Redox geochemistry in organic-rich sediments of a constructed wetland treating colliery spoil leachate

Applied Geochemistry 24 (2009) 44–51

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Redox geochemistry in organic-rich sediments of a constructed wetland treating colliery spoil leachate Michelle I. Morrison, Andrew C. Aplin * School of Civil Engineering and Geosciences, Drummond Building, Newcastle University, Newcastle upon Tyne, NE1 7RU, UK

a r t i c l e

i n f o

Article history: Received 7 July 2008 Accepted 1 November 2008 Available online 11 November 2008 Editorial handling by Dr. R. Fuge

a b s t r a c t The results are reported of a geochemical study of sediment cores and surface waters taken over an annual cycle from the compost-based constructed wetland at Quaking Houses, NE England. The wetland was built to treat acidic and metalliferous waters emanating from colliery spoil. The influent waters contain up to 10 mM SO2 4 , total Fe around 100 lM, and a mean pH of 6.2. The organic-rich sediments sustain a coupled redox cycle of Fe and S which occurs throughout the year but which is more intense in the summer months. Throughout the sediments, reduction of Fe(III) and SO2 4 apparently occur within the same macroscopic volume of sediment, along with oxidation of sulfide and Fe(III). In the top 2 cm of the sediment, pore water Fe concentrations reach a maximum of 1 mM in the presence of high concentrations of Fe oxides and the occurrence of SO4 reduction. Partial re-oxidation of sulfide is indicated by the presence of significant elemental S. Total Fe in the surface 10 cm has a mean value of 7–9% of sediment dry weight, of which sulfide–Fe represents 7–12% and reactive amorphous Fe 20–45%; around 90% of the solid phase sedimentary Fe occurs as oxides or oxyhydroxides. It is suggested that the downwards diffusion of dissolved Fe from the near-surface maximum is sustained by the precipitation of Fe oxides as a result of radial O2 loss from roots in the dense rhizosphere. Pore water pH is between 7.2 and 7.8 and alkalinity increases downwards, coupled to microbial SO4 and Fe reduction. Transport processes occurring at and across the sediment–water interface are sufficiently rapid in the 20 h residence time of the waters to: (a) remove 70–90% of influent Fe and 15–25% of influent SO4 into surface sediments and (b) increase both the pH and alkalinity of effluent waters. Coupling of the Fe and S cycles is fundamental to effective remediation in terms of both alkalinity generation and the retention of metals. Ó 2008 Elsevier Ltd. All rights reserved.

1. Introduction Acidic, metalliferous pollution arising from abandoned mine workings and spoil heaps is a major environmental threat to substantial lengths of river courses worldwide (Younger, 1997). Depending on the chemical nature of the waters and their contaminant loadings, treatment solutions can include solely active or solely passive systems, or a combination of both (Younger et al., 2002; Watzlaf et al., 2004 and references therein). Active treatment systems are employed where flows and contaminant loads are high, using chemicals such as flocculants and pH adjusters to ameliorate the contamination. They require constant inputs of energy, maintenance and manpower and are consequently expensive to operate. Passive treatment systems do not require chemicals or energy inputs, but instead direct the flow of contaminated water over or through constructed, compost-based systems. In using composted materials as a substrate for the treatment of mine-impacted waters, the aim is to use the natural biogeochemical C–S–Fe cycle to: (a) increase pH and alkalinity and (b) remove metal contami* Corresponding author. Tel.: +44 191 222 6513; fax: +44 191 246 4961. E-mail address: [email protected] (A.C. Aplin). 0883-2927/$ - see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.apgeochem.2008.11.001

nants such as Fe and Al. For the treatment of net acidic waters, anaerobic systems including wetlands, barriers and RAPS (reducing and alkalinity producing system) are employed since they have the potential to increase alkalinity and pH through the microbially mediated oxidation of organic C; aerobic systems (reed beds) are of limited use in net acidic waters because although Fe can be successfully removed by oxidation and hydrolysis – and trace metals subsequently removed from solution by scavenging onto the surface of the oxyhydroxides (Johnson and Thornton, 1987) – the oxidation and hydrolysis of both Fe and Al generate significant acidity. Ultimately, successful remediation hinges on the appropriate rate and balance of the C–S–Fe redox cycles such that the rate of proton generating reactions such as Fe and sulfide oxidation are outweighed by proton consuming/alkalinity generating reactions such as microbial SO4 and Fe reduction. Whilst these concepts and reactions have been very extensively explored in marine and freshwater sediments (e.g. Berner, 1985; Holmer and Storkholm, 2001; Kostka and Luther, 1994; Lin and Morse, 1991; Thamdrup et al., 2000; Van Cappellen and Wang, 1996), very few comparable studies have been undertaken within constructed wetlands (Fortin et al., 2000; Webb et al., 1998; Woulds and Ngwenya, 2004). This

M.I. Morrison, A.C. Aplin / Applied Geochemistry 24 (2009) 44–51

paper outlines the results of a geochemical study of the Quaking Houses constructed wetland in County Durham, England, UK. 2. Methods 2.1. Site location and design The Quaking Houses artificial wetland is situated in County Durham, NE England, UK, at grid reference 417780, 550961, or tile NZ 179 508. The wetland is a surface flow system consisting of two interconnected ponds plus a naturally formed aerobic reed bed (Fig. 1). A surface flow wetland design was chosen in view of the limited hydraulic head available at the site (up to 1 m), plus the relatively low flow rates associated with the discharge (mean rate of 98 L min1). The wetland system was built in 1997 to treat a net acidic colliery spoil leachate emanating from a perched aquifer within a large spoil heap. The main chemical characteristics of the leachate during the initial period of treatment were: pH 4.5, zero bicarbonate alkalinity, 0.13 mM Fe, 0.49 mM Al, 28 mM Na, and 21 mM Cl (Younger et al., 1997; Jarvis, 13.6 mM SO2 4

45

2000). At the time of this study, influent water quality had improved significantly as a result of the application of a 0.5–1 m clay cap to the spoil heap (Gandy and Younger, 2003). The wetland is lined with a few centimeters of compacted pulverized fuel ash (PFA) to form an impermeable layer beneath the substrate. The substrate of the two ponds (referred to hereafter as ‘sediment’) consists of a mixture of horse manure, cow manure and municipal waste compost, in an approximate 30:40:30 ratio. The two ponds have a collective surface area of 440 m2 and were filled with 30 cm of compost. Baffles plus two islands in the second pond (away from the influent) were constructed using PFA as the base layer and were designed to minimize the potential for hydraulic short-circuiting through the second pond. Two species of wetland plant were planted around the edges of the ponds, Phragmites australis and Typha latifolia, chosen in part for their tolerance to metals. Other plants such as the Flag Iris also colonized the wetland. The effluent from pond 2 flows into a small, naturally formed reed bed, before rejoining the Stanley Burn. Over the period of this study (16 months), there was a marked increase in vegetation within both ponds, resulting in almost complete plant cover.

Fig. 1. Schematic diagram of Quaking Houses wetland (not to scale). The water enters the wetland at the influent distribution point, flows across ponds 1 and 2, then exits into a naturally formed reed bed before re-entering the Stanley Burn. The shaded rectangle indicates the coring area, which was densely vegetated. The four circles, as indicated by the legend, represent surface water sampling points. The ponds are of approximately equal size, with surface areas of around 200–250 m2.

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2.2. Sampling and geochemical analyses Cores and surface waters were taken on four occasions between May 2001 and September 2002, representing an annual cycle of changing temperature and plant growth; additional samples of influent and effluent waters were taken over a 3-a period between May 2001 and June 2004. During each of the four main sampling trips, a sediment core was taken (Fig. 1), as well as water samples and field measurements. Sediment cores were taken from a similar central location of pond 2 using a 600 mm  50mm section of HDPE plastic piping. The core tube was securely sealed at both ends to prevent loss of material and ingress of O2. Surface water samples were taken from four sample points representing the influent, pond 1, pond 2 and the effluent from pond 2 (Fig. 1). Cation samples were preserved using 1% v/v of 36% HCl (AnalaR grade) and all water samples were manually filtered onsite through 0.45 lm membrane filters (Whatman, cellulose nitrate, 47 mm diameter) using a 60 mL plastic syringe and a 47 mm Swinnex filter holder. The height of the water column above the sediment was between 5 cm and 20 cm, depending on the season (lower during the summer, higher during the winter). Surface water samples were taken midway between the air–water and sediment–water interfaces. On-site field measurements of pH, Eh, conductivity, alkalinity and temperature were taken using a HACHÒ alkalinity test kit (model AL-DT, 10–4000 mg/L), and a MyronÒ L Company Ultrameter. All sediment and water samples collected were stored at 4 °C. Water samples were analysed within a week of collection. Concentrations of major anions and cations in surface waters were determined by Ion Chromatography and Flame Atomic Absorption Spectroscopy, respectively. Aqueous Fe(II) and Fe(III) were measured using a modified version of Stookey’s colorimetric method (Viollier et al., 2000). The pH and bicarbonate alkalinity were determined in the surface water samples using titrimetric methods. Sediment cores were extruded in a glove box which had been flushed with N2 for 40 min to remove O2. Samples were taken at 1 cm depth intervals, placed in clean 25 mL sample tubes and stored at 4 °C prior to processing. Following centrifugation of the sediment samples (2000 rpm for 10 min), pore waters were extracted from the sediments in a glove box under anoxic conditions. Pore waters were placed into 2 mL microcentrifuge tubes (pore waters for Fe(II)/Fe(III) determination were preserved with 1% v/v of 36% HCl) and were subsequently analysed for SO2 4 , Fe(II)/Fe(III), and pH using the same methods as for the surface waters. HCO 3 The speciation of solid phase S was carried out according to the Duan et al. (1997) method. Sediments were sequentially extracted to determine operationally defined pools of acid volatile sulfur (AVS), chromium reducible sulfur (CRS) and elemental sulfur (ES). In this study, AVS is taken to represent solid phase FeS plus aqueous phase H2S, while CRS represents FeS2. Approximately 1 g of wet sediment was placed into a pre-weighed glass centrifuge tube (Quickfit), in a N2 filled glovebox, and then re-weighed. The sample was treated with 5 mL of 20% w/v Zn acetate (ZnAc) solution to fix any free sulfide. This was then placed in a Denley Ultrafreezer at 76 °C for 2 h. The frozen sample was freeze–dried in an Edwards Modulyo freeze–drier for at least 24 h. Once dry, approximately 25 mL of DCM was used to extract the S0 fraction from the sample over a 24 h period, on a shaker. The sample was then centrifuged at 3000 rpm for 10 min, and the supernatant isolated. This was repeated with two further washings of DCM (5 mL each). The isolated DCM fractions were pooled into a 100 mL round bottomed flask (RBF), and the remaining sediment transferred into a different 100 mL RBF, with a little ethanol. The sediment fraction underwent a cold 6 M HCl extraction for 1 h, to isolate the AVS fraction, followed by a hot 1 M CrCl2 (in 1 M HCl) extraction for 1 h to isolate

the CRS fraction. The DCM supernatant containing the ES fraction was reduced to near dryness using a rotary evaporator (Büchi Rotorvapor). Any remaining solid was extracted with hot CrCl2 plus an additional 2 mL of 6 M HCl for 1 h to isolate the ES fraction. The liberated S from each extraction step was collected in a double trap setup containing 5% ZnAc solution and precipitated as ZnS. Concentrations of (Zn) sulfide in the 3 S pools were determined using an iodometric back-titration method. In addition to sulfidic Fe phases, sedimentary oxide phases were extracted using either the ascorbate (Kostka and Luther, 1994) or dithionite (sodium dithionite/sodium citrate buffered to pH 4.8 with 50:50 glacial acetic acid; Raiswell et al., 1994; Poulton and Canfield, 2005) extraction methods. Ascorbate is considered to dissolve poorly crystalline Fe oxyhydroxide phases whereas dithionite dissolves both poorly crystalline phases and crystalline phases such as lepidocrocite, akaganeite and goethite (Kostka and Luther, 1994; Poulton and Canfield, 2005). Total Fe was determined following complete dissolution using HF/HNO3/HClO4. The 3 extraction procedures were not carried out sequentially, but on discrete samples. The liquor remaining following sulfide extraction of the September core samples was analysed for Fe by FAAS, to compare with total sediment Fe data. Total organic C, total inorganic C and total S analyses were determined on selected samples using a Leco-244 C–S Analyser. XRD analysis on a few samples was performed using a PANalytical X’Pert Pro Diffractometer. Approximately 1 g (accurately weighed) of wet sediment was dried for 24 h in an oven at 105 °C for wet/dry weight conversions. 3. Results 3.1. Surface water chemistry Compositional data for influent and effluent waters are shown in Table 1. The general characteristics of the influent waters are that they are brackish (high concentrations of Na+, Cl and SO2 4 ) and have a range of pH values between 5.6 and 6.7 (to give a mean  value of 6.2, Table 1). No PO3 4 or NO3 is detected. The origin of the +  high concentrations of Na and Cl are uncertain but may relate to the storage of road salt in the immediate catchment. Alkalinities of the influent range from 18 to 81 mg L1 as CaCO3 and, like pH, are highest during November and January. The total Fe content of the influent waters vary between 12 and 157 lM, with a wide range of Fe(II)/Fe(III) ratios (Fig. 2). In November 2001 and January 2002,

Table 1 Summary of influent and effluent surface water composition over the period May 2001–April 2004. Determinant

Mean influent value

Range (n)

Mean effluent value

Range (n)

Temperature (°C) pH Eh (mV) Conductivity (mS cm1) Alkalinity (mg L1 as CaCO3) Na (mM) K (mM) Ca (mM) Mg (mM) Fe (lM) total Fe2+ (lM) Fe3+ (lM) SO2 4 (mM) Cl (mM)

11.4 6.2 112.5 4.5

5.9–15.8 (13) 5.6–6.7 (13) 26–277 (13) 2.5–11.9 (12)

11.1 6.4 98 3.5

3.8–14.7 (13) 6.1–6.7 (13) 56–293 (13) 2.6–6.4 (12)

52

18–81 (13)

78

35–117 (13)

29.0 2.2 4.9 2.8 89.0 27.1 52.5 6.6 30.5

23.9–35.4 (4) 1.6–4.1 (4) 3.6–6.2 (4) 2.0–3.8 (4) 11.9–157.3 (7) 1.4–50.7 (6) 0–109.6 (6) 3.7–10.1 (6) 7.4–49.4 (6)

28.3 2.1 4.2 2.3 23.9 5.9 15.3 5.1 28.0

25.2–32.6 (4) 1.5–3.8 (4) 3.0–5.5 (4) 1.6–3.3 (4) 3.8–74.1 (6) 0–26.4 (6) 1.6–47.7 (6) 2.8–8.5 (6) 10.2–41.4 (6)

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M.I. Morrison, A.C. Aplin / Applied Geochemistry 24 (2009) 44–51

SO42- (mM) 0

Sept02

2

2

4 6

4

10 12 14

May01

14

16 18

Nov01

16

20

Jan02 Sept02

18

0 pond 1

pond 2

350

300

250

200

150

100

50

0

0

Jan02

0

0

Sept02

2

2 4

10 12 14

0

20

18 20

22

22

180 7

Nov01

140

2

Sept02

4 Depth (cm)

120

80 60

1400

0

Jan02 100

7.2

pH

May01

160

7.6

effluent

10 12

16

7.4

pond 2

8

14

16 18

pond 1

6

8

20

influent

4

10 12

7.8

Fe 40

2

6

8

Depth (cm)

Depth (cm)

(µM)

4 6

60

1200

HCO3- (mM)

Fe(III) (µM)

80

1000

22

Nov01 100

800

20

22

effluent

120

Total Fe (µM)

8 10 12

10

influent

600

6

8

8

30 20

3+

0

0

Depth (cm)

40

0 1 2 3 4 5 6 7 8 9

Jan02

Depth (cm)

Fe

2+

(µM)

50

Fe(II) (µM) 400

Nov01

200

60

6 8 10 12 14

40

16

20

18

0 influent

pond 1

pond 2

effluent

Fig. 2. Surface water concentrations of Fe(II), Fe(III) and total Fe over an annual cycle. Note that samples were filtered through a 0.45 lm paper so that some of the Fe may be colloidal.

Fe(II) comprises 50% and 30%, respectively, of the total Fe. Iron and concentrations are highest in surface water samples from SO2 4 May and September. Changes in water chemistry across the wetland indicate 15–25% removal of SO2 4 , 70–90% removal of Fe and a gain in alkalinity. 3.2. Pore water chemistry Concentrations of pore water SO2 4 in May and September decline steadily from 7–8 mM to <1 mM at 12 cm depth (Fig. 3). In contrast, SO2 4 profiles for November and January show a subsurface maximum at 1–2 cm depth, followed by a decline to <1 mM at 12 cm depth.

20 22 Fig. 3. Pore water data for each core. Depth 0 cm represents the sediment–water interface and the data point at depth 0 cm represents the surface water concentration for that chemical species. Standard deviations for n = 2 are plotted, but in most cases are smaller than the data points themselves (<5% RSD). Pore water pH and alkalinity data were not determined for the May 2001 core. Note that samples were filtered through a 0.45 lm paper so that some of the Fe may be colloidal.

Bicarbonate alkalinity profiles (Fig. 3) are similar throughout the year, increasing rapidly over the upper 4 cm of the profile from approximately 2 mM to 8 mM, then more slowly over the remain2 ing 18 cm from 8 mM to 11 mM. Based on the 2HCO 3 :1SO4 reaction stoichiometry for bacterial SO4 reduction (BSR), the alkalinity increase is less than that predicted from the observed reduction of 2 pore water profile suggests the generation of SO2 4 . The SO4 approximately 8 mM HCO 3 alkalinity over the top 4 cm plus a further 7–8 mM over the rest of the profile depth. Pore water pH ranges from 7.1 to 7.9 with no clear spatial or temporal trends (Fig. 3).

M.I. Morrison, A.C. Aplin / Applied Geochemistry 24 (2009) 44–51

0

2

2

1600

1400

1200

1000

800

600

0

0

400

Fe (µmol/g) Nov 01 1600

1400

1200

1000

800

600

400

200

0

Fe (µmol/g) May 01

200

48

Depth (cm)

Depth (cm)

4 6 8 10 12

4 6 8 10

14 12 16

dithionite-Fe

18

14

sulphide-Fe 16

20

1800

1600

1400

1200

1000

800

600

200

0

2500

2000

1500

1000

500

0

400

Fe (µmol/g) Sept 02

Fe (µmol/g) Jan 02 0

0

2 4 Depth (cm)

Depth (cm)

5 10 15

6 8 10 12 14

20

16 25

asc. Fe sulphide-Fe total Fe

30

18 20

asc.Fe sulphide-Fe total Fe HCl total Fe

22

Fig. 4. Solid phase Fe data from ascorbate (asc-Fe), dithionite, sulfide and total Fe extractions. There is no total Fe data for the May core. The September graph also shows the data for Fe remaining in the liquor following sulfide extraction (in effect, extracted by 6 M HCl) as ‘HCl total Fe’.

Concentrations of aqueous Fe(II) in the pore waters increase rapidly through the top 2 cm of sediment, with concentrations reaching 1100–1300 lM in May and September, and peaking at approximately 850 lM during November and January (Fig. 3). In addition to the HCO 3 generated as a result of BSR, Fe reduction  would generate up to 0.3 mM HCO 3 , based on a 1HCO3 :4FeOOH reaction stoichiometry. Concentrations of aqueous Fe(II) decline erratically below 2 cm, reaching values of 200 lM at the base of the cores. Surprisingly, based on the pH, Fig. 3 also shows that most of the pore waters contain 20–50 lM of aqueous Fe(III), with peaks of up to 300 lM. Pore water sulfide values were below detection limit, consistent with the high concentrations of dissolved Fe. 3.3. Sediment chemistry The Quaking Houses sediments typically contain 12–15% organic C, 0.1% inorganic C and 0.5–1.5% S. Solid phase Fe and S data are shown in Figs. 4 and 5, respectively. Total Fe in the top 6 cm of the sediments is typically 1200–1600 lmol g1, with a value of 2300 lmolg1 at the surface of the January 2002 core. These values correspond to 7–13% of the sediment dry weight and are substantially greater than the mean value quoted for sediments in salt marshes (Kostka and Luther, 1994) and fresh water lakes (0.1– 2.3%; Luther et al., 2003). In deeper sediment, solid phase Fe con-

centrations decline steadily to values around 300–600 lmol g1 (Fig. 4). Selective leaching of the sediments shows that sulfidic Fe comprises only 7–12% of the total Fe, except in the some of the deepest sediments where total Fe is less (Fig. 4). The amount of sulfidic Fe as a proportion of both total Fe and ascorbate leachable Fe is greatest in winter and lowest in summer. The authors believe for several reasons that the majority of the non-sulfidic Fe comprises Fe oxides. Firstly, ascorbate leachable Fe, which is thought to comprise amorphous Fe oxyhydroxides (Kostka and Luther, 1994; Poulton and Canfield, 2005), represents between 20% and 45% of total Fe. Secondly, the Fe concentrations extracted by citrate–dithionite (which extracts most Fe oxide and oxyhydroxide phases; Kostka and Luther, 1994; Poulton and Canfield, 2005) from the May 2001 core are much higher than those extracted by ascorbate and are similar to the total Fe concentrations in the November and September cores. Thirdly, it appears that 6 M HCl, which will dissolve oxides and carbonates but not organic matter, extracts essentially all the Fe from the September core: analysis by FAAS of the liquor remaining after sulfide extraction of the September core samples gave Fe concentrations very similar to those achieved through the ‘total Fe’ extraction procedure. Since inorganic C data suggests that there can be little FeCO3 present, it is inferred that most of the Fe in the Quaking Houses sediments occurs as oxides, except in some of the deepest sections (Fig. 5). Nevertheless, it was

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AVS CRS

500

400

300

AVS CRS ES

400

350

300

250

200

150

100

0 Depth (cm)

50

S (µmol/g) Sept02 250

200

150

100

50

0 Depth (cm)

200

0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30

S(µmol/g) Jan 02 0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30

100

0

250

200

150

100

50

0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30

S (µmol/g) Nov 01

Depth (cm)

Depth (cm)

0

S (umol/g) May 01

0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30

Fig. 5. Solid phase S data from extractions with 6 M HCl and CrCl2 are shown, where AVS is acid volatile S (H2S + FeS), CRS is chromium reducible S (FeS2) and ES is elemental S (S0). The CRS fraction for May represents FeS2 plus S0.

recently deposited surface layers; transport of Fe and S into the sediment has thus occurred. The C–S–Fe cycle within the Quaking Houses wetland commences with the delivery of SO2 4 and Fe(III) to the sediment–water interface. During the summer months, pore water data show that SO2 4 is lost and bicarbonate alkalinity is generated, consistent with reduction; the occurrence of AVS and pyrite also bacterial SO2 4 reduction. Since the increase in HCO indicates SO2 4 3 is less than that calculated from the 2:1 stoichiometry of bacterial SO2 4 reduction, it is also possible that some SO2 4 is removed by the precipitation of Fe oxyhydroxysulfates such as schwertmannite, although this was not detected using XRD. Given the organic-rich nature of the Quaking Houses sediment reduction in the sediments, and the occurrence of bacterial SO2 4 it is perhaps surprising that: (a) most of the sedimentary Fe occurs as oxides and (b) there are substantial concentrations of dissolved Fe in the pore waters. During May and September, in the top 2– 4 cm of this organic-rich sediment, pore water data suggest that reduction occur in the same macroboth Fe reduction and SO2 4 scopic sediment volume. The rate of Fe reduction exceeds the net rate of sulfide generation and builds up pore water Fe concentrations greater than 1 mM. Although the data cannot be used to distinguish between microbial and chemical Fe reduction, it is likely that both are occurring; the presence of elemental S in sediments containing both substantial Fe oxide and dissolved Fe suggests that oxidation of reduced S by Fe oxides is a likely reaction:

2FeOOH þ H2 S þ 4Hþ ! 2Fe2þ þ S0 þ 4H2 O

ð1Þ

0

ð2Þ

3H2 S þ 2FeOOH ! S þ 2FeS þ 4H2 O

The reaction of dissolved, reduced S with Fe oxides is important from a treatment perspective because it inhibits complete re-oxidation to SO2 4 by O2 and thus the regeneration of acidity: þ H2 S þ 2O2 ! SO2 4 þ 2H

ð3Þ

In the winter months, profiles of dissolved Fe and SO2 4 2 suggest that both SO4 and Fe reduction continue (Fig. 3). However, the pore water SO2 4 profile displays a maximum concentration at 1–2 cm into the sediment, indicating that SO2 4 is lost from the sediments by diffusion to the overlying water column. The observed seasonal change in the pore water SO2 4 profile can in principle relate to two processes. The lower SO2 4 concentrations observed in the winter surface waters will force the diffusive re-equilibration of the pore water profile (Fig. 3). In addition, lower ratios of sulfidic Fe to both total Fe and ascorbate Fe in winter compared to summer argues for the partial re-oxidation of solid phase S (Fig. 4), as previously observed in some salt marsh and wetland sediments (e.g. Kostka and Luther, 1995; Koretsky et al., 2003; Neubauer et al., 2005; Koretsky and Miller, 2008). Over an annual cycle, the solid phase data show that significantly more Fe oxide than reduced S accumulates in the sediment. In summer, approximate calculations of the rates of addition of Fe and S to the wetland (Fe addition based on influent and effluent data; S addition from the SO2 4 diffusion profile using Fick’s law) suggest that similar amounts of S and Fe are added to the wetland. A similar calculation for the winter months shows continued addition of Fe, but substantial loss of S; the balance of the winter and summer cycles result in the observed preponderance of oxidized Fe over reduced S in the wetland solid phases. The pore water Fe data shows the characteristic maximum reported previously in both natural marine (Canfield et al., 1993; Thamdrup et al., 2000; Wijsman et al., 2001) and fresh water sediments (Roden and Wetzel, 1996). Concentrations of dissolved Fe are higher than those determined for marine sediments (e.g. Haese et al., 2000; Thamdrup et al., 1994), but are between those measured in saltmarsh sediments (Koretsky et al., 2003; HCO 3,

not possible to identify Fe oxide phases by XRD, albeit against the high background imposed by the organic-rich nature of the Quaking Houses sediments. Solid phase S geochemistry is shown in Fig. 5. There are few clear trends in the data; in some cases AVS dominates CRS, and vice versa. Elemental S accounts for more than half the total reduced S at the surface of the September 2002 core; more typically it comprises 10% of total reduced S in the November and January cores. 4. Discussion Whilst sharing many common characteristics, the sediment from an artificial wetland such as Quaking Houses differs from natural sediments in two important ways. Firstly, as an engineered system, the organic C-rich sediment (compost) is installed en masse during construction. Potentially reactive organic C is thus present throughout the sediment, in contrast to natural sediments where reactive organic C is added to the system over time at the sediment–water interface. Equally, any material which is deposited subsequent to commission of the wetland can only be incorporated deeper into the sediment through processes such as benthic mixing, diffusion and advection. Secondly, because the chemical composition of the influent and effluent of the wetland is monitored, the approximate extent to which elements have been retained by or lost from the wetland sediments is known. At Quaking Houses, changes in water chemistry across the wetland show that Fe and S have been gained by the sediment, and that bicarbonate alkalinity has been gained by the overlying water. It is also known that both Fe and S occur throughout the sediment profile and not just in the

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Lowe et al., 2000) and certain acid mine lake sediments (Blodau et al., 1998; Peine et al., 2000). Unlike the seasonally affected profiles, aqueous Fe(II) diffuses up to the sediment/water SO2 4 interface throughout the year, as well as diffusing further down into the sediment. Comparisons of influent and effluent Fe concentrations show that Fe is sequestered into the wetland sediments throughout the year (Table 1; Jarvis, 2000; Morrison, 2005). On average, therefore, Fe diffusing towards the sediment–water interface does not return to the water column but must be significantly re-oxidized to Fe oxide within the top 2 cm, where the highest concentrations of sedimentary Fe oxide also occur (Fig. 4). The decline in dissolved Fe below the maximum indicates loss of Fe to solid phases deeper in the sediment. In marine and freshwater sediments without vegetation, downward diffusion of dissolved Fe is driven by the deeper formation of solid phase Fe sulfides or carbonates. In Quaking Houses sediments, a significant proportion of the Fe throughout the sediment column occurs as oxides or oxyhydroxides, with a much smaller percentage as sulfides (Fig. 4). Furthermore, 20–70 lM concentrations of Fe(III) occur in pore waters throughout the sediment cores, and elemental S is also present at all depths. All these parameters point to an oxidative biogeochemical pathway in the wetland sediment. The occurrence of ‘dissolved’ Fe(III) is provocative since the equilibrium concentration at the pH of these pore waters is many orders of magnitude less than the measured values. Although it cannot be ruled out that the results are an analytical artefact, it is also possible that the Fe(III) is present as sub 0.45 lm colloids, which are common in mine waters (Mudashiru et al., 2008) and have been suggested to occur in pore waters from other wetlands receiving acidic mine waters (Peiffer et al., 1999), potentially stabilized by organic ligands (Taillefert et al., 2000). It is proposed that the oxidation of Fe in the deeper parts of the wetland is related to radial O2 loss from the roots of the dense populations of T. latifolia and P. australis within the wetland. Similar conclusions have been reached by others in a variety of wetland environments (Crowder and Macfie, 1986; Emerson et al., 1999; Batty and Younger, 2002; Gribsholt and Kristensen, 2002; Colmer, 2003; Sundby et al., 2003; Weiss et al., 2004). T. latifolia has been shown to be particularly effective in supplying O2 to root tips and the surrounding pore water (Michaud and Richardson, 1989), and the rhizosphere has been shown to be a region of intense, microbially driven Fe-cycling at very fine spatial resolution (Emerson et al., 1999; Weiss et al., 2003, 2004). Although there are no direct measurements of radial O2 loss from plant roots, the authors have observed macroscopic oxide coatings on plant roots and have seen microscopic oxide annuli in backscattered electron micrographs which may also be root coatings (Morrison, 2005). Since the pore water profiles indicate that diffusive transport dominates advection or bioturbation, supply of O2 through roots is the most likely explanation of the extensive occurrence of oxidized Fe phases in these sediments. The root system effectively acts as an Fe pump, drawing dissolved Fe(II) from pore waters near the sediment–water interface into the rhizosphere, where it is oxidized and precipitated as Fe oxides or oxyhydroxides. Whilst this is an effective way of redistributing and trapping Fe in the wetland sediments, there will be little overall gain in alkalinity or pH as the reductive and oxidative parts of the Fe cycle are quite closely balanced through reactions such as:

CH2 O þ 4FeOOH þ 7Hþ ! 4Fe2þ þ HCO3 þ 6H2 O 1 1 Fe2þ þ O2 þ Hþ ! Fe3þ þ H2 O 4 2 Fe3þ þ 2H2 O ! FeOOH þ 3Hþ

ð4Þ ð5Þ ð6Þ

5. Summary and conclusions The Quaking Houses wetland has successfully remediated acidic, Fe-rich mine waters for over a decade (Jarvis and Younger, 1999; Jarvis, 2000); this study has revealed some of the biogeochemical processes underpinning the remediation process over an annual cycle. Transport processes occurring at and across the sediment– water interface are sufficiently rapid on the 20 h residence time of the waters to: (a) remove most influent Fe and some SO2 4 into surface sediments and (b) increase both the pH and alkalinity of effluent waters. Sulfate and Fe reduction occur throughout the sediments. Rates of Fe reduction exceed rates of SO2 4 reduction in the top 2 cm of the sediment, resulting in pore water Fe concentrations up to 1 mM. Iron diffuses both up and down from the pore water maximum; upward diffusing Fe is oxidized and precipitates as Fe oxides in the top 2 cm of the sediment. Downward diffusing Fe also precipitates largely as Fe oxides, probably as the result of radial O2 loss from roots in the dense rhizosphere of the wetland plants Typha and Phragmites. The presence of elemental S throughout the sediments also indicates an oxidative part to the S cycle. The rhizosphere thus plays a critical role in the downward redistribution of Fe in the sediments, and in the overall C–S–Fe redox cycle of the wetland. Pore water SO2 4 profiles show that S is added to surface sediments in summer but lost during the winter months. Iron is gained throughout the year, resulting in the accumulation of 5–10 times more Fe oxide than reduced S in the wetland solid phases. The C–S–Fe redox reactions inferred in this study all generate or consume alkalinity and protons; the balance of reactions in the Quaking Houses sediments is such that pH is sustained between 7.2 and 7.8 and bicarbonate alkalinity is generated within the pore 2 waters. Much of the HCO 3 results from SO4 reduction; although Fe is reduced, re-oxidation results in little net gain of alkalinity. Iron oxides do however play an important role in ensuring that reduction is retained the dissolved sulfide resulting from SO2 4 within the sediment and does not re-oxidize to SO2 4 and regenerate protons; the coupling of the Fe and SO4 cycles are thus central to effective remediation. Acknowledgements This work was funded by a NERC Ph.D studentship to Michelle Morrison and EPSRC Grant No. GR/S07247/01. We thank Byrne Ngwenya and three other anonymous reviewers for their thoughtful comments which substantially improved this paper. References Batty, L.C., Younger, P.L., 2002. Critical role of macrophytes in achieving low iron concentrations in mine water treatment wetlands. Environ. Sci. Technol. 36, 3997–4002. Berner, R.A., 1985. Sulfate reduction, organic-matter decomposition and pyrite formation. Phil. Trans. R. Soc. London. Ser. A-Math. Phys. Eng. Sci. 315, 25– 38. Blodau, C., Hoffmann, S., Peine, A., Peiffer, S., 1998. Iron and sulfate reduction in the sediments of acidic mine lake 116 (Brandenburg, Germany): rates and geochemical evaluation. Water Air Soil Pollut. 108, 249–270. Canfield, D.E., Thamdrup, B., Hansen, J.W., 1993. The anaerobic degradation of organic-matter in Danish coastal sediments – iron reduction, manganese reduction, and sulfate reduction. Geochim. Cosmochim. Acta 57, 3867–3883. Colmer, T.D., 2003. Long-distance transport of gases in plants: a perspective on internal aeration and radial oxygen loss from roots. Plant Cell Environ. 26, 17– 36. Crowder, A.A., Macfie, S.M., 1986. Seasonal deposition of ferric hydroxide plaque on roots of wetland plants. Can. J. Bot. 64, 2120–2124. Duan, W.M., Coleman, M.L., Pye, K., 1997. Determination of reduced sulphur species in sediments – an evaluation and modified technique. Chem. Geol. 141, 185– 194.

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