural alkalinity generation determines the critical acidity load by groundwater and surface inflow into the lake. Mining authorities and environmental managers want to know if a given or future lake will stay neutral or may acidify. If the lake is to be used for fisheries and recreation purposes, it is essential to know the robustness of the alkalinity budget and the sensitivity of the lake to episodic acidity inputs.

If the acidity inflow exceeds the critical acidity load, an extremely acidic pit lake forms (Klapper and Schultze, 1995; Herlihy and Mills, 1985; Brugam et al., 1988; Castro and Moore, 2000; Pellicori et al., 2005). The pH of the lake water is often <3 and lower than the pH of the corresponding AMD (Geller et al., 2000). Such lakes are typically more mineralized than waters affected by acid rain, often by two or more orders of magnitude (Schultze and Geller, 1996). Methods for the biotechnological remediation of such lakes are currently under development (Wendt-Potthoff et al., 2002; Frömmichen et al., 2004).

During the 1980s, internal alkalinity generation in lakes was intensively studied in the context of atmospheric acidification, especially in North America and Scandinavia (Schindler, 1986). It is well known that processes in the lake sediment dominate internal alkalinity generation (Peine and Peiffer, 1996). Processes identified to be important are sulfate reduction, denitrification, and to a lesser extent N burial (Schindler, 1986). Models have been developed to quantify alkalinity production in soft water, acidified lakes (Kelly et al., 1987; Baker and Brezonik, 1988). These models were based on water residence time, mean depth, and mass transfer coefficients that describe the fluxes of sulfate, nitrate, and ammonium between water and sediment. Sulfate reduction is assumed to be first-order dependent on the sulfate concentration in the water. Because studies on atmospheric acidification focused on shallow, nutrient poor, and weakly buffered small lakes, very similar removal coefficients for sulfate were obtained across a number of lakes (Baker et al., 1986). Existing models based on soft water lakes might not be suitable for neutral lakes in mining areas, which are characterized by high sulfate concentrations and varying trophic status and depth.

The aim of this manuscript is to estimate the overall potential and the significance of individual processes of alkalinity generation in lakes. By compiling rates for different alkalinity-producing processes in various lakes, we provide estimates of probable and maximum alkalinity production in lakes of different trophic status. These data can then be compared with acidity loads to predict the probable future of a lake. We also aim to quantify the relative importance of the different processes, discuss their regulation by environmental factors, and identify gaps in our knowledge of those processes.


   IN-LAKE GENERATION OF ALKALINITY
TOP
 ABSTRACT
 INTRODUCTION
 IN-LAKE GENERATION OF ALKALINITY
RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Alkalinity, or the acid-neutralizing capacity, can be estimated in circumneutral waters as:

Formula 1[1]
with A– = organic acid anion (Stumm and Morgan, 1981; Cook et al., 1986) and all concentrations given in mmol(–) L–1 (= meq L–1). The alkalinity can be determined by an alkalinity titration or it can be calculated by the following equation considering electroneutrality:

Formula 2[2]

Alkalinity is generated if sulfate or nitrate anions are reduced and removed from the water. Other anions are of low concentration in most waters (e.g., P) or not reactive (Cl–) and were not considered here (see Other Processes below). Changes in the alkalinity budget can occur either by a mass imbalance of ions between the inflow and outflow or by processes within the lake. The present article focuses on in-lake processes. There are two possible ways to remove anions by in-lake processes: burial in the sediment or conversion to a gas, which then escapes to the atmosphere.

From a literature survey we calculated most probable and maximum rates for in-lake alkalinity-producing processes. We grouped lakes into those of high (eutrophic/hypertrophic) and low productivity (oligotrophic/mesotrophic). If possible, we followed the trophic classification from the original reference. If no information of trophic status was given, we searched the literature for other references about that particular lake at the same time. The summarized results are presented in Table 1. The original data did not show a normal distribution. To calculate the 25% and 75% quartiles, we transformed the data logarithmically, which resulted in a fairly symmetric distribution around the mean (Fig. S1 in online supplemental information). The compilation of literature data is published online as supplemental information (Tables S1GoGoGo–S5). All the processes are assumed to be taking place predominantly in the lake sediments. Rates of alkalinity generation are therefore related to sediment area (= lake area) rather than to lake volume. Rates of sulfate reduction and methanogenesis are usually orders of magnitude lower in the water column than in the sediment (Ingvorsen et al., 1981). In the anoxic hypolimnion of Lake Vechten (Netherlands), sulfate reduction consumed 8 to 19% of the C input, but the sulfide produced was re-oxidized after mixing in autumn (Hordijk et al., 1985). We assume that most of the reduced products formed in an anoxic hypolimnion (H2S, Fe2+, Mn2+, NH4+) are re-oxidized after mixing, which results in no net alkalinity gain (Cook et al., 1986). Except for denitrification, we therefore did not consider net microbial alkalinity production by reductive microbial processes in the water column to be significant for our calculations.


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  		Table 1. Contribution of autochthonous processes to alkalinity generation in lakes as summarized from the literature (detailed data in the online supplemental information). Calculations were done according to Eq. [2]. n = number of studies. The 25 to 75% range encloses 50% of measurements around the mean (means ± quartiles of log-transformed data). The net sulfate reduction refers to waters having low or moderate SO42– concentrations. PSR is the potential sulfate reduction due to inhibition of methanogenesis in sulfate-rich waters. The group of less productive lakes encompasses those having an oligotrophic or mesotrophic status compared to eutrophic and hypertrophic lakes.

 

Figure 2
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  		Figure S1. Frequency distribution of alkalinity generating processes in the dataset used for the present study. (A) log2 nitrogen burial, (B) log2 denitrification, (C) log10 net sulfate reduction and (D) log10 potential net sulfate reduction when methanogenesis is inhibited (PSR). Process rates were converted to alkalinity generation according to Eq. [2]. Labels of x axes denote class minima.

 

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  		Table S1. Nitrogen burial rates in freshwater lake sediments.

 

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  		Table S2. Sulfate reduction gross rates and sulfate concentration in lakes of different tropic status{dagger} measured with 35S tracer techniques.

 

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  		Table S3. Net sulfate reduction rates and sulfate concentrations in lakes of different trophic status.

 

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  		Table S4. Methanogenesis in lakes of different trophic status.

 

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  		Table S5. Denitrification in lakes of different trophic status.

 
Burial of Nitrogen in the Sediment
In the process of photosynthesis, alkalinity is generated by the uptake and incorporation of nitrate (assimilatory nitrate reduction). Alkalinity is only generated if nitrate is consumed. Utilization of ammonia by the autotrophs leads to an increase of acidity (Eq. [2]). The uptake of both N sources depends on their concentrations, on phytoplankton composition, and on temperature (Reay et al., 1999), and no generalizations can be made. Recycling processes in the food web and sediment diagenesis both cause N to be released back into the water column (Kuivila and Murray, 1984). Therefore, to predict long-term alkalinity generation we considered only the amount of N permanently buried in deeper sediment layers. Furthermore, only studies with N analyses in dated sediment cores were considered. Nitrogen burial rates from four studies (Table S1 in online supplementary information) were derived from organic carbon (Corg) burial using a Corg/N weight ratio of 9:1 (range 7.6–14; Dean and Gorham, 1998).

Estimates of alkalinity generation derived from N burial must be lowered by N2–fixation as fixation of N2 from the atmosphere has no net effect on the alkalinity budget. In less productive lakes, planktonic N2 fixation is insignificant and lower than 0.02 mmol m–2 d–1 (Howarth et al., 1988). In productive lakes, N2 fixation ranges between 0 and 1.80 mmol m–2 d–1 with a median of 0.10 mmol m–2 d–1. Since the potential effect on the alkalinity budget is rather low compared to rates and error ranges of other processes (Table 1), N2 fixation is neglected in our estimations. Thus, our estimate of alkalinity generation by burial of N refers to an upper limit.

The medians of N burial were 0.20 and 0.87 mmol m–2 d–1 in less productive and productive lakes, respectively (Table 1, Table S1 in online supplemental information). The median value for productive lakes was close to 1.2 mmol m–2 d–1, a value which had been calculated from N mass balances for 58 nutrient-rich Danish lakes (Jensen et al., 1990). In 46 lakes in Minnesota, a mean organic C (Corg) accumulation rate of 16 mmol C m–2 d–1 was measured (Dean and Gorham, 1998). Mulholland and Elwood (1982) reported rates of 6 and 21 mmol C m–2 d–1 in oligotrophic and meso-eutrophic lakes, respectively. Assuming a mean Corg/N mass ratio of 9:1 for sediments of autochthonous origin (Dean and Gorham, 1998), these rates correspond to N burial rates between 0.7 and 2.3 mmol N m–2 d–1, which is somewhat higher than our estimates.

Denitrification
The dissimilatory reduction of nitrate (denitrification) converts nitrate to gaseous products

Formula 3[3]

Under anoxic conditions nitrate and nitrite are rapidly converted to N2 but also to N2O and NO (Seitzinger, 1990). N2 and to a lesser extend N2O can be considered as chemically stable and not subject to re-oxidation. This means that denitrification in anoxic sediments and also in an anoxic hypolimnion produces alkalinity even if the water column becomes oxic during mixing events. NO is a minor product of denitrification and can be neglected in these considerations.

Precise measurements of in situ denitrification rates are not trivial and several methods have historically been employed. These include the acetylene inhibition technique, N2 flux method, NO3– flux method, 15N isotope pairing, and N mass-balance approaches (Steingruber et al., 2001; and references therein). Results obtained with the acetylene method were not considered here since the method tends to underestimate denitrification (Bollmann and Conrad, 1997). Most studies were performed in lakes with oxic hypolimnia, and consequently denitrification was only measured in the sediments. The medians of sedimentary denitrification were calculated to be 0.16 mmol m–2 d–1 for less productive lakes and 2.3 mmol m–2 d–1 for productive lakes, respectively (Table 1). In the review of Seitzinger (1990), rates of 0.05 to 1.4 and 0.5 to 6.3 mmol m–2 d–1 were reported for less productive and productive lakes, respectively. From N mass-balance data of 58 shallow eutrophic Danish lakes, a mean denitrification rate of 4.5 mmol m–2 d–1 (median 3.1) was calculated (Jensen et al., 1990). Our estimates are consistent with these data. Under the assumption of an anoxic hypolimnion during summer stratification and the complete denitrification of an initial nitrate concentration of 39 µmol L–1 (mean value for eutrophic lakes from Table S5 in online supplemental information), 39 mmol m–3 are denitrified each year. This corresponds to an annual mean denitrification rate of 0.1 mmol m–3 d–1. Thus, in productive lakes denitrification in a 23-m-thick anoxic hypolimnion equals the annual rate of benthic denitrification.

Sulfate Reduction and Burial of Sulfur in the Sediment
Particle-bound S can reach the sediment after uptake and incorporation by organisms. This assimilatory reduction of SO42– to S esters takes place mainly in the water column and the fixed S can then be transported down by sedimentation (Rudd et al., 1986). In addition, dissimilatory sulfate reduction converts S into gaseous H2S:

Formula 4[4]

This reaction is eventually followed by the precipitation of sulfides (Mitchell et al., 1990). Two moles of alkalinity are produced per mole of sulfate reduced. There are three possible fates of the produced sulfides and organic S components: re-oxidation, formation of metal sulfides, or formation of organically bound sulfides (Giblin et al., 1990). The net alkalinity gain is proportional to the burial of S in the sediment (= net sulfate reduction).

Gross rates of sulfate reduction in sediments have been determined in many freshwater lakes using 35S techniques. Our literature survey of 22 lakes gave rates between 0 and 354 mmol m–2 d–1 (Table S2 in online supplemental information). The rates were lower in less productive compared to productive lakes (p = 0.05, two-tailed U test). The median was calculated as 2.0 mmol m–2 d–1 (n = 13) and 6.1 mmol m–2 d–1 (n = 11) for both types of lake, respectively. These values are higher than those given in the review of Cook and Kelly (1992), probably because these authors did not differentiate between gross and net rates.

The literature database for net sulfate reduction rates was more heterogeneous across the different methods employed. Methods considered were S budget calculations, flux determination from SO42– or H2S gradients, Fe sulfide accumulation, or core incubations (Table S3 in online supplemental information). Rates of net sulfate reduction in 18 lakes ranged from 0 to 7 mmol m–2 d–1. The median was 0.07 mmol m–2 d–1 for less productive (n = 13) and 0.96 mmol m–2 d–1 for productive lakes (n = 5, Table S3 in online supplemental information). The difference between gross and net rates gives the rates of re-oxidation, which is of greater importance in less productive lakes (97% re-oxidation) compared to productive lakes (84% re-oxidation). This is in accordance with other published re-oxidation rates (Table 2).


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  		Table 2. Rates of re-oxidation of reduced S compounds in lake sediments.

 
Sulfate reduction in freshwater lakes is restricted to those upper sediment layers containing sulfate and is often limited by sulfate availability (Sinke et al., 1992; Urban et al., 1994). Measured rates are often low and vary seasonally (Bak and Pfennig, 1991); methanogenesis is the dominant terminal process of organic matter oxidation in lakes. During the mineralization of organic matter, usually the electron acceptor with the highest potential energy yield is used first. Neutralized mining lakes still contain considerable concentrations of sulfate (e.g., 2.2 mmol L–1 in Lake Senftenberg; Werner et al., 2001). Under such conditions, methanogenesis might be inhibited and electrons will be channeled to SO42– and not to CO2 (Kleeberg, 1998). If we assume that the organic matter mineralized by methanogenesis in natural lakes serves as an electron source for sulfate reduction in neutral, sulfate-rich lakes, we can calculate potential gross sulfate reduction rates from rates of methanogenesis. In our comparison of data from 21 lakes (Table S4 in online supplemental information), rates of methanogenesis ranged from 0.7 to 110 mmol m–2 d–1, and the median was 7.2 mmol m–2 d–1 (Table S4 in online supplemental information). The standard deviation was high and there was no significant relationship to the trophic status.

Under natural conditions, H2/CO2 and acetate are the dominant substrates of methanogenesis (Conrad 1999):

Formula 5[5]

Formula 6[6]
The oxidation of H2 and acetate by sulfate-reducing bacteria can be described by the following equations:

Formula 7[7]

Formula 8[8]
Regardless of the substrate, the ratio of CH4 produced to SO42– potentially reduced is 1:1. Therefore, we converted rates of gross methanogenesis obtained under low-sulfate concentrations (Table S4 in online supplementary information) to potential gross sulfate reduction rates in sulfate-rich lakes by a ratio of 1:1. We then derived potential net sulfate-reduction rates (PSR) from individual potential gross rates by assuming a re-oxidation of 97% in less productive and 84% in productive lakes (see above). The resulting median PSR is 0.56 in less productive and 1.76 mmol(–) m–2 d–1 in productive lakes, respectively (Table 1).

Iron Reduction
Alkalinity is produced when Fe is reduced and fixed in the sediment either as Fe sulfides or non-sulphidic-mineral Fe (II) (NSFe[II]) such as FeCO3 (siderite), Fe3(PO4)2 (vivianite), Fe3O4 (magnetite), or clay-bound Fe (II):

Formula 9[9]
(Vile and Wieder, 1993). In lake sediments, sulfides and siderite are probably the predominant solid phase products of Fe reduction (Postma, 1982; Davison, 1993).

Since Fe-reducing bacteria can use solid-phase Fe minerals as electron acceptors, Fe reduction can also be important in systems with low concentrations of dissolved Fe (Roden and Wetzel, 1996). The exact amount of alkalinity produced depends on the source of ferric Fe. For example, the reduction of dissolved Fe3+ is a proton-producing process, while the reduction of jarosite consumes one mole protons per mole of Fe reduced.

If Fe (II) is completely recycled due to oxidation, no net gain of alkalinity occurs (Anderson and Schiff, 1987). Thus, analogous to sulfate reduction, only the net burial of Fe (II) can be considered in an alkalinity balance. Rapid recycling of Fe (III) can lead to an active Fe cycle with a very low pool of available Fe (III) and no alkalinity gain (Sobolev and Roden, 2002).

Iron reduction has long been overlooked as an important process in sediments and, compared to other processes, has rarely been quantified. This is especially true for freshwater habitats. Studies of atmospherically acidified lakes did not consider Fe reduction because of the geochemical setting of these lakes in mainly granitic environments of low geogenic Fe. In freshwater sediments, Fe reduction is usually measured by the accumulation of Fe (II) in batch assays. In marine sediments, the standard method is to measure the total C oxidation rate and subtract the sulfate reduction rate, which is measured by 35S techniques. The remaining mineralization activity is assumed to be the sum of Fe and Mn reduction (Thamdrup 2000). Iron reduction rates vary considerable (Table S6 in online supplemental information) and may contribute significantly to C cycling in sediments (Thamdrup, 2000).


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  		Table S6. Iron reduction rates in different habitats.

 
In most cases, the predominant product is Fe sulfide. In this case, the alkalinity gain is already included in the S budget according to Eq. [10] (Anderson and Schiff 1987):

Formula 10[10]

While a lot of research has been done on the role of sulfides in sediments, not much is known about NSFe (II) formation, speciation, and stability (Thamdrup, 2000). The work of Postma (1981, 1982) showed that at low H2S concentrations, the pore water of freshwater sediments became oversaturated with respect to siderite, but the precipitation kinetics were very slow. The formation of siderite was accompanied by vivianite precipitation. Considerable amounts of NSFe(II) will only be formed if the Fe reduction rate exceeds gross sulfate reduction, which can be the case in situations with high Fe (III) input fluxes to the sediment. We conclude that Fe reduction does not have to be considered for internal lake alkalinity production. It might only be important for alkalinization under special conditions when high organic C supply meets a high Fe concentration, such as in acidic mining lake treatments (Wendt-Potthoff and Neu, 1998). Furthermore, Fe reduction may affect a lake's alkalinity budget indirectly, since it competes with sulfate reduction for electron donors and may limit alkalinity production by sulfate reduction.

Other Processes
Several other processes can alter the alkalinity of lakes. The degradation of autochthonous organic matter is associated with a loss of acidity. In three acidified Bohemian Forest lakes, alkalinity generation by organic C transformation amounted to 0.15 to 0.52 mmol(–) m–2 d–1 (Kopacek et al., 2003). These values can be regarded near to an upper estimate, as the lakes received significant dissolved organic C inputs (net in-lake retention 1.1–5.6 mg C L–1 d–1). In other studies, alkalinity generation by degradation of humic acids was insignificant (Cook et al., 1986). Therefore, we did not consider organic acids in our estimations.

Hard water lakes have often accumulated calcite-bearing sediments. The dissolution of calcite in the process of acidification will increase the alkalinity. How much stored calcite dissolves depends on the quality of the groundwater and transport processes. For example, higher rates can be expected if acidic groundwater seepage interacts with sediments in a large area, but lower rates are probable if acidic surface inflows affect only a thin sediment surface layer. High Fe concentrations in the groundwater may lead to coating and inactivation of calcite, and the persistence of calcitic sediments in acidic pit lakes is possible (Benthaus and Uhlmann, 2006). Since these processes are strongly lake specific, we have not included calcite dissolution in our estimate. In special cases, the alkalinity generation from calcite dissolution must be determined by other means, e.g., by sediment core incubations.

Macrophytes may take up nitrate, but vegetative assimilation of N is of minor importance compared to other N retention processes (Saunders and Kalff, 2001). Macrophyte N uptake is included in the N burial budget. The production of volatile S compounds might remove acidity from lakes, but the process is insignificant compared to S retention in the sediment (King et al., 1974; Cook and Kelly, 1992).


   RESULTS AND DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 IN-LAKE GENERATION OF ALKALINITY
 RESULTS AND DISCUSSION
CONCLUSIONS
 REFERENCES
 
Rates of Alkalinity Production
All processes considered were relevant for alkalinity production. Their significance depended on the lake's trophic status (Table 1). In the group of less productive lakes, N burial, denitrification, and sulfate reduction were of equal importance. High rates of re-oxidation may explain the low importance of S burial in less productive lakes. In productive lakes denitrification and sulfate reduction were one order of magnitude higher than in less productive lakes whereas N burial was only threefold higher. The sum of the medians of all single processes was 0.50 mmol(–) m–2 d–1 in less productive lakes and 5.09 mmol(–) m–2 d–1 in productive lakes. These values are probable rates of alkalinity generation and can be considered to be probable limits of acidity load in neutral lakes that can be compensated for by internal lake processes. Lake managers can take these numbers and compare them with the acidity load of a lake to predict the probable fate of that lake. A conservative estimate would use the lower boundary of the 25% quartile, which leaves a 25% risk that the real alkalinity production might be lower than that value.

The mean sulfate concentration within lakes in our literature survey was 325 µmol L–1 compared to a mean nitrate concentration of 39 µmol L–1. One may hypothesize that the relative abundance of both electron acceptors could imply that S retention might be more important than N retention. This was obviously not the case, which highlights the significance of re-oxidation in the S-cycle.

In productive lakes, denitrification was estimated to be the most important mechanism of internal N removal (Table 1), which agrees with the study of Jensen et al. (1990) who found that denitrification accounted for 77% of N removal in 58 Danish lakes. The conclusion, that denitrification is the primary mechanism of N retention in lakes (Saunders and Kalff, 2001), however, applied only to the group of productive lakes. In less productive lakes, N burial and denitrification were of comparable importance.

If the rate of alkalinity generation in a lake is to be measured and compared with the acidity load, it is important to consider possible uncertainties. Acidic groundwater is usually the major source of acidity in mining lakes. Uncertainties related to the quantification of acidity input in lakes are typically high. Independent estimates of groundwater flow can differ by more than 100% (Winter, 1981) and depend on the quality of available groundwater information and hydrogeological heterogeneity. For this reason, it appears unnecessary to measure N burial, since its contribution to alkalinity generation and variability among lakes was low. Furthermore, in young lakes, N burial rates cannot be measured.

Net sulfate reduction contributed more than 25% to total alkalinity production. Rates were highly variable and therefore cannot be predicted by simple models (see below). Since the rates of re-oxidation were high compared to gross sulfate reduction and not directly measured, S retention in the sediment could only be quantified if net sulfate reduction were determined independently. If the accumulation of S in the sediment cannot be determined (e.g., in very young lakes), it is questionable if the method for measuring gross sulfate reduction and making an empirical correction for a mean re-oxidation rate is more accurate than taking a net sulfate reduction rate from this literature survey.

Except for N burial, our estimations are based on mixed data representing either annual means or single measurements. If we exclude the latter from the dataset (5 out of 13 and 7 out of 31 references, respectively, Tables S3, S5), the recalculated medians (25–75% quartiles) are as follows: net sulfate reduction (LP) 0.10 (0.01–0.22), denitrification (LP) 0.15 (0.05–0.18), sum (LP) 0.46 (0.17–0.76), net sulfate reduction (P) 1.92 (0.18–4.2), denitrification (P) 1.49 (0.55–3.62), and sum (P) 4.28 (1.41–8.84) mmol(–) m–2 d–1. Thus, the summed alkalinity generation rates were about 10% (productive lakes) and 30% (less productive lakes) lower compared to the complete dataset.

Alkalinity production by benthic processes requires organic C as an electron donor. By comparing C sedimentation with rates of benthic denitrification, methanogenesis, and gross sulfate reduction, we can crosscheck the plausibility of our rate estimates. Carbon sedimentation equals the sum of benthic aerobic and anaerobic mineralization and C accumulation in the sediment. In 12 North American Shield lakes with low productivity (chlorophyll a = 0.5–8 mg m–3), medians of gross photosynthesis and planktonic community respiration were 24 and 13 mmol C m–2 d–1, respectively (Carignan et al., 2000). Therefore, an estimated median export of 11 mmol C m–2 d–1 from the trophogenic layers can be expected. In 39 North American lakes (chlorophyll a 0.5–19 mg m–3), the median of measured sinking fluxes amounted to 13 mmol C m–2 d–1 (Baines and Pace, 1994). Expected rates of benthic denitrification, methanogenesis (Table 1), and gross sulfate reduction (according to our literature survey, Table S2 in the online supplemental information) in less productive lakes required a C supply of 0.2, 7.1, and 4.0 mmol C m–2 d–1, respectively (Eq. [3GoGo–6]). With a C burial rate of 6.2 mmol C m–2 d–1 (Mulholland and Elwood, 1982), this summed to 17.4 mmol C m–2 d–1 and roughly matched sinking fluxes. Hence, within error ranges, productivity and vertical C flux in this category of lakes could account for the estimated alkalinity generation.

The consideration of PSR makes our estimate of alkalinity generation a maximum estimate. The sum of the medians of net sulfate reduction and PSR from our literature survey (0.35 and 1.84 mmol m–2 d–1 for less productive and productive lakes, respectively) lies within the range of literature data of net sulfate reduction in neutral sulfate-rich mining lakes. From the accumulation of acid volatile S in the sediment of five neutral mine lakes, Brugam et al. (1988) calculated a mean net sulfate reduction rate of 2.6 mmol m–2 d–1, while Peine and Peiffer (1998) calculated much lower rates between 0.005 and 0.15 mmol m–2 d–1 in three former German mining lakes. PSR is, however, a hypothetical rate that depends on the assumption of complete inhibition of methanogenesis in sulfate-rich lakes. Evidence from marine sites supports this assumption. When the PSR is included, all the individual processes summed to 1.06 mmol(–) m–2 d–1 in less productive lakes and to 6.85 mmol(–) m–2 d–1 in productive lakes, respectively. To evaluate the significance of PSR, we need more information on methanogenesis and sulfate reduction in sulfate-rich, freshwater lakes. There is recent evidence that anaerobic methane oxidation by sulfate can occur in freshwater environments (Grossmann et al., 2002; Savvichev et al., 2005; Eller et al., 2005). But if methane is oxidized by sulfate, PSR cannot be added to the sulfate reduction rate, because this would count the same reaction twice.

Regulation of Nitrogen Retention
It is well known that denitrification in the sediment depends on the nitrate concentration in the overlying water as well as on coupled nitrification–denitrification at the sediment–water interface (Steingruber et al., 2001). These interactions depend on the microscale stratification of oxygen and microbial processes at the sediment–water interface (Christensen et al., 1990; Rysgaard et al., 1994), organic matter and nitrate supply, and temperature. In most published studies, nitrate from the water column was the dominant substrate of denitrification; but at low nitrate concentrations, coupled nitrification–denitrification became important (Steingruber et al., 2001).

Due to the complexity of the system, it is not feasible to predict a lake-wide N budget from single-process studies. There are, however, several studies on N retention on an ecosystem level. Nitrogen retention is usually calculated on an areal rather than on a volumetric basis because denitrification occurs at the sediment surface (Prairie and Langvin, 1990). It has been shown that N retention in lakes depends on the N load (Saunders and Kalff, 2001):

Formula 11[11]

Denitrification depends on the concentration of dissolved inorganic N (DIN) (Höhener and Gächter, 1993):

Formula 12[12]

Nitrogen retention can also be estimated from lake area and water residence time using a lake-specific loss coefficient (Prairie and Langvin, 1990). The N concentration in the water can be modeled from the N input and the water residence time using empirical relations (Jensen et al., 1990), however, with a high uncertainty (Höhener and Gächter, 1993).

There is obviously a positive feedback between nitrate load (= acidity influx) and denitrification (= acidity removal). With respect to the alkalinity balance, it makes no sense to increase denitrification by adding nitrate to a lake. The DIN load of a lake is often not known. In such cases, typical rates of denitrification as presented in this study may be more useful than predictions using unknown parameters or complex catchment models.

Regulation of Sulfate Reduction
The internal removal coefficients of sulfate reflect limited sulfate availability at the site of sulfate reduction (Baker et al., 1986; Kelly et al., 1987). By dividing the sulfate reduction rates by the sulfate concentration (Table S3), we can calculate mass-transfer coefficients (Baker and Brezonik, 1988). The median value of 0.3 m yr–1 for the less productive lakes (n = 13) was close to that reported by Baker et al. (1986) of 0.5 m yr–1, which was attributed to diffusion limitation of sulfate reduction in the sediment. For nutrient-rich lakes, we calculated a mass-transfer coefficient of 2.0 m yr–1 (n = 5), demonstrating that the models of Baker and Kelly do not apply to these lakes. Variable mass-transfer coefficients were also observed by Giblin et al. (1990). Furthermore, the high standard deviations indicate that it is not feasible to use a simple median mass-transfer coefficient for all lakes. In the acidic Lake Anna, sulfate reduction rates were higher than the diffusive flux of sulfate to the sediment (Herlihy et al., 1987). The authors speculated that sulfate adsorbed to particles was transported to the sediment.

It has also been suggested that both sulfate and organic C are potential regulating factors for sulfate reduction in lakes (Giblin et al., 1990; Prairie and Langvin, 1990; Mitchell et al., 1990; Cook and Kelly, 1992; Dornblaser et al., 1994). Our set of literature data supports the view that sulfate reduction in lakes is not directly regulated by the sulfate concentration in the water column. We did not find any significant relationship between sulfate concentration and either gross or net sulfate reduction (Fig. 1). We conclude that, except for a special group of well-studied oligotrophic lakes, it is not feasible to predict sulfate retention in lakes from their sulfate concentration. Our working hypothesis, that neutral lakes in mining areas exhibit higher rates of alkalinity production due to their higher sulfate concentrations compared to normal lakes, is therefore not supported by the data. The data do not justify the inclusion of potential sulfate reduction rates in our estimate of alkalinity generation. Although lakes with very high sulfate concentrations tended to have high sulfate reduction rates (Herlihy et al., 1987; Skovgaard Jensen and Östergaard Andersen, 1987), it remains to be investigated under which conditions PSR rates become significant.


Figure 1
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  		Fig. 1. Gross (squares) and net (dots) sulfate reduction rates plotted against sulfate concentrations in different lakes. Data from Tables S2 and S3 (in online supplemental information). Slopes of regressions (not shown) were not significant (p = 0.50, t = 0.68 and p = 0.32, t = 1.03 for net and gross rates, respectively).

 
The high rates of re-oxidation show that the stability of reduced S compounds in the sediment is crucial for net S accumulation in the sediment (Anderson and Schiff, 1987; Peine and Peiffer, 1998). Unfortunately, there are only a few direct measurements of the oxidation of reduced S compounds in lake sediments (Jørgensen, 1990; Elsgaard and Jørgensen, 1992). Re-oxidation is usually taken as the difference between gross and net rates of sulfate reduction. Oxidation is not regulated by sulfate but by the stability of the reduced compound and by the oxidant (O2, NO3–, Fe [III], Mn) availability (Holmer and Storkholm, 2001). This is not taken into account by existing models. The oxidative S cycle in anoxic sediments is complex and includes biological as well as chemical reactions with thiosulfate (Jørgensen, 1990), elemental S, or polysulfides (Holmer and Storkholm, 2001) as intermediates.

The proportions of organic and inorganic S bound to the sediment differ considerably between lakes. This is important, because the long-term stability of different forms of bound S might vary. High organic C supply (Smith and Klug, 1981; Herlihy et al., 1988) and low Fe concentration (Carignan and Tessier, 1988) favor the formation of organic sulfides. However, not only the quantity but also the quality of organic matter might influence organic sulfide formation (Cook and Kelly, 1992). In sulfate- and Fe-rich lakes, organically bound sulfides are probably of minor importance for acidity fixation in the sediment (Herlihy et al., 1988; Herlihy and Mills, 1989; White et al., 1989). But not much is known about the stability of different sulfides. An experiment with 35SO42– addition to sediments revealed equal initial incorporation into organic and inorganic S compounds; but after 8 months, more tracer remained in the organic fraction, implying that organic sulfides were more stable than inorganic sulfides (Rudd et al., 1986). In Fe-rich sediments of mining lakes, the oxidation of H2S by Fe (III) is of special importance (Peiffer, 1994). High concentrations of Fe (III) in the sediment may promote the instability of reduced S compounds. More research is needed to determine the long-term behavior of different S compounds in sediments.

The availability of oxidants in the sediment is directly related to the respiratory processes at the sediment surface and can be modified by bioturbating animals which transport oxidized Fe or Mn to deeper sediment layers (Holmer and Storkholm, 2001) or by aquatic plants which supply oxygen from their roots (Holmer et al., 1998). High rates of organic matter supply lead to consumption of oxygen, nitrate, and Fe (III) and keep the sediment more reduced. Consequently, our estimate of the re-oxidation rate (Table 2) is lower in productive lakes. Re-oxidation is also influenced by the mixing regime of a lake (Rudd et al., 1986; Urban and Monte, 2001) and sediment permeability (Holmer and Storkholm, 2001).


   CONCLUSIONS
TOP
 ABSTRACT
 INTRODUCTION
 IN-LAKE GENERATION OF ALKALINITY
 RESULTS AND DISCUSSION
 CONCLUSIONS
REFERENCES
 
On the basis of a thorough literature survey, we conclude that in circumneutral lakes the critical acidity load that can probably be compensated for by internal lake processes is estimated to be 5.09 mmol(–) m–2 d–1 in productive lakes and 0.50 mmol(–) m–2 d–1 in less productive lakes. The most important processes were denitrification, sulfate reduction, and N burial in the sediment. The largest uncertainties were related to the stability of reduced S compounds in the sediment and to the rates of re-oxidation of sulfides. The prediction of the resulting net sulfate reduction in lakes remains an interesting challenge.


   ACKNOWLEDGMENTS
 
Special thanks to Katrin Wendt-Potthoff and Carolyn Oldham for discussion and to Walter Geller for information about some lakes.


   REFERENCES
TOP
 ABSTRACT
 INTRODUCTION
 IN-LAKE GENERATION OF ALKALINITY
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 





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