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Soil Science Society of America Journal 66:1377-1389 (2002)
© 2002 Soil Science Society of America

DIVISION S-10—WETLAND SOILS

Seasonal Transformations of Manganese in a Palustrine Emergent Wetland

Matthew J. La Force*,a, Colleen M. Hanselb and Scott Fendorfb

a Dep. of Geosciences, San Francisco State Univ., San Francisco, CA 94132-4163
b Dep. of Geological and Environmental Sciences, Stanford Univ., Stanford, CA 94305-2115

* Corresponding author (laforce{at}sfsu.edu)


   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Naturally occurring wetlands need to be investigated to assess their role in contaminant sequestration. Accordingly, aqueous- and solid-phase analyses were conducted to monitor Mn response to changing physicochemical conditions within a Palustrine Emergent Wetland. A rhodochrosite-like phase (MnCO3) was the dominant Mn bearing phase (by mass) within the solids throughout the year. Aqueous-phase Mn concentrations were highly variable, but only two sampling sites in the interior of the wetland had pore-waters supersaturated with respect to rhodochrosite. Despite that a rhodochrosite-like material was the dominant quantity of Mn in the solids through the year, the reactivity of Mn solids, as measured by selective sequential extractions, did change seasonally. In spring, Mn was preferentially associated with the hydrofluoric acid-extractable pool, comprising ~35% of the total extractable Mn. The sodium acetate (SA)/acetic acid (AA) fraction appears elevated from spring through fall. Amorphous sulfide phases denoted by the difference between hydrochloric- and oxalic-acid extractable Mn increased during the summer months and then decreased in early fall as the site dried. Furthermore, as the site dried, Mn associations with the MnCl2 (water soluble and exchangeable) extractable phase increased significantly (P < 0.05), comprising between 26 and 43% of the total extractable Mn. Manganese removed using hydroxylamine-hydrochloride/AA (crystalline oxide) increased significantly (P < 0.05) in the summer to ~40% of total extractable Mn. It is therefore apparent that seasonal changes in temperature and water level, with associated redox status, drive changes in surficial coatings of Mn phases and thus its reactivity.

Abbreviations: AA, acetic acid • AOD, ammonium oxalate in the dark • AVS, acid volatile sulfides • CdAR, Coeur d'Arlene • EXAFS, extended x-ray absorption fine structure spectroscopy • HA, hydroxylamine-hydrochloride • ICP, inductively, coupled plasma • PVC, polyvinyl chloride • RSD, relative standard deviation • SA, sodium acetate • SSE, selective sequential extraction • XAFS, x-ray absorption fine structure • XAS, x-ray absorption spectroscopy


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 
KNOWLEDGE OF ELEMENTAL CYCLING in natural wetlands is needed for their preservation and to optimize their remediation potential. Wetlands commonly diminish the amount of suspended solids, nutrients, and heavy metals from water that migrates through them (Gambrell, 1994; Cooke, 1994; Crites et al., 1997), and they have been used as repositories for sewage sludges, petroleum wastes, industrial wastes, and mine wastes (Masscheleyn et al., 1992; Tarutis and Unz, 1995; Thomas et al., 1995). In wetlands, redox-sensitive elements such as Mn influence contaminant cycling (Murray, 1975; Sakata, 1985). Seasonally induced changes in physiochemical conditions will undoubtedly influence Mn stability and subsequent bioavailability of sequestered contaminants. Therefore, understanding the transformation of Mn in wetlands is of critical importance if wetlands are to be implemented for improving and maintaining environmental quality in contaminated regions. Accordingly, this investigation elucidates the seasonal changes of Mn in both the aqueous- and solid-phase within a Palustrine Emergent wetland.

Manganese averages 950 mg kg-1 in the earth's crust and is an essential element for plant and animal growth (Moore, 1991). Although Mn is not quantitatively a dominant element (<0.1 wt %) in the terrestrial environment, its hydrous oxides are a primary contributor in trace element attenuation (Loganathan and Burau, 1973; Murray, 1975; Loganathan et al., 1977; McKenzie, 1979; Zasoski and Burau, 1988; Fendorf and Zasoski, 1992). Manganese oxides and oxyhydroxides have a high surface area (often near 200 m2 g-1), low zero point of charge (commonly <3), exist as concretions or coatings on soil solids, are highly reactive, and are among the most powerful oxidants found in nature (Hem, 1978, 1981). Moreover, Mn oxides and oxyhydroxides have varying solubilites and are found in nature in the (IV), (III), and (II) oxidation states—Mn(II) being the reduced and most soluble species (Hem, 1980). Because of manganese's redox activity, reductive dissolution of Mn oxides can result in a release of Mn(II) and trace elements to the aqueous phase (Hem, 1978, 1980). Manganese cycling has been investigated in marine (Balzer, 1982; Burdige and Gieskes, 1983; Thomson et al., 1986; Middelburg et al., 1987; Mucci, 1988; Jakobsen and Postma, 1989; Wartel et al., 1990; Burdige, 1993) and freshwater (lacustrine) environments (Robbins and Callender, 1975; Wilson, 1980; Pederson and Price, 1982; Trefry and Presley, 1982; Friedl et al., 1997).

The purpose of this study is to define the seasonal cycling of Mn within a Palelustrine Emergent Wetland. Palustrine Wetland classification includes all nontidal wetlands situated on river floodplains or shoreward of lakes, river channels, or estuaries, which are dominated by trees, shrubs, persistent emergents, emergent mosses, or lichens (Cowardin et al., 1979). In the USA, the overwhelming majority of wetlands are Palelustrine (Cowardin et al., 1979; Fretwell et al., 1996); therefore, investigating Mn at this site will have considerable implications to other wetland ecosystems found in the USA. To accomplish our objective, we quantify aqueous and solid phase Mn concentrations; solid-phase analysis includes defining the mineralogical composition of Mn as well as the reactivity of the specific phases.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Site Description
The Cataldo Mission Flats constitute the floodplain of the Main Stem of the Coeur d'Alene (CdAR) and are located ~30 km southeast of the city of Coeur d'Alene and 50 km east of Lake Coeur d'Alene (La Force et al., 1998) (Fig. 1) . Fluvially deposited and dredged mine tailings from the Main Stem of the CdAR are situated upon the floodplain (La Force et al., 1998), and consequently the flats and wetland soils contain elevated concentrations of As (130 mg kg-1), Fe (40000 mg kg-1), Mn (1600 mg kg-1), Pb (1800 mg kg-1), and Zn (1300 mg kg-1). Flooding of the CdAR results in Palustrine Emergent Wetlands within the contaminated floodplain, one of which is the focus of this study and is hereinafter referred to as the Cataldo wetland.



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  		Fig. 1. Location map of the Coeur d'Alene wetlands.

 
Soils at the site have been mapped as Slickens which are poorly drained accumulations of medium-textured materials, separated in ore-mill operations, overlying stratified moderately fine- and fine-textured soils and organic material (USDA, 1975). The soil consists mainly of mine tailings from the CdAR mining district and alluvium from yearly overflow. From 1882 to 1960, over 72 million tons of mine wastes were released into the CdAR river system (Bennett et al., 1989).

The primary economic minerals mined in the district are galena (PbS), sphalerite (ZnS), and Ag-bearing tetrahedrite (Cu12Sb4S13). The gangue minerals within the region are chiefly quartz (SiO2), sericite [fine-grained muscovite, KAl2(AlSi3-O10(OH)2)], ferroan dolomite [Ca(Mg, Fe2+)(CO3)2], and siderite (FeCO3), but the region also hosts an abundance of calcite (CaCO3), pyrite (FeS2), chalcopyrite (CuFeS2), and pyrrhotite (Fe1-xS, x = 0–0.2) (Bennett et al., 1989; Hobbs et al., 1965). Minor gangue minerals also include the carbonates, dolomite [CaMg(CO3)2], and ankerite [Ca(Mg, Fe2+, Mn)(CO3)2]. The influence of carbonate gangue minerals is partially responsible for keeping the site buffered from the effects of acid mine drainage.

In the CdAR district, the average maximum temperature, for the years 1905 through 1999, occurs in July (29°C) and the average minimum temperature occurs in January (4°C). Additionally, the total annual precipitation averages 78 cm and annual snowfall averages 138 cm, with the majority occurring during January. Plants that occur at this site are cattails (Typha latifolia L.), small fruited bulrush (Scirpus microcarpus J. & K. Presl.), common rush (Juncus effusus L.), reed canary-grass (Phalaris arundinacea L.), and common horsetails (Equisetum arvense L.).

Sample Collection and Preparation
Cores were taken from the same three locations within the site (herein Site 5, 10, and 15) at each sampling time; each core was homogenized and triplicate subsamples were used for solid-phase analysis. Three sampling sites are situated along a transect through the deepest part of the wetland (Site 5) protruding into a dense cattail stand (Site 15). Cores were removed from the site using a 4-cm diam. polyvinyl chloride (PVC) piston coring device (La Force et al., 2000). Cores collected via piston coring were then immediately capped and placed upright into a N2 purged sample transporter (La Force et al., 2000). Upon returning to the laboratory, the cores were sectioned into 2-cm intervals in a N2 purged atmosphere glovebox (Labconco, Kansas City, MO) and each section was homogenized. The top 2 cm of the sediment is in contact with the overlying water column and is considered the most reactive and bioavailable soil fraction; therefore, we focused our solid phase analysis on this fraction of the wetland solids. Particle-size distribution was performed using pipette and centrifugation procedures (McDaniel et al., 1993); loss on ignition was utilized to determine mineral content; and pH was determined as a wet paste (USDA, 1996).

Samples were collected on 5 Mar. 1998 (late-winter), 13 May 1998 (spring), 8 July 1998 (early summer), 15 Sept. 1998 (late-summer), and 27 Nov. 1998 (fall), and 25 Jan. 1999 (early winter). Temperature measurements at the sediment-water interface from the interior of the wetland (Site 10) were monitored daily using a Hobo XT temperature logger (Ben Meadows, Atlanta, GA).

Aqueous Phase Sampling and Thermodynamic Calculations
Membrane separated diffusion cells (peepers) have been used extensively to monitor metal ions in solution (Carignan 1984; Carignan et al., 1985; Brandl and Hanselmann, 1991). Peepers used in this experiment were modified after Hesslein (1976) by placing 1-cm diam. (5-mL) cells containing sterile 0.20-µm nylon filters (Centrex, MF-5 sterile, disposable centrifugal microfilters; A. Daigger, Lincolnshire, IL) at 2-cm intervals for Mn depth profiles (La Force et al., 2000). The centrifuge tubes were filled with double-deionized water and placed in the wetland for 3 wk, at which point it was assumed that equilibrium was reached. At the time of sampling a few peepers were sacrificed to record pH and EH (mV) measurements. Direct electrode potentials (EH, relative to the standard H electrode) were determined using a Corning redox combination Pt electrode with Ag/AgCl reference electrode (Corning, Inc., Corning, NY) and pH was measured with a combination electrode (Accumet, Fischer Scientific, Santa Clara, CA) in each cell after a stable reading was obtained. Aqueous samples were removed from the peepers, acidified with concentrated trace element grade HCl, and kept on ice until metal analysis 3 h later. Samples were collected on 16 Mar. 1998 (late-winter), 13 May 1998 (spring), 18 July 1998 (summer), and 28 Nov. 1998 (fall). Peeper data from late-summer were not available as the site had dried; whereas late-winter peeper data were limited owing to sampler disturbance. At all sampling times, a field blank of double-deionized water with 2 drops of concentrated HCl was also run with the samples as a quality control; metal concentrations in the blank were negligible.

To determine if the system was supersaturated with respect to Mn carbonates the aqueous phase data (Mn [II] concentration, pH, alkalinity, temperature, and ionic strength was calculated from species concentrations) from the three sites were used in conjunction with the MINTEQA2 chemical speciation program (Allison et al., 1990). Thermodynamic data from Johnson (1982) were used for rhodochrosite solubility. Additionally, we used three levels of PCO2 (3.19 x 10-5, 3.19 x 10-4, and 3.19 x 10-3 MPa or 10-3.5, 10-2.5, and 10-1.5 atm) to span the range of measured and expected CO2 partial pressures commonly found in anaerobic environments (Stumm and Morgan, 1995).

Statistical Analysis
Peeper and selective sequential extraction (SSE) data from all three sites were analyzed using a SAS-GLM procedure and a SAS-CORR procedure (SAS Institute Inc., 1985) to determine at the 95% confidence level (P = 0.05) the correlation coefficients in seasonal trends among the three sites for a given data set.

Analytical Procedures
Trace elements were measured in the aqueous phase by inductively coupled plasma (ICP) optical emission spectrophotometry (Thermo Jarrell Ash IRIS ICP-OES; Franklin, MA) with a 10% accuracy range; quality control was checked every 15 samples. Detection limits were defined as 3{sigma}, where {sigma} is the standard deviation of 7 blanks (Klesta and Bartz, 1996). Detection limits for Mn were 0.01 mg L-1. All glass and plasticware used in the experiment were rinsed in 0.5 M HCl prior to use.

X-ray Absorption Fine Structure Analysis
X-ray absorption fine structure (XAFS) spectroscopy is ideal for identifying low concentrations of amorphous and crystalline components in soils in situ; it has been applied to speciate elements residing in natural materials (see for example, Friedl et al., 1997; O'Day et al., 1998, 2000). Solids were mounted as wet pastes in acrylic plates and sealed with Kapton polymide film (Dupont, Circleville, OH) to prevent moisture loss while minimizing x-ray absorption. Unfortunately, because of time restrictions imposed by access to the synchrotron facility, we only analyzed samples collected on 5 Mar. 1998, 24 June 1998, 17 Dec. 1998, 25 Feb. 1999, and 5 Apr. 1999 with XAFS spectroscopy.

To analyze bulk Mn solids, XAFS spectroscopy was performed on beamlines IV-1, IV-2, and IV-3 (beamline IV is an eight pole wiggler) at the Stanford Synchrotron Radiation Laboratory (SSRL), running under dedicated conditions. The ring operated at 3 GeV with a current ranging from ~100 to ~50 mA. Energy selection was accomplished using a Si (220) monochromator typically with an unfocused beam. Depending on the beamline, higher-order harmonics were either eliminated by a harmonic rejection mirror or by detuning ~60%. Absorption spectra were recorded by fluorescent x-ray production using a wide-angle ionization chamber (Lytle et al., 1984) for model compounds or a 13-element Ge semiconductor detector (Cramer et al., 1988) for unknown samples. A Cr filter along with Soller slits were used to minimize the effects of scattered primary radiation when using the fluorescent ionization chamber. Incident and transmitted intensities were measured with in-line ionization chambers. The energy range studied was -200 to +500 eV about the K-edge of Mn (6539 eV); higher energies could not be studied because of interference from the Fe absorption edge (7111 eV).

Each scan was calibrated internally by placing an elemental Mn foil between the second and third in-line ionization chambers. Between three and ten individual spectra were averaged for each sample. The reported extended x-ray absorption fine-structure (EXAFS) spectra were obtained at 11 K to reduce the thermal disorder of the specimen; all other spectra were collected at ambient temperature and pressure.

X-ray absorption spectra of model Mn compounds used for analysis included feitkechtite (ß-MnOOH), bixbyite ({alpha}-Mn2O3), rhodochrosite (MnCO3), manganite ({gamma}-MnOOH), pyrolusite (ß-MnO2), and birnessite ({delta}-MnO2). Manganese standards of pyrolusite, bixbyite, and rhodochrosite were purchased from Fischer Scientific (Pittsburgh, PA), Strem Chemicals (Newburyport, MA), and Alpha Aesar (Hertsfordshire, UK), respectively. Hausmanite, feitkechtite, and manganite were synthesized from procedures modified after Hem and Lind (1983), Hem et al. (1982), and Bricker (1965) as described in Fendorf et al. (1999); birnessite was made following procedures of Fendorf and Zasoski (1992). The purity of all mineral standards was confirmed using x-ray diffraction.

Both x-ray absorption near-edge structure (XANES) and EXAFS spectroscopies were used to identify solid-phase forms of Mn. Following data reduction, maximum first-derivative curve intensity was used to differentiate between Mn(II) and Mn(IV) in the solid-phase. Quantitative analysis was achieved using a linear combination of the standard Mn-phase spectra described above and comparing them to the unknown; a Levenberg-Marquardt least-squares algorithm in the WinXAS code (Ressler, 1997) was used to minimize the error between the unknown and reconstructed spectrum (Ressler et al., 2000). Energy shifts were constrained during fitting to <1.0 eV while minimizing the fit residual.

We used the extended portion of the XAFS spectrum (EXAFS) to confirm linear combination fits achieved during XANES analysis and decipher the structural characteristics of Mn in our unknown samples. The EXAFS spectra were analyzed using EXAFSPAK computer software (George, 1995). At least eight spectra were averaged and processed following procedures of Manning et al. (1998) and Eick and Fendorf (1997). Because we were confirming bulk characteristics of Mn within the samples, we used only the structural data for rhodochrosite, as defined by Effenberger et al. (1981) and Friedl et al. (1997), to test and confirm our XANES data; refinement of the rhodochrosite structural parameters was unwarranted owing to the fact that the unknown sample was similar in structure to rhodochrosite (see below).

Selective Sequential Extractions
Selective sequential extractions are common procedures that provide useful information on trace element partitioning within soils and sediments [see reviews by Pickering (1981), Chao (1984), and Martin et al. (1987)]. It is important to note that although sequential extractions attempt to isolate and dissolve one particular soil fraction, the efficiency of chemical extractions depends on the affinity and specificity of the extracting chemical for the target phase. Therefore, soil extractions are at best operationally defined, and hence are referred to by their method of extraction and not the target phase (Tessier and Cambell, 1991; Kim and Fergusson, 1991). Nevertheless, in light of the shortcomings of selective extractions, they provide valuable information regarding general partitioning patterns and qualitative estimates of reactive Mn phases within soils. Furthermore, they provide a means to assess the reactivity of the material that is not afforded by microscopic or spectroscopic methods.

Oxygenation of reduced soil upon collection leads to redistribution and repartitioning of contaminants (Rapin et al., 1986; La Force et al., 1999). To circumvent this undesired phenomenon, analyses of reduced phases were performed in a N2 purged glovebox. Sequential extractions were initiated using homogenized 4 g wet soil (~2 g dry weight equivalent) subsamples from the top 0 to 2 cm of cores taken from the three sampling sites. Following each extraction, the soil residue was washed with double-deionized water. Supernatants were filtered through a 0.2-µm membrane filter and acidified with concentrated trace element grade HCl prior to ICP analysis. Extractions were performed in triplicate along with appropriate controls and blanks. Operationally defined Mn sums varied slightly throughout the year (see below); thus, data were normalized to circumvent this phenomenon. Relative fractions were determined by summing all extractants to create an operationally defined total. Comparison of a total digestion via HNO3/HCl/HF with our operationally defined sums demonstrated excellent Mn recovery using the extraction sequence (relative standard deviation [RSD] generally <13%). The extraction procedure utilized in this study aspired to optimize the selectivity of each extraction while preserving sample integrity. Most importantly, however, we sought to determine the reactivity of Mn within different extractable phases using the defined extraction sequence.

The first step in the extraction series involves the removal of water-soluble and exchangeable Mn by the addition of 1 M MgCl2 at pH 7 followed by shaking for 1 h (Tessier et al., 1979). Second, Mn associated with the operationally defined carbonate phase was determined using a 1 M SA/AA at pH 5 for 5 h (Tessier et al., 1979). Following the SA/AA extraction, each soil sample was split with one half undergoing extraction to remove acid volatile sulfides (AVS) and noncrystalline (hydr) oxides and the other half removing solely noncrystalline (hydr)oxides. The AVS extraction involved the addition of 1 M HCl followed by shaking for 12 h (Huerta-Diaz and Morse, 1992). The noncrystalline (hydr)oxides were removed by adding 0.2 M ammonium oxalate at pH 3 and allowed to react in the dark (AOD) for 3 h (Schwertmann, 1973; Fey and LeRoux, 1977; Jackson et al., 1986). The difference between these two extractions (HCl-AOD) was used to approximate Mn association with amorphous sulfides. The sediment remaining after the HCl extraction was then treated with sodium hypochlorite (NaOCl) at 95 ± 5°C three times to enhance oxidation of organic phases and subsequent removal of Mn (Lavkulich and Weins, 1970; Hoffman and Fletcher, 1981; Shuman, 1983). Next, 1 M hydroxylamine-hydrochloride (HA)/AA was added and allowed to react for 6 h at 95 ± 5°C to dissolve crystalline (hydr)oxides (Tessier et al., 1979). Removal of residual materials and the silicate fraction was undertaken by reaction with 10 M HF for 16 h (Huerta-Diaz and Morse, 1990, 1992). Finally, the non-AVS-sulfide (pyritic) extractable fraction was removed by reacting the remaining solids with concentrated HNO3 for 2 h (Huerta-Diaz and Morse, 1990; Huerta-Diaz and Morse, 1992).


   RESULTS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Physical Site Conditions
The mineral content of the soil was 87% and the pH was 6.3. The particle-size distribution was fairly uniform across the site and was ~7% sand (0.05–2-mm), ~75% silt (0.05–0.002-mm), and ~18% clay (<0.002-mm). The average temperature at the sediment-water interface in the wetland increased from 7 to 20°C between early-spring (1 Mar. 1998) and summer (15 July 1998) (Interval A, Fig. 2) . Water levels were elevated at all sites in early-spring as a result of rainfall and snow melt; they then decreased as the wetland warmed. In September (15 Sept. 1998), the wetland dried and the highest sediment temperatures were recorded (25°C) (Interval B, Fig. 2). During fall, the wetland was replenished with water from rainfall and temperatures decreased. Finally, temperatures decreased to ~6°C in winter (Interval C, Fig. 2).



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  		Fig. 2. Seasonal sediment-water interface temperatures (Site 10) and water depths (all sites).

 
Aqueous Phase
A representative pH and EH depth profile from the middle of the wetland (Site 10) is illustrated in Fig. 3 . The pH of the site ranges from 7.3 during late spring to a low of 3.6 during late fall. The change in pH is consistent with cycles expected for reducing processes during the spring (increased pH because of bicarbonate liberation from microbial respiration) and oxidizing processes during the late summer into the fall (decreased pH resulting from Fe and S oxidation). The EH of the system ranged from -150 to 325 mV throughout the year. Redox potentials are semi-quantitative but for circumneutral pH values are considered oxic at EH > 400 mV, suboxic between 100 and 400 mV, and anoxic at EH < 100 mV (Sposito, 1981). During the spring and summer, the redox potential of the water column is indicative of a suboxic state while the decreasing values in the pore-water represent an anoxic status. Additionally, the redox boundary generally occurs within 10 cm of the sediment-water interface. In winter, biologic activity decreases and the EH remains poised above 250 mV.



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  		Fig. 3. Representative pH and EH depth profiles from Site 10. Samples were collected in (A) late winter (16 Mar. 1998), (B) late spring (13 Mar. 1998), (C) summer (8 July 1998), and (D) fall (23 Nov. 1998).

 
High variability exists in the aqueous phase Mn data temporally and spatially (Fig. 4) . Additionally, no statistically significant correlation between the sites could be determined from the given data set. However, throughout the year, soluble Mn(II) concentrations are often times lower in the oxic overlying water column because of (probable) Mn oxidation and formation of solid MnO2 phases; whereas, occasional peaks of aqueous Mn occur at the sediment-water interface and at specific points deeper in the sediment (Fig. 4). Increases in soluble Mn occurred at Site 5 below the sediment-water interface; whereas at Site 10, substantial levels of soluble Mn(II) (60 mg L-1) were detected within 6 cm (above and below) the sediment-water interface (Fig. 4A, B). On the basis of measured Mn(II) levels, we only noted oversaturation with respect to rhodochrosite in pore-water at Site 10 during summer sampling. If one considers the highest possible elevation of carbonate levels because of microbial respiration (Stumm and Morgan, 1995), pore waters may be slightly oversaturated with respect to rhodochrosite at Site 5 in the spring. However, it is clear that we did not detect a high degree of oversaturation with respect to rhodochrosite throughout the wetland during our seasonal sampling.



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  		Fig. 4A. Seasonal aqueous phase Mn concentration from Site 5 (the deepest part of the wetland).

 


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  		Fig. 4B. Seasonal aqueous phase Mn concentration from Site 10 (the middle of the wetland).

 


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  		Fig. 4C. Seasonal aqueous phase Mn concentration from Site 15 (near cattail vegetation). Samples were collected in late winter (16 Mar. 1998), late spring (13 May 1998), summer (8 Sept. 1998), and fall (23 Nov. 1998); the site had no standing water by September of 1998.

 
Solid Phase Manganese Speciation
The positioning of the main absorption edge in the XANES spectrum and corresponding first-derivative curves can be used to differentiate Mn oxidation states in soil (higher oxidation states occur at greater energies) (Manceau et al., 1992; Schulze et al., 1995; Friedl et al., 1997; Fendorf et al., 1999). For example, Mn(II) and Mn(IV) are separated by ~4 eV in their first-derivative XANES spectra (Fig. 5) . Additionally, the pre-edge intensity of MnCO3 is diminished relative to higher valent Mn (hydr)oxides (Manceau et al., 1992; Schulze et al., 1995). Pre-edge peak amplitudes and main absorption edges were similar in all Mn samples collected at the site and are consistent with the rhodochrosite spectrum; therefore, Mn(II) is the dominant solid phase oxidation state and appears to have a rhodochrosite-like structure. Linear combinations of the unknown Mn species were also best reproduced using the rhodochrosite spectrum (Fig. 6) . A final confirmation of the Mn structural state was achieved with EXAFS spectroscopy. The Fourier transformed and unfiltered k3-weighted {chi}[k] EXAFS functions are shown in Fig. 7 ; the unknown spectra are compared with predicted spectra for rhodochrosite calculated using the first two shells as described by Effenberger et al. (1981) and Friedl et al. (1997).



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  		Fig. 5. First-derivative XANES spectra of Mn standards used for linear combination fitting.

 


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  		Fig. 6. Unknown samples (solid lines) and optimal fits (dotted lines) based on linear combinations of the standards.

 


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  		Fig. 7. Fourier transform spectrum and corresponding k3 weighten {chi}(k) EXAFS function (insert) of a soil sample collected on 28 Mar. 1998 from Site 5. Location of Mn-O and Mn-Mn bond distances corresponds to a rhodocrosite-like structure.

 
Unfortunately, it is difficult to decipher whether the MnCO3 is a pure species or a mixed mineral phase. Regardless, the rhodochrosite-like species detected in the solids is the dominant Mn bearing phase and remains stable in the system independent of sampling location or time (Fig. 6). The importance of Mn carbonates in the Mn cycle has been previously noted (Pederson and Price, 1982; Thomson et al., 1986; Middelburg et al., 1987; Jakobsen and Postma, 1989; Friedl et al., 1997). However, in the present case we cannot determine whether the rhodochrosite phase is of primary, inherited with the mine waste deposition, or of authigenic origin.

Reactive Manganese Fractions
X-ray absorption fine structure spectroscopy is ideal for measuring the bulk properties of a sample; in contrast, SSE data provide useful measurements regarding the reactivity and mobility of Mn in the solid phase. It is important to note that the average operationally defined Mn totals throughout the year at the three sites ranged from 1400 to 1700 mg kg-1. The SSE data for all three sites throughout the period of investigation indicate variable trends in Mn reactivity (Fig. 8) . The AOD (23.5–175 mg kg-1), NaOCl (8.4–59.7 mg kg-1), and HNO3 (0.8–10.3 mg kg-1) extractable fractions contain appreciably less Mn than the MgCl2 (53.2–1150 mg kg-1), SA/AA (46.1–542 mg kg-1), HCl minus AOD (112–807 mg kg-1), HA/AA (40.9–1350 mg kg-1), and HF (77.3–430 mg kg-1) extractable fractions at all three sites. Thus, the following discussion will focus on the latter, more appreciable Mn fractions.



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  		Fig. 8. Reactive Mn solid phase fractions for all three sites throughout the period of investigation. Magnesium chloride (exchangeable and water soluble) (A), sodium acetate/acetic acid extractable (carbonate) (B), hydrochloric acid minus AOD extractable (amorphous sulfides) (C), hydroxylamine-hydrochloride/acetic acid extractable (crystalline oxide) (D), hydrofluoric acid extractable (silicate) extractable fractions (E) of the solid phase.

 
Manganese removed during the MgCl2 (exchangeable) extraction is water soluble or loosely bound to exchange sites and thus potentially more available for biological uptake. Manganese removed during this extraction varies from 2 to 43% of the total extractable Mn (Fig. 8A). Seasonal comparison indicates that Mn is significantly correlated (P < 0.05) among all three sites (Table 1). In late-summer, no standing water remains and Mn in the exchangeable and water-soluble phase increased significantly (P < 0.05)—approximately five to eight fold at the three sites comprising ~25 to 43% of the total extractable Mn (Fig. 8A). It is not surprising that as the site dries, Mn partitions onto exchange sites or within a soluble solid (e.g., an evaporite) and is dissolved using a weak extractant (i.e., MgCl2) (Bartlett and James, 1980). Consequently, as the wetland dries, Mn is appreciably retained within an easily extractable (i.e., highly available) pool.


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  		Table 1. Correlation coefficients for seasonal trend between sampling sites. Only the greatest fractions of extractable Mn are shown.

 
The SA/AA (carbonate) fraction varies among the three sites and comprises between 5 to 25% of the total Mn detected in the solid phase throughout the year (Fig. 8B). The seasonal trends for the sampling sites are not significantly correlated (P > 0.05) (Table 1). The SA/AA (carbonate) extractable fraction is elevated in the spring at Site 10 and Site 15 and in the late-summer at Site 5. It appears that increases in this extractable fraction occur as water levels decrease; however, they are highly variable within the wetland. Manganese is known to partition with carbonates in anoxic sediments (Pedersen and Price, 1982; Thomson et al., 1986; Jakobsen and Postma, 1989), consistent with the x-ray absorption spectroscopy (XAS) analysis. There are a number of reasons why the results obtained by XAS are not consistent with those from SSE analysis; in the following section we provide details on the origin of these differences.

The HCl extraction subtracted from the AOD extraction estimates the fraction of amorphous sulfides within the soil. In anoxic portions of wetlands, amorphous sulfides may play an integral role in Mn and trace element sequestration (Morse et al., 1987; Huerta-Diaz and Morse, 1992; Arakaki and Morse, 1993). Manganese removed using HCl minus the AOD extraction varies from 10 to 33% throughout the year (Fig. 8C). The seasonal trends for this chemical extraction indicate that Site 5 to Site 10 and Site 10 to Site 15 are significantly correlated (P < 0.05) (Table 1). Regardless of the sampling site, concentrations in this extractable pool increase in the summer as the wetland warms and biological activity increases; whereas, in the late-summer the site dries and Mn in this extractable fraction decreases. Coincidentally, elevated Mn concentrations in the AVS pool occur at the same time we detect a pungent (H2S[g]) odor emanating from the site.

The HA/AA (crystalline oxide) extractable fraction varies from 4 to 49% of the total extractable Mn throughout the year (Fig. 8D). The seasonal trends for this chemical extraction are significantly correlated (P < 0.05) among the three sampling sites (Table 1), indicating that this extractable pool responds uniformly at the site over the period of our investigation. Moreover, Mn increases significantly (P < 0.05) at all three sites in the summer when water levels decrease and remains elevated through early-winter.

The HF(residual materials and silicates)-extractable Mn fractions comprise between 3 and 38% of the total extractable Mn within the wetland (Fig. 8E). The seasonal trends for the HF extraction are significantly correlated (P < 0.05) among sampling sites (Table 1). Additionally, the HF extractable Mn fractions are significantly (P < 0.05) higher in the spring at all the sites within the wetland (Fig. 8E). Elevated Mn associated with the HF-extractable pool in the spring was unexpected; interestingly, Mn concentrations are inversely related to the HA/AA (crystalline oxide) extractable fraction (Fig. 8D, E).


   DISCUSSION
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 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
CONCLUSIONS
 REFERENCES
 
Differences between SSE and X-ray Absorption Spectroscopy Analyses: Reactivity versus Quantity
Appreciable differences between the XAS and SEE analysis are apparent and these discrepancies at first may seem distressing. However, if one considers what is being measured by each method, the results should be independent and complementary rather than redundant (i.e., they are not measuring the same thing). X-ray absorption spectroscopy characterizes the chemical and structural state of an element throughout a sample; surface sensitivity can be gained with special configurations but only for flat samples (generally single crystals) or ones exposed to a He(g) (or vacuum) environment. Chemical extractants, in contrast, measure the reactivity of a sample; they thus attack the surface of the solid phases. While chemical extractants provide a means to measure the reactivity of the solids, one should note that there specificity for a target phase is limited (Tessier et al., 1979; Nirel and Morel, 1990; Kim and Fergusson, 1991).

We should next ask whether XAS data will allow you to predict the reactivity of the phase. The answer is "not always". If the bulk of a material is coated with a fine veneer of organic matter or a secondary precipitate (e.g., metal hydroxide or metal sulfide), then the reactivity will be appreciably different (the M&M scenario) than the bulk measurement would suggest. Thus, we believe that a powerful mixture of information is obtained by using XAS in combination with SSE; XAS provides a bulk measurement while SSE give the reactivity of the material. The two methods are used here as measures of independent parameters, one to obtain the dominant form of Mn and the other to assess its reactivity. In combination, they reveal that the bulk of Mn (>95%) remains locked in a rhodochrosite-phase, which we believe is largely of primary origin (deposited with the mine tailings) since it does not undergo seasonal alteration. In contrast, the reactivity of Mn, as measured by SSE, is altered appreciably with seasonal changes. We thus postulate that surface transformations of, and coatings on, the rhodochrosite particulates accounts for the observed data.

Cycling of Manganese
The general conception of Mn is that in aerobic waters, such as the surficial waters of a wetland, oxidized forms of Mn (IV, III) commonly persist; in anaerobic portions of the soil or sediment, Mn(II) predominates as a result of reducing conditions—microbially and chemical induced (Postma, 1985; Myers and Nealson, 1988; Canfield et al., 1993; Burdige, 1993). Homogeneous oxidation of Mn(II) by O2(g) proceeds at a limited rate (Diem and Stumm, 1992); however, the process can be microbially mediated and catalyzed by mineral surfaces (Hem, 1981; Murray et al., 1985; Davis and Morgan, 1989; Junta and Hochella, 1994). Thus, soluble Mn(II) may mobilize toward aerobic portions of the soil or water and undergo oxidative precipitation; however, it is also possible that Mn(II) may precipitate to form carbonates or sulfides (Hem, 1978; McBride, 1979; Balzer, 1982; Franklin and Morse, 1983; Middleburg et al., 1987; Canfield, 1993; Wang and Van Cappelan, 1996; Lebron and Suarez, 1999).

The Mn cycle within the Cataldo Wetland appears largely consistent with the general paradigm described above: It is driven by a combination of microbial mediated and surface catalyzed processes. The SSE data captures these transformations. In spring, the wetland floods, temperatures gradually increase, yielding a more biologically and chemically active environment than during the winter—one that leads to decreased redox potentials. The majority of Mn is removed using HF (silicate), HCl-AOD (amorphous sulfides), and SA/AA (carbonate) extractants, all of which are considered Mn(II) bearing phases. With the progression of summer, water levels begin to decrease as temperatures increase. These two factors have somewhat offsetting impacts on the biogeochemistry of the wetland; lower water levels promote aeration while increased temperature stimulates microbial activity and thus anoxia. However, on the basis of extractable Mn, it appears that water levels are the overriding factor during the summer months. Manganese in the HF(silicate)-extractable pool decreased and Mn is partitioned into the HA/AA (crystalline oxide)—HCl-AOD (amorphous sulfides) and SA/AA (carbonate) extractable phases remain appreciable. In the late summer, the wetland dries leading to desiccation induced partitioning of Mn into the MgCl2-extractable fraction. Finally, in winter (25 Jan. 1999) temperatures decline, water levels increase, and the site becomes biotically and chemically less active as denoted by elevated redox potentials.

Neglecting the exchangeable fraction owing to desiccation induced inputs from the water column, the composite of the remaining Mn(II) bearing phases (carbonate + sulfide + silicate) and Mn(III/IV) phases (oxide + amorphous + crystalline) gives rise to trends consistent with those expected for Mn cycling (Fig. 9) . Manganese (II) bearing pools increase slightly during fall as the site wets and then dramatically increase during the spring as it warms—a result of stimulated anaerobic metabolism. As the water begins to subside and the wetlands become more oxygenated in summer, the oxide fraction is enhanced. Increased water levels during the fall result in only modest changes but do generally lead to slight elevation of the reduced pools. It is apparent that the Mn cycle is dictated dominantly by water level and secondarily by temperature. Temperatures not restricting, Mn pools respond directly to water levels, with increased flooding leading to reduction.



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  		Fig. 9. Seasonal distribution of Mn in reduced and oxidized-extractable pools.

 
An interesting observation within the Cataldo wetland is the apparent presence of a Mn-silicate pool controlling a large fraction of the Mn-reactivity, as noted by the HF extraction. There are a few plausible explanations that can account for the increased Mn removed via HF. First, an unknown recalcitrant solid phase was not removed with the previous reagents which may in part be responsible for an increase in Mn removed via HF. However, this seems unlikely given the seasonal changes in the HF-extractable pool. Alternatively, a Mn(II)-silicate phase formed. While such solids are uncommon, soils within the CdAR region are heavily impacted by volcanic glass, material which leads to appreciable concentrations of dissolved silica. The final question returns as to why we do not detect this phase in the x-ray absorption spectrum? Because the extractions measure surficial changes in solids phase but XAS does not (given the experimental configuration used here), we postulate that surface transformation on rhodochrosite grains, in addition to discrete phases, explain the observed trends. It is likely that silica may adsorb on the mineral surfaces and thus gives the appearance, on the basis of extractions, that a discrete silicate phase formed.

Thus, the hypothesized reaction scenario is that during the summer months, aeration promotes the surficial oxidation of rhodochrosite grains. Additionally, desiccation of the wetland leads to the deposition of evaporites. Upon inundation of the wetland with water during the winter, evaporites undergo dissolution and modest reduction of Mn(IV/III) transpires. As the site warms during the spring, biological activity is promoted and reductive dissolution leads to the formation of Mn(II) bearing phases—or surface coatings. Dissolved silica resulting from volcanic glass within the wetlands promotes the formation of Mn(II) silicate phases; nucleation is expected to occur at the site of reduction and thus leads to a surface coating of the silicate on the rhodochrosite substrate. Progression back into summer leads to the oxidative dissolution of the Mn(II) solids.


   CONCLUSIONS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
REFERENCES
 
Using XAS, we are able to determine the bulk characteristics of Mn within the solids; the reactivity of these phases were then assessed using chemical extractants. In combination, we determined that the dominant solids and their reactivities changed seasonally within this mining impacted wetland. Rhodochrosite appears to be the principal form of Mn within the wetland throughout the year. Although pore-water near the sediment-water interface is oversaturated with respect to this mineral phase only seasonally, it remains as the major phase throughout the year. However, the reactivity of Mn does change through the course of our investigation. Manganese is primarily extracted within MgCl2 (exchangeable and water soluble), HCl-AOD (amorphous sulfide), HA/AA (oxide), and HF (silicate) fractions. Because the bulk form of Mn does not change during the year but the reactivity does, we suspect that the Mn carbonates develop an oxide and silicate rind seasonally. Additionally, an evaporite results during complete desiccation of the wetland in late summer. Rewetting of the site in fall results in a pulse of dissolved Mn and thus represents the greatest point of concern for elevated levels of metals similar to Mn(II) within the water column or pore-waters. Changes in the surface chemistry of the Mn phases may also result in alterations in contaminant retention properties throughout the year.


   ACKNOWLEDGMENTS
 
We gratefully acknowledge support provided by the Exploratory Research Program of the US EPA (grant number R-825399). We also thank Benjamin Bostick for assistance with XAFS data collection and Dr. G.C. Li for assistance in statistical analysis. Special thanks are given to the staff at the Stanford Synchrotron Radiation Laboratory (SSRL). The SSRL is operated by the Department of Energy, Office of Basic Energy Sciences; support for the Biotech program is provided by the National Institute of Health.

Received for publication April 3, 2000.


   REFERENCES
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 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
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 CONCLUSIONS
 REFERENCES
 





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