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DIVISION S-10-WETLAND SOILS

Chemical Fluxes from Sediments in Two Adirondack Wetlands

Effects of an Acid-Neutralization Experiment

^a Dep. of Geology, State Univ. of New York College at Cortland, Cortland, NY 13405 USA
^b Dep. of Civil and Environ. Engineering, Syracuse Univ., Syracuse, NY 13244 USA
^c Dep. of Art and Archaeology, Princeton Univ., Princeton, NJ 08544 USA

cirmoc{at}snycorva.cortland.edu

	ABSTRACT

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As a strategy of acid deposition mitigation, the application of neutralizing agents to hydrologic source areas has received substantial attention for the past decade. To compare mass balance–determined fluxes with field measurements at the sediment–water interface, we used benthic enclosures to determine chemical fluxes from the sediments of a reference beaver pond (no chemical treatment) and a beaver pond within the watershed of an acid-neutralization experiment (CaCO₃ treatment). Baseline O₂–consumption rates, the effects of reacidification, and the effects of CaCO₃ and CaCl₂ additions were determined. Oxygen consumption rates in pond sediments were higher in the CaCO₃–treated wetland, indicating stimulation of microbial activity and the subsequent enhancement of organic-matter decomposition. In the reference wetland, anoxia was followed by the sequential consumption of NO^-₃ and SO^2-₄, basic cation (C_B) and Fe²⁺ release, and the production of acid-neutralizing capacity (ANC), while the release of Ca²⁺ from cation-exchange sites dominated ANC in the treated wetland. Reacidification of CaCO₃–treated sediments caused an immediate increase in Al concentration in the water column, initially in the inorganic monomeric form (Al_IM), followed by increasing concentrations of the organic monomeric form (Al_OM). Hydrolysis of Al inputs from upland drainage, complexation of Al with dissolved organic carbon (DOC), and the formation of less toxic Al_OM were all observed. Our evidence reveals that these sediments may act as sinks for inputs of strong acid anions (e.g., SO^2-₄ and NO^-₃) from atmospheric deposition, and as sinks and transformation zones for Al associated with acidic upland drainage.

Abbreviations: Al_IM, inorganic monomeric form • Al_OM, organic monomeric form • ANC, acid-neutralizing capacity • C_B, basic cation • DOC, dissolved organic carbon • DO, dissolved oxygen • Exp., Experiment • SI, saturation index

	INTRODUCTION

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THE ADDITION OF ACID-NEUTRALIZING AGENTS directly to sediments of lakes, streams, or wetlands received little attention in the literature until the early 1990s. The production and consumption of ANC in freshwater systems has been estimated through mass-balance calculations based on inlet–outlet hydrology, coupled with concentrations of chemical constituents responsible for ANC (Cirmo and Driscoll, 1996; DeVito and Dillon, 1993). The representative redox reactants (electron acceptors) contributing to ANC are shown in Table 1 . Using these reactions and their stoichiometry, we calculated ANC based on the following formula and the prevalence of these particular ions in our study area:

(1)

View this table: [in this window] [in a new window]   Table 1 Acid-neutralizing capacity (ANC) generating and consuming reactions (per mole of reactant)

The dissolution of sediment-bound calcite is thought to delay reacidification of overlying waters and to act as a net ANC source for a period of time after addition (Sverdrup and Bjerle, 1983). Driscoll et al. (1989) calculated a net consumption of ANC after liming in some lakes, resulting from Al hydrolysis. Some processes, such as microbial and benthic faunal activity, as well as sediment cation exchange, may serve to sequester additional Ca²⁺ in sediments (Young et al., 1989). Burial of calcite and surface deactivation by organic material, Fe oxides, or FeCO₃ (Sverdrup et al., 1984) limits further dissolution. The potential for increased benthic or sediment microbial activity after liming is supported by studies in Scandinavia (Ivarson, 1977; Gahnstrom, 1988). The redox status of sediments has also been shown to be critical to specific chemical-flux rates. Davison and Woof (1990), using steady-state (flow-through) sediment–water reactors, observed that O₂ was quickly depleted from water overlying anoxic sediments. Calcium release dominated under oxic conditions, while SO^2-₄ reduction and Fe²⁺ release were prevalent during anoxia. Oxygen and organic matter content, the size of electron acceptor pools (Table 1), and pH are important in determining the contribution of various chemical process to ANC. The sensitivity of macrophyte communities and Sphagnum spp. to elevated pH and liming is also well established (Hultberg and Andersson, 1982; Bukaveckas, 1988; Mackun et al., 1994). Decomposition of macrophyte litter elicited by pH changes may affect cation-exchange capacity or cause an increase in the O₂–uptake rate. This would in turn increase the spatial extent of anoxia within a wetland, altering the processes responsible for controlling ANC dynamics. Indeed, the release of CO₂ and bicarbonate from decaying organic material is known to accelerate CaCO₃ dissolution (DePinto et al., 1989).

	Chemical-Flux Determinations in Wetland Sediments

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In situ measurements of chemical flux have not been previously described for the sediments within beaver ponds or similar freshwater wetlands. The determination of the chemical-reaction rate and the sequence of electron-acceptor removal, at the sediment–water interface, are possible using sediment enclosures, laboratory sediment-core experiments, or modeling (Grenz et al., 1991; Ciceri et al., 1992). For most aquatic systems, chemical fluxes from sediments are controlled by a variety of hydrodynamic processes, including molecular and eddy diffusion and advective transport. One approach used to determine chemical flux is the direct measurement of concentration changes over time within an enclosure placed over the sediments. A drawback to this method is the effect of the accumulation of solutes in the enclosure, along with subsequent effects on additional chemical fluxes affected by the altered biogeochemical environment (e.g., redox status). This concern requires that estimations of flux rates be limited to the linear portion of measured concentration change. Sediment fluxes can also be estimated based on sediment pore water profiles through the benthic boundary layer into the overlying water column, assuming conservative diffusive flux (Maran et al., 1995).

A somewhat more integrated estimate of chemical-flux rates from wetland sediments (or, more appropriately, whole wetlands) can be obtained by calculating the chemical mass flux using the continuous monitoring of surface and subsurface water hydrology, precipitation chemistry, and relevant chemical concentrations related to these hydrologic fluxes (i.e., the mass-balance approach). Chemical flux–rate calculations made in this way may represent overall chemical input or output to a wetland but do not reveal information specific to in situ chemical-flux rates. Mass-balance estimates may in fact mask the inherent heterogeneity in flux rates for any specific site within the wetland (Cirmo and Driscoll, 1993). In addition, this type of total flux determination is prone to a propagation of errors within the calculation of a mass balance, which depends on calculations of an overall water budget before calculation of flux (DeVito and Dillon, 1993). Chemical and physical processes within the diffusive boundary layer between the sediments and water column can greatly affect chemical fluxes from sediments. In turn, chemical fluxes inferred from concentration–depth profiles in standing water, through laboratory pore water investigations or through modeling studies, may be a pragmatic approach to estimating a parameter that is difficult to measure in situ (Ciceri et al., 1992; Miller-Way et al., 1994).

	Study Objectives

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An estimate of the in situ boundary-layer solute flux in wetlands within acidic catchments may prove useful in understanding the capacity of sediments to be sources or sinks of chemical solutes associated with acid deposition and the potential buffering of upland drainage inputs. In turn, the effects of a neutralizing agent on intrinsic flux rates and the overall impact on ANC generation within sediments are relatively unstudied. We chose to address the aforementioned issues by comparing the in situ chemical flux rates (both uptake and release) of constituents related to ANC production and consumption (Eq. [1], Table 1) from the sediments in a relatively undisturbed beaver pond, with those in a beaver pond that had experienced the addition of an acid neutralizing agent (pelletized CaCO₃). Our results may be indicative of the overall comparison of the in situ chamber measurement method with a mass-balance, hydrological approach. These results should be useful (i) for interpreting the effects of such treatment on innate sediment fluxes and processes and on the actual intrinsic chemical changes elicited on-site, and (ii) for comparing the estimation of areal flux rates based on both in situ and mass-balance estimates.

	Methods and materials

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Study Sites
In this study, chemical fluxes were observed in two separate beaver ponds and associated wetlands in the Adirondack Mountains of New York, during two successive summer periods. Both wetlands are located in the south-central region of the Adirondack Mountains of northern New York State (Fig. 1) . The treatment wetland is located in the Woods Lake watershed (42°52' N, 71°58' W) and has been extensively studied as part of the Experimental Watershed Liming Study (EWLS; Driscoll et al., 1996). The biogeochemical dynamics and chemical mass fluxes from this pond have been described for both pre- and post-liming periods (Cirmo and Driscoll, 1996). The reference wetland is located within the drainage system of Pancake-Hall Creek (43°50' N, 74°52' W) 15 km southeast of Woods Lake. An 8-yr record of stream chemistry and hydrology has been established at this site (Goldstein et al., 1987; Driscoll et al., 1987; Cirmo and Driscoll, 1993). The watershed and water column of this beaver pond have not been disturbed and are assumed to be representative of beaver ponds in the region.

View larger version (19K): [in this window] [in a new window]   Fig. 1 Location of the CaCO3–treated beaver pond (Woods Lake) and the reference beaver pond in the Adirondack Mountain region of New York

Benthic Enclosures
The experimental design involved the use of static benthic enclosures. Enclosures were made from 35.5-cm-high sections of clear PVC cylinder (60 cm diam., 0.7 cm thick) with an effective volume of 100 L. The top of the enclosure was covered with a cemented PVC plate containing two one-way valves, one for injection and one for sampling. Each enclosure housed a 0.3-A DC stirrer motor enclosed in a watertight housing on top of the enclosure. This motor rotated a PVC stirring paddle at a rate of 1.5 rpm. Exposed sediment surface area within each enclosure was 0.28 m². The bottom of the enclosure was placed 2.5 cm into the sediment (up to a set of stabilizing flanges) to effectively isolate the sediment interactions to those occurring below the enclosed water column. A one-way bubble hole in the top of the enclosure allowed air to escape while the device was lowered into place. Water samples (500 mL) were collected with a hand pump through a flexible PVC tube from the one-way sampling port. The sampled volume was replaced by water entering the one-way injection valve (<1% of total enclosure volume per sampling). Sampling frequency and strategy were determined by initial determinations of the O₂–consumption rate during preliminary experiments conducted in June 1990. Each enclosure was sampled at 2- to 4-h intervals except during the evening hours, when sampling was conducted at 6- to 8-h intervals. Since near-sediment surface processes dominate sediment–water solute exchange, a slow rate of mixing slightly reduces the benthic boundary layer separating oxic from anoxic conditions. A slow, constant rate of mixing of the overlying water column was used to approximate eddy diffusion conditions, while allowing minimal resuspension of sediment particles (Glud et al., 1996). Experimental manipulations of the water column and sediments within the enclosures were performed to simulate the effects of various perturbations on chemical flux rates.

Chemistry and Data Analysis
Standard chemical methods were used for all analyses and quality assurance–quality control (QA–QC) procedures were followed, which included sample and analytical replicates, blanks, and blind audit samples, as described in Driscoll et al. (1996). We used a modification of the o-phenanthroline method for field-fixation of Fe²⁺, with use of a portable spectrophotometer. Dissolved O₂ was fixed in the field and titrated after collection using a modified Winkler titration. Measurements of pH were done on-site.

The chemical equilibrium model ALCHEMI (version 4.0; Schecher and Driscoll, 1995) was used to calculate Al equivalency (Alⁿ⁺) based on analytical measurements of total monomeric Al (Al_TM) and organic monomeric Al (Al_OM). Inorganic monomeric Al (Al_IM) was calculated as the difference between these two measurements. Water column chemistry was evaluated with regard to the solubility (saturation index [SI]) of several aluminum hydroxide mineral phases [Al(OH)₃(s)]. These included synthetic gibbsite (log*K_so = 8.11), natural gibbsite , microcrystalline gibbsite , and amorphous aluminum hydroxide . Calcite saturation was also evaluated, and organic strong-acid equivalency was estimated using DOC measurements and a triprotic organic acid analog within the ALCHEMI model. The chemical equilibrium model MINEQL+ (version 2.2; Schecher and McAvoy, 1991) was used to estimate saturation with respect to the solubility of siderite , and all saturation indexes were calculated assuming a system closed to atmospheric CO₂.

At the beginning of each experiment, chemical concentrations of all constituents were determined in the water column at the site of placement of the enclosure. Subsequently, concentration time-series plots were generated for each chemical parameter, and the slope of the initial linear portion of the plot (where applicable) was calculated by a least squares fitting procedure. These slopes were then used to estimate chemical fluxes. Replication at any one site was not logistically possible for the treatments conducted because of the expense of the enclosures. Sediment-flux measurements in the beaver pond at Woods Lake were not made before the watershed CaCO₃–treatment in October 1989. These experiments did allow a qualitative site comparison of intrinsic chemical fluxes and processes controlling ANC generation and consumption.

Experimental Design
Enclosure experiments were conducted during two summers (1990 and 1991) following the watershed CaCO₃ application at Woods Lake. A preliminary test of the effects of water column and sediment isolation on the O₂–uptake rate was conducted at the Pancake-Hall Creek (reference) site in May 1990. One enclosure was placed into the edge sediments of the pond with black polyethylene sheeting covering the sediments, while the other was placed where it was exposed to the sediments. Results showed the effectiveness of the sediment in controlling water column chemistry. Based on these preliminary results, a set of experiments was conducted in the summer of 1990 involving the CaCO₃–treated (Woods Lake) and reference (Pancake-Hall Creek) beaver ponds: Experiment (Exp.) 1. These experiments were designed to compare chemical fluxes, ANC generation and consumption, and O₂–uptake rates in different pond microsites (shallow vs. deep), as well as to observe the effects of light in the shallow sites. At both CaCO₃–treated and reference ponds, one enclosure was placed in relatively deep water near the beaver dam (2.5 m deep); the other two enclosures were placed in shallow (<1 m), moss-covered areas. One of these shallow enclosures was covered with aluminum foil to exclude light, and the other was left exposed to light. This set of experiments was conducted for 6 d at each beaver pond, with background samples sampled along with enclosure contents at each sampling time.

A second set of experiments was conducted in the summer of 1991: Exp. 2. These experiments were designed to observe the effects of CaCO₃ and CaCl₂ additions followed by reacidification (at the reference site), along with reacidification experiments at the CaCO₃–treated sediments at Woods Lake. To simplify interpretation of the results, we excluded light from all enclosures and used only shallow wetland sites for the experiments. This simplification was warranted since (i) we needed to minimize the external factors that might lead to time-specific differences (e.g., sunlight on one day, clouds another), and (ii) the shallow sites represented a typical wetland depth, since both wetlands had a median depth of 0.4 to 0.6 m (Cirmo and Driscoll, 1993; Cirmo and Driscoll, 1996), the same range of depth used for the shallow chambers in all cases. At the treated wetland (Woods Lake) one enclosure was allowed to run undisturbed for the duration of the experiment without manipulation (control Exp. 2). The other two enclosures were acidified, one with HCl (0.12 M) and the other with an H₂SO₄–HNO₃ mixture (0.10 M and 0.075 M, respectively). Acid additions were made only after the enclosures reached near-anoxic conditions. Acid was added in sufficient quantities to lower the pH to 4.5, with subsequent additions used to maintain an acidic pH.

At the reference site (Pancake-Hall Creek), 75.8 mmol of CaCO₃ were added to one enclosure to simulate liming. This CaCO₃ was added as a slurry suspended in 50-mL batches and added to the enclosure during a 4-h period. The other enclosure at this site was treated with 75.8 mmol of CaCl₂ to evaluate cation exchange with the sediments. These two enclosures were reacidified 5 d after the manipulations using HCl (1.0 M) injections at regular intervals. Enclosures were sampled for a total of 8 d for all chemical parameters contributing to ANC, as well as Al fractions and H₄SiO₄.

	Results and discussion

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Oxygen Consumption Rate
Results from the preliminary experiment in June 1990, using enclosures where sediments were exposed to light, revealed similar trends in increasing pH, diurnal temperature fluctuations, and O₂ consumption. Exposure of the water column to sediments resulted in an initial O₂–consumption rate of 67.6 mmol O₂ m^-2 d^-1. Dissolved O₂ content in the sediment-covered enclosure declined slightly within the first few hours, but remained at 6.0 mg L^-1 thereafter. It was concluded that the sediments were responsible for O₂ consumption and subsequent water-column chemistry, even in the presence of incident light, which would allow photosynthetic production of O₂.

Experiment 1: Chemical Fluxes
Baseline Chemical Fluxes
Rapid O₂ uptake was evident in the shallow–dark enclosure (64.9 mmol O₂ m^-2 d^-1), similar in magnitude to that observed in the preliminary experiment (67.6 mmol O₂ m^-2 d^-1). Small diurnal fluctuations in dissolved oxygen (DO) were evident in the shallow, light enclosure, most likely due to macrophyte photosynthesis. Initial DO concentration in the deep enclosure was near detection limits, and calculation of a DO-consumption rate was not possible. Temporal increases were evident for pH, ANC, and Fe²⁺ concentration in all enclosures, the shallow–dark enclosure exhibiting the largest overall change. Sulfate consumption and NH⁺₄ production were observed only in the shallow–dark enclosure (Fig. 2) .

View larger version (39K): [in this window] [in a new window]   Fig. 2 Time trends in SO2-4, NH+4, and Fe2+ concentrations for enclosure Experiment 1 in the reference beaver pond

View larger version (41K): [in this window] [in a new window]   Fig. 3 Time trends in SO2-4, NH+4, and Fe2+ concentrations for enclosure Experiment 1 in the CaCO3–treated beaver pond

Reacidification at the CaCO₃–Treated Site (Woods Lake)
Chemical trends in the control enclosure were similar to 1990 results, with a slight decline in pH over time and an O₂–consumption rate of 62.9 mmol O₂ m^-2 d^-1 (Fig. 4) . Thermodynamic calculations revealed lower SI_FeCO3 values (-2.5) and greater undersaturation with respect to FeCO₃. Both Al_IM and Al_OM concentrations were low (<1 and <2 µmol L^-1, respectively) throughout the experiment (Fig. 5) , and Al_IM remained near saturation with respect to natural gibbsite . In addition, the increase in concentration of H₄SiO₄ (from 38–60 µmol L^-1) supports evidence of an in situ silica release under elevated pH and DOC concentration (Hiebert and Bennett, 1992). Silica release was also demonstrated with mass-balance calculations at this same wetland in a previous study (Cirmo and Driscoll, 1996).

View larger version (41K): [in this window] [in a new window]   Fig. 4 Time trends in acid-neutralizing capacity (ANC), dissolved organic carbon (DOC), and Ca2+ concentrations for enclosure Experiment 2 in the reference beaver pond

View larger version (38K): [in this window] [in a new window]   Fig. 5 Time trends in Fe2+, H4SiO4, inorganic monomeric Al (AlIM), and organic monomeric Al (AlOM) concentrations for enclosure Experiment 2 in the reference beaver pond

The release of Al upon acidification (Fig. 5) is consistent with the hypothesis that Al has been stored and immobilized by hydrolysis within the confines of the pond during the period following watershed liming (Cirmo and Driscoll, 1996). Complexation of Al_IM by organic ligands as it accumulated in the enclosure (as evidenced by a steady increase in Al_OM concentration) is a likely explanation. Inorganic Al concentration increased in response to subsequent maintenance acid additions at 110 and 140 h, but declined rapidly soon thereafter. Acid was needed at least once per d to maintain a pH between 5.0 and 5.5; and a total of 75 meq HCl m^-2 d^-1 and 87 meq HNO₃/H₂SO₄ m^-2 d^-1 were neutralized during the course of the experiment.

Liming Simulation and Acidification at Reference Site (Pancake-Hall Creek)
At the reference site, water chemistry in the unmanipulated enclosure was similar to 1990 results, including an O₂–consumption rate of 55.5 O₂ m^-2 d^-1, a steady increase in pH, and diurnally fluctuating temperature. The production of ANC at the rate of 4200 meq m^-2 yr^-1 and the release of DOC and Ca²⁺ (Fig. 6) were similar to findings from the summer of 1990. Additions of CaCO₃ and CaCl₂ elicited different responses, with pH and ANC increases noted in the CaCO₃–treated enclosure, but an initial decline in pH and ANC in the CaCl₂–treated enclosure. The decline in Ca²⁺ concentration after CaCl₂ treatment could be the result of cation exchange with sediments and Sphagnum. Additional buffering capacity associated with the addition of carbonate resulted in high ANC in the CaCO₃–treated enclosure. Since equimolar amounts of CaCO₃ and CaCl₂ were added to the enclosures (75.8 mmol Ca²⁺), a nominal concentration of >1500 µeq L^-1 would be expected with no exchange loss of Ca²⁺. The Ca ion was lost to sediments in the CaCl₂ manipulation (1100 µeq L^-1 or 55.6 mmol) compared with the CaCO₃–treated enclosure (850 µeq L^-1, 42.9 mmol) due to slow dissolution of the CaCO₃ slurry. Theoretically, higher pH in the CaCO₃–treated enclosure would favor deprotonation of exchange sites, while conversely, lower pH in the CaCl₂–treated enclosure should lower the effective cation-exchange capacity (CEC). Our results do not support this theoretical response, possibly because of the loss of active exchange by Sphagnum, due to mortality from higher pH conditions. Chemical equilibrium calculations indicated that the CaCO₃–treated enclosure approached saturation with respect to at several periods from 53 to 140 h. Indeed, a stabilization of Fe²⁺ concentration was observed in this enclosure (Fig. 7) . Even after acidification, saturation with respect to FeCO₃ was approached.

View larger version (44K): [in this window] [in a new window]   Fig. 6 Time trends in acid-neutralizing capacity (ANC), dissolved organic carbon (DOC), and Ca2+ concentrations for enclosure Experiment 2 in the CaCO3–treated beaver pond

View larger version (38K): [in this window] [in a new window]   Fig. 7 Time trends in Fe2+, H4SiO4, inorganic monomeric Al (AlIM), and organic monomeric Al (AlOM) concentrations for enclosure Experiment 2 in the CaCO3–treated beaver pond

	Conclusions

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Although the chemical flux rates calculated in this study may not represent in situ sediment fluxes for an entire wetland, their measurement allows comparisons between sites and manipulations and may assist in the determination of overall ecosystem impact of watershed and wetland liming. Baseline measurements at the Panake-Hall Creek reference site confirmed the importance of basic cation and Fe²⁺ release, and of SO^2-₄ reduction in the generation of ANC in these wetlands. Redox processes here are driven by organic matter decomposition in an anoxic environment. From the results of the enclosure experiments at the treated site, we predict the following overall effects of a CaCO₃ treatment on wetland sediments:

• Increased O₂–uptake rate and more widespread anoxic conditions
  
• Maintenance of circumneutral pH by undissolved CaCO₃ pools
  
• High Ca²⁺ concentrations
  
• Creation of circumneutral pH sinks for upland drainage water inorganic Al
  

Our findings suggest that metabolic activity of sediment bacterial communities may respond to liming with increased consumption of other electron acceptors, such as SO^2-₄ and Fe³⁺. Calcium release from CaCO₃ dissolution, and release from cation-exchange sites, became the dominant processes of ANC generation after the liming, and the overall importance of organic decomposition and the uptake of acidic anions was diminished. Reacidification of the CaCO₃–treated sediments resulted in accelerated Ca²⁺ release from undissolved CaCO₃, or from cation-exchange sites.

Vegetation and light in the shallow areas of the reference pond appear to inhibit widespread anoxic conditions through O₂ production. We observed a large diurnal range in dissolved O₂ and pH in the shallow areas of the unlimed pond. Treatment of these shallow areas with CaCO₃ seemed to enhance O₂ uptake, since photosynthetic O₂ production was not sufficient to prevent rapid development of anoxic conditions. Potential mortality of vegetation adapted to acidic conditions (e.g., Sphagnum and Utricularia purpurea Walter) could result in a decline in primary production and an increase in the heterotrophic status in the wetland.

Dissolution and transport of hydrolyzed Al, by reacidification, was demonstrated in the CaCO₃–treated sediments at Woods Lake and with CaCO₃–treated sediments within enclosures in the reference pond at Pancake-Hall Creek. The high buffering capacity of the sediments at the Wood Lake pond persisted, as reflected in rapid pH recovery of reacidified enclosures, along with a decline in the Al release rate. If this observed release of Al from reacidification holds true for the whole wetland, then the critical time for Al release in a CaCO₃–treated pond may be in the first few months after liming. During this time, a large pool of hydrolyzed Al may exist in the sediments, especially at inlet areas where acidic upland inlet streams meet circumneutral pond water.

	ACKNOWLEDGMENTS

 
Funding for this project was provided by the Electric Power Research Institute, the Empire State Electric Energy Research Corporation, the United States Fish and Wildlife Service, and Living Lakes, Inc. We wish to thank the Woods Lake Club and the Steinhausen family for access to the properties used in the study, as well as the Covewood Lodge and the Bowes family for assistance during the field component of the study. We also wish to thank B. Aulenbach, C. and D. Bowes, J. Brasher, B. Cirmo, M. Cirmo, R. Geary, D. Niehaus, and R. Rader for assistance in the field, and W. Wang for help in the laboratory at Syracuse University. This is manuscript number 199 of the Upstate Freshwater Institute.

Received for publication July 20, 1998.

	REFERENCES

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