Fate and distribution of arsenic in laboratory-scale subsurface horizontal-flow constructed wetlands treating an artificial wastewater

Ecological Engineering 37 (2011) 1214–1224

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Fate and distribution of arsenic in laboratory-scale subsurface horizontal-flow constructed wetlands treating an artificial wastewater Khaja Zillur Rahman a , Arndt Wiessner b , Peter Kuschk b,∗ , Manfred van Afferden a , Jürgen Mattusch c , Roland Arno Müller a a b c

Centre for Environmental Biotechnology, UFZ–Helmholtz Centre for Environmental Research, Permoserstrasse 15, 04318 Leipzig, Germany Department of Environmental Biotechnology, UFZ–Helmholtz Centre for Environmental Research, Permoserstrasse 15, 04318 Leipzig, Germany Department of Analytical Chemistry, UFZ–Helmholtz Centre for Environmental Research, Permoserstrasse 15, 04318 Leipzig, Germany

a r t i c l e

i n f o

Article history: Received 28 July 2010 Received in revised form 20 February 2011 Accepted 20 February 2011 Available online 26 March 2011 Keywords: Arsenic Artificial wastewater Juncus effusus Mass balance Retention Subsurface horizontal-flow constructed wetland

a b s t r a c t Knowledge regarding the fate, accumulation and distribution of arsenic inside constructed wetlands is still insufficient. Based on a complete mass balance analysis, the aim of this study was to investigate the fate and distribution of As in distinct wetland compartments and different segments along the wetland gradient. Experiments were carried out in laboratory-scale wetland systems, two planted with Juncus effusus and one unplanted, using an As-containing artificial wastewater. The obtained results revealed that the planted wetlands have a substantially higher As-mass retention capacity (59–61% of the total As inflow) than wetlands without plantation (only 44%). However, different loads of organic carbon within the inflowing artificial wastewater showed no remarkable influence on As-mass retention in the planted wetlands. Nearly 47–52% of the total inflowing As mass was found to be retained within the first half of the planted wetlands and this retention decreased step by step along the flow path. In contrast, only 28% of the total inflowing As mass was retained within the first half of the unplanted wetland. In general, a different fate and distribution of As was observed inside the planted and unplanted wetlands. Higher As concentrations were exhibited by the plant roots (51.5–161.5 mg As kg−1 dry wt.) compared to the shoots (1.1–6.4 mg As kg−1 dry wt.). Analysis of the total As-mass balance in the planted wetlands revealed that nearly 44–49% of the total inflowing As was recovered or concentrated within the plant roots, only 1% was sequestered within the plant shoots, 7–10% were entrapped or deposited within the gravel bed sediments, 2–3% were retained in the standing pore water, 39–41% were flushed out as outflow and the remaining 1–2% is still considered to be unaccountable. Total As accumulation in the plant shoots made a small contribution to the mass balance, and plant root biomass was found to be the most important compartment for As retention. In contrast, nearly 11% of the total inflowing As were found in the sediment, 2% in the standing pore water, 57% in the outflow and a substantially higher portion (nearly 30%) remained unaccountable in the unplanted bed, which might be released as volatile As compounds or lost from the system due to various unknown reasons. The results indicate that plants have a remarkable effect on As retention and stability of already retained As; hence planted wetlands might be a suitable option for treating As-contaminated wastewater. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Arsenic contamination appears to be widespread in soil and water due to anthropogenic activities or from geogenic sources. In recent years, there has been an increasing contamination of water, soil and crops by this metalloid in many regions of the world (Fitz and Wenzel, 2002; Tripathi et al., 2007), particularly in some countries of southern Asia (Abedin et al., 2002; Meharg,

∗ Corresponding author. Tel.: +49 341 235 1765; fax: +49 341 235 1471. E-mail address: [email protected] (P. Kuschk). 0925-8574/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2011.02.016

2004). The situation becomes markedly problematic in the regions of West Bengal and Bangladesh, where several millions of humans are exposed to drinking, ground and surface water contaminated with arsenic (Chowdhury et al., 1999). The adsorption and coprecipitation of arsenic on hydrous oxides of metals and iron sulphides is an important sink for arsenic fixation (Jacks et al., 2002). Terrestrial plants are able to accumulate arsenic to a substantial extent (Visoottiviseth et al., 2002). The so-called hyperaccumulators take up more than 1000 mg kg−1 dry weight of the pollutant (Brooks et al., 1977). Mkandawire and Dudel (2005) demonstrated in another study that aquatic macrophytes can also

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be used for water treatment, because some species can accumulate arsenic from water to a greater extent. Inorganic arsenic species have been studied extensively in terms of uptake and accumulation by aquatic macrophytes (Mkandawire and Dudel, 2005; Robinson et al., 2003). Constructed wetlands (CW) are a low-cost, natural technology for wastewater treatment and have been successfully applied for treating various pollutants such as organic compounds, nutrients as well as heavy metals with promising treatment performance (Faulwetter et al., 2009; Garcia et al., 2010; Knight et al., 2000). In the past decades, many research studies have demonstrated the potentials of CW in the removal of several heavy metals, such as cadmium, chromium, copper, lead, nickel, and zinc (Cheng et al., 2002; Maine et al., 2009; Vymazal, 2005). The study of Buddhawong et al. (2005) demonstrated that more than 95% of As at an initial concentration of 0.5 mg l−1 in the water was removed in their batch experiment with lab-scale CW units planted with Juncus effusus. Another study by Rahman et al. (2008a) showed that the immobilisation of arsenic was more stable in planted beds than in an unplanted bed. Both systems (planted and unplanted) were suitable to treat As-containing wastewater particularly under microbial sulphate reducing conditions. The behaviour of metals in aquatic systems is complex and may include interactions among or between the major wetland compartments, above-ground plant parts, roots, litter, biofilms, soil, and water (Kadlec and Knight, 1996). Volatilisation of metals into a gaseous phase occurs with mercury, selenium, and arsenic to a lesser degree. Dissolved metals can adsorb onto particles, or exist complexed to inorganic and organic ligands, or be present in solution in the free-ion state. Chemical reactions, such as acid–base, precipitation, complexation, oxidation/reduction, and sorption, all play a role in removing metal ions from the water column, resulting in a metal-ion complex more or less rapidly settling to the sediments (Yong, 1995). Moreover, dissimilatory reduction caused by iron- and sulphate-reducing bacteria is widely considered the primary mechanism responsible for the rapid As reduction and release observed in anaerobic environments (Islam et al., 2004; Kirk et al., 2004). In the past, little attention was paid to the accumulation and distribution of the total As mass within the major wetland compartments and also in different segments along the flow path. Rahman et al. (2008a) showed redox transformation and dynamics of arsenic species in laboratory-scale subsurface horizontal-flow constructed wetlands treating an artificial wastewater. However, it is necessary to understand the mass balance of As in treatment wetlands to realise As-retention capacities, both in the presence and absence of wetland plants. Singhakant et al. (2009) studied the removal of As in pilot-scale subsurface-flow constructed wetlands by comparing constructed wetland units with and without vetiver grass in order to determine the roles of vetiver grass affecting As removal. But in fact, no sufficient information directly addresses the fate of As under constructed wetland conditions and the distribution of As within wetland gradients along the flow path. Thus, several essential questions remain that have not yet been comprehensively addressed. Within the total frame of our research work in this field, several fundamental aspects and mechanisms of As fixation, influences of dynamic redox processes on As biotransformation, long-term As stability, physiological reactions and responses of plants to As, bioaccumulation and uptake of As and sulphur in plant biomass, sludge sediment analysis, mass-balance of As and S, etc., were investigated under laboratory-scale constructed wetland conditions. The main objectives of this study were (a) to investigate the distribution of total As in distinct wetland compartments and different segments along the flow path; (b) to use a mass balance approach to iden-

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tify the main removal pathways of As in constructed wetlands; (c) to identify the role and importance of wetland plants in removing As; and (d) to assess the feasibility of applying a constructed wetland system to treat As-containing municipal wastewater. All these processes were closely investigated in both planted and unplanted laboratory-scale wetlands using an artificial wastewater which resembled a secondary As-containing domestic effluent. 2. Materials and methods 2.1. Experimental design: laboratory-scale wetlands The three experimental wetlands used in this study consisted of plastic containers (length: 100 cm, width: 15 cm, height: 35 cm). The containers were uniformly filled with 65.7 kg gravel (with a diameter of 2–6 mm, a density of 1.665 g cm−3 and a porosity of 40%) up to a height of 30 cm. The water level was adjusted to 5 cm below the surface of the gravel bed and thereby a free pore water volume was calculated as 15 l. Sieves of perforated stainless steel were placed 3 cm in front of the inflow and outflow of the gravel bed. This free liquid volume was able to ensure an equal distribution of the inflow and a laminar liquid flow through the gravel bed. The schematic diagram of laboratory-scale subsurface horizontal-flow constructed wetland is illustrated in Fig. 1. Two wetlands (W1 and W2) were planted with rush (J. effusus) and one wetland remained unplanted (W3) as a control. The planted wetlands were loaded with different amounts of organic C in order to evaluate the influence of C loading on As retention. 2.2. Experimental conditions The experiments were conducted over a period of 491 days and wetlands W1, W2 and W3 were fed with As-containing artificial wastewater characterized by a hydraulic retention time of 3 days and defined concentrations of As(V) in the influent. In general, an inflow rate of 5 l d−1 was adjusted for all systems, which corresponds a hydraulic loading rate of 33.3 mm d−1 . Different experimental phases were carried out to investigate both how the dynamics of As was influenced by the different redox conditions (e.g. C-deficient and oxidised conditions, C-surplus and reducing conditions with defined SO4 2− concentrations and the discontinuing of organic C and As(V) dosages with only traces of SO4 2− , etc.) and how the stability of immobilised As varies within the wetlands (Rahman et al., 2008a). Each wetland was fed separately from a feeding tank (50 l). In order to keep an anoxic environment inside the feeding tank containing the artificial wastewater, a continuous purging of nitrogen gas (N2 ) through the headspace of the feeding tank was maintained throughout the whole operation period (Fig. 1). Before starting the experiment, wetlands (W1 and W2) were planted uniformly with J. effusus in August 2006 with a mean shoot density of approximately 800 and 733 shoots m−2 , respectively. By October 2006, the plant shoots were well established and covered the entire surface of the model wetlands. Feeding with artificial As-containing wastewater started from November 2006 after confirming the same initial conditions in both planted wetlands. The model wetlands were placed in a greenhouse with a day length of 16 h and were operated under defined environmental conditions with a temperature of 16–24 ◦ C simulating an average summer day in moderate climates (Wiessner et al., 2005b). One lamp (Master SON-PIA 400W, Phillips, Belgium) per wetland was switched on during daytime as an additional artificial light source whenever the natural light fell below 60 klx. The operational period lasted from November 2006 to March 2008.

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Fig. 1. Schematic diagram of a laboratory-scale subsurface horizontal-flow constructed wetland: (1) feeding storage tank; (2) pump; (3) inflow distribution chamber; (4) gravel bed; (5) outflow chamber; (6) plant biomass; (7) flow meter; (8) outflow storage tank.

The artificial wastewater that was prepared and used in this investigation contained As and simulated a typical secondary domestic effluent with a theoretical COD value of about 340 mg l−1 and a TOC of 122.3 mg l−1 , which was derived from a combination of two different organic carbon sources (acetate and benzoate) and SO4 2− (approx. 50 mg S l−1 ). The inflow concentrations of the used ingredients were (in mg l−1 ): 204.9 CH3 COONa, 107.1 C6 H5 COONa, 117.8 NH4 Cl (30.8 N), 28 K2 HPO4 (5 P), 7 NaCl, 3.4 MgCl2 ·6H2 O, 4 CaCl2 ·2H2 O, 0.89 Na2 SO4 (0.2 SO4 2− –S, when traces of SO4 2− were applied), 0.2 As(V) (Titrisol® As2 O5 in H2 O, Merck, Germany), and 1 ml l−1 of trace mineral solution which was adapted from Buddhawong et al. (2005). These compounds were dissolved both in tap water (providing approx. 50 mg l−1 of SO4 2− –S, experimental phase I) and in deionised water (when traces of SO4 2− –S were supplied, experimental phases II and III). The operation conditions in all the phases within the model wetlands W1, W2 and W3 are listed in Table 1. Artificial wastewater was freshly prepared every 3 days to prevent microbial degradation during storage and operation (Fig. 1). N2 gas was vigorously purged through and bubbled out of the liquid phase of the wastewater for approximately 20–25 min after each preparation of fresh feeding solution in order to remove any traces of dissolved oxygen from the feeding tank. From the start through to day 267 of the experiment (phase I), all the wetlands were fed by a continuous inflow of As-contaminated

artificial wastewater (200 ␮g As l−1 ) along with organic C (with a COD of ∼340 mg l−1 in the planted W1 and the unplanted W3; a COD of ∼680 mg l−1 in planted W2 wetland) under reducing conditions (with an Eh ∼ from −84 to −189 mV). In experimental phase II, the inflow of As dosage and organic C were stopped until day 372 (Table 1). During this time (from day 267 to day 372), only artificial wastewater with traces of SO4 2− –S (0.2 mg S l−1 ) and other reagents (except As and organic C dosage) was fed under highly oxidized conditions (with an Eh from ∼305 to 714 mV) in all the three wetlands. After day 372, the supply of As(V) and organic C dosage started again in the inflow (phase III) until day 463 (with an Eh from ∼30 to −33 mV). Maintaining the same organic C dosage as in phase III, the supply of As(V) and traces of SO4 2− –S (0.2 mg S l−1 ) were completely stopped until the termination of the experiment on day 491 (phase IV). During the whole experimental period, the growth status of the plant biomass in terms of green and healthy shoot density increased to a maximum of 14,693 and 13,353 shoots m−2 in planted wetlands W1 and W2, respectively. 2.3. Sample collection and parameter analysis Pore water samples were collected on a weekly basis at a depth of 15 cm below the wetland surface and five consecutive locations at 25 cm intervals along the flow path at 0, 25, 50, 75 and 100 cm

Table 1 Operation conditions (phases I, II, III and IV) within the model wetlands W1, W2 and W3 accomplished by defined arsenic, organic carbon and sulphate inflow concentrations (mean ± SD) of the artificial wastewater. Wetland

Parameter

Unit

Experimental phases I

n

II

III

n

IV

n

W1 (planted)

As(V) SO4 2− –S TOC

␮g l−1 mg l−1 mg l−1

195 ± 14 45 ± 2 104 ± 33

30 30 22

bdl 0.2 ± 0.1 bdl

14 10 8

190 ± 29 0.2 ± 0.1 101 ± 3

13 10 10

bdl 0.2 ± 0.1 114 ± 2

6 6 5

W2 (planted)

As(V) SO4 2− –S TOC

␮g l−1 mg l−1 mg l−1

190 ± 13 45 ± 3 244 ± 24

30 30 22

bdl 0.2 ± 0.1 bdl

14 10 6

174 ± 33 0.2 ± 0.1 114 ± 30

13 10 10

bdl 0.2 ± 0.1 112 ± 3

6 6 5

W3 (unplanted)

As(V) SO4 2− –S TOC

␮g l−1 mg l−1 mg l−1

196 ± 14 45 ± 2 104 ± 33

30 30 22

bdl 0.2 ± 0.1 bdl

14 10 7

188 ± 33 0.2 ± 0.1 102 ± 12

13 10 10

bdl 0.2 ± 0.1 112 ± 2

6 6 5

bdl: below the detection limit (<0.3 ␮g As l−1 ; <1 mg TOC l−1 ), n: number of samples.

n

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from the inflow with a syringe and a long needle (Fig. 1). Sampling locations at 0 and 100 cm represented the inflow and outflow of the wetlands, respectively. The needle was rinsed in advance with deionised water and also with N2 gas between each sampling in order to prevent auto-oxidation of sample ingredients. The preservation technique of the collected samples for measuring the total arsenic content and the analytical procedure of hydride generation atomic adsorption spectrometry (HG-AAS) have already been described by Daus et al. (2002) and Schmidt et al. (2004). The volatile arsenic species were analysed by gas chromatography/mass spectrometry (GC–MS, Shimadzu-GC-17A-Shimadzu GC-MS-Qp5000) with electron ionisation and quadrupole analyser, using the method described by Pantsar-Kallio and Korpela (2000). The analyses were performed isothermally at 50 ◦ C, and helium was used as a carrier gas. A SenTix ORP electrode connected to a multiline P4 (WTW, Germany) and pH-electrode Sentix 41 were used for measuring the redox potential (Eh ) and pH, respectively. The proper functioning of the electrodes was tested and calibrated regularly using WTW solution and standard pH buffer (pH 4.01 and pH 7.00) solutions for redox potential (Pt/Ag/AgCl in 1 M KCl, +220 mV/25 ◦ C) and pH measurement, respectively. Evapotranspiration was measured by balancing the inflow and outflow amounts of water. The total amount of water loss (data not shown) was divided by the time and the area to calculate the specific water loss (evapotranspiration rate). After the termination of the experiment, plant biomass samples (shoots and roots) and sludge sediments were collected from each wetland segment of 0–25, 25–50, 50–75 and 75–100 cm in order to investigate which segment of the wetlands had the highest potential for As-removal efficiency. Plant samples were sectioned into their shoot and root components after collecting them from each segment. The segments were manually divided carefully by inserting a sharp galvanized steel sheet or plate (with a length of 35 cm, a width of 15 cm, and a thickness of 0.25 cm) one after another to mark the correct segment length so that the plant biomass sample, gravel or even pore water could not pass through from one segment to another. The roots were first thoroughly washed with tap water and then with deionised water to remove any gravel aggregate or sludge sediment. The plant shoots and roots collected from each wetland segment were freshly weighed, dried at 105–108 ◦ C for 3 d, allowed to cool, and then the dry weights were determined and the water content was calculated. These dry weights are used throughout the text unless otherwise specified. The dried samples were ground to a fine powder using a mortar and pestle under liquid nitrogen in order to obtain a homogeneous sample and then preserved in sealed plastic bottles for analysis. For the analysis of the total arsenic concentration in plant biomass (shoots and roots), the homogenised powdered samples were digested by microwave extraction (PE Anton Paar GmbH, Graz, Austria). For digestion, 2 ml of digestion mixture (HNO3 :HCl = 4:1) were added to 0.5 g powder in a Teflon pressure bomb and heated to 260 ◦ C for 1 h. After the digest had cooled down, it was filled with deionised water to a total volume of 10 ml, mixed and filtered using a 0.45 ␮m syringe filter (Satorius AG, Goettingen, Germany). The filtrate solution was analysed for total As by using hydride generation atomic absorption spectrometry (HG-AAS) with a detection limit of 0.3 ␮g As l−1 . Acid blanks were analysed in order to assess possible contamination. All analyses were performed in duplicate. Analysis of the sludge sediment samples was performed by the energy dispersive X-Ray fluorescence (EDXRF) spectrometer XLAB 2000 (SPECTRO Instruments) running with the software package XLAB Pro 2.2. Collected sludge sediments from each segment were dried at 105 ◦ C for 24 h using oven MA4O (Satorius, Germany) and

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Fig. 2. Cumulative total As-mass inflow and retention in the three laboratory-scale constructed wetland systems (W1, W2 and W3) during the whole experimental period of 491 days.

were ground by means of an agate ball mill (Retsch). Well-ground sample material (1 g) was mixed with stearine wax (Hoechst, Germany) for XRF-analysis as a binder in a ratio of 80:20 (w/w) and subsequently pressed at 150 MPa to pellets (with an internal diameter of 32 mm) and analysed by means of energy dispersive Xray fluorescence analysis (EDXRF). The mean value of two replicates was calculated. The relative error of the method was 2–3%. 2.4. As mass balance calculations After 491 days of the experiment, a complete mass balance of As was investigated in each wetland unit and also within the intermediate segments (at 25 cm intervals along the flow path of each bed) by considering the total As-mass input, the total As-mass output, and the total As retained in the plant biomass (the shoots and the roots), in the pore water and in the sludge sediments. The remaining (loss or gain) of the As mass from the mass balance calculation was considered to be unaccountable. The total As-mass input and output in each wetland was calculated from the cumulative total As-mass inflow and outflow during the whole operation time period (Fig. 2). The same methodology was applied for calculating the total As-mass inflow and outflow for the intermediate segments (0–25, 25–50, 50–75 and 75–100 cm) of each wetland. Therefore, the outflowing As mass of one particular segment was the inflowing As mass of the next segment and so on. The total wet-

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land bed is represented as 0–100 cm in the whole text. In both cases (within a segment or in the total wetland), a simple As-mass balance was computed according to the following Eq. (1) (Singhakant et al., 2009): Asin = Asout + Asplant + Assed + Aspw + Asunaccount

(1)

where Asin is total As mass in the inflow (g); Asout is total As mass in the outflow (g); Asplant is total As mass in the plant biomass (g), including plant roots and shoots; Assed. is total As mass retained in the sludge sediment (g); Aspw is total As mass retained in the pore water or standing water (g); Asunaccount is the total As mass that was unaccountable (g) includes the loss or gain of As from the mass balance calculation. 3. Results 3.1. As-removal efficiencies of model wetlands Fig. 2 shows the total amount of the As-mass inflow and the total As retention in all three corresponding wetlands during the whole study period of 491 days. A total inflow mass of 335.2, 323.2 and 335.8 mg As was fed and a total of 138.8, 127.3 and 189.7 mg As mass was flushed out through the outlet of wetland systems W1, W2 and W3, respectively. Therefore, a cumulative total mass of 196.4, 195.9 and 146.1 mg As was retained, which resulted in nearly 59%, 61% and 44% of the total As-mass retention in the corresponding wetlands W1, W2 and W3, respectively. During the two stoppage periods of As(V) supply (phases II and IV), a total amount of at least 26.04 mg As mass was flushed out or remobilized from the unplanted wetland W3. During these stoppage times, no remarkable loss of the total As mass was observed in the two planted wetlands, W1 and W2. Within the first 267 days of operation (phase I), the corresponding area specific mean removal rate (which considered water loss due to evapotranspiration) had values of 3.76 ± 0.8, 3.52 ± 1.2 and 3.9 ± 1.5 mg As m−2 d−1 and thus contributed to a mean As retention of 63%, 62% and 67%, in wetlands W1, W2 and W3, respectively. In contrast, during the experimental period from day 372 to day 463 (phase III), the area-specific mean removal rate data demonstrated a slightly higher As retention of 77% and 69% in W1 and W2, respectively and only 62% in the unplanted W3. Only traces (2–3 ␮g As l−1 ) of volatile arsenic compound [gaseous arsine (AsH3 )] was found in these two experimental phases and only in the planted wetlands W1 and W2. 3.2. Changes of redox (Eh ) and pH Table 2 presents the summing-up of the changes in pH and the redox potentials (Eh ) within the corresponding wetlands W1, W2 and W3 during the whole operation period of 491 days. The results indicated a variation of mean pH and Eh values along the wetland gradient from the inflow. The mean inflow Eh values did not show any sudden fluctuations throughout the whole experimental period and were measured as a mean value of nearly 425 ± 34 mV at the inflow feeding zone in all wetlands. The mean outflow values were measured as 234 ± 44, 216 ± 35 and 72 ± 29 mV in the planted W1, W2 and in the unplanted W3 wetlands, respectively. At a 50 cm distance along the flow path, the mean redox values decreased to 177 ± 38 and 137 ± 73 mV in the planted wetlands W1 and W2, respectively. The mean redox potential values exhibited an increasing tendency after the first half (50 cm) of each planted wetland. In the unplanted wetland W3, the mean redox values decreased step by step along the flow path from the mean inflow value of

Fig. 3. Concentration of total arsenic in plant shoots (a) roots (b) and sediment (c) measured in each segment of subsurface horizontal-flow constructed wetlands W1, W2 and W3 (the values are means of two replications and the error bars are standard deviations; 0–100 cm represents the mean ± SD values of all four segments in each wetland).

424 ± 34 mV to 130 ± 16, 88 ± 32 and 71 ± 23 within the intermediate locations of 25, 50 and 75 cm, respectively. This decreasing tendency of the mean redox values along the unplanted wetland gradient was quite opposite to the redox dynamics in the planted wetlands. The mean inflow pH values were consistent to within 7.0 ± 0.5 in the inflow feeding zone of all model wetland systems but showed a decreasing tendency along the flow path in both planted wetlands. The mean pH values of the outflow zone were measured as 5.8 ± 1.4 and 5.7 ± 1.3 in the planted W1 and W2, respectively. In contrast, no significant changes were observed in the mean pH values along the flow path in the four segments of the unplanted wetland W3. The values within the planted wetlands were substantially lower in the range of 3.0–7.7 and 3.4–7.8 in W1 and W2, respectively, as compared to the range of 5.2–7.7 in the unplanted wetland W3 (Table 2). 3.3. Concentration of total As in plant biomass The analytical results of the total As concentrations in plant biomass are presented in Fig. 3. The total dry weight of the plant shoots were 0.13, 0.28, 0.30 and 0.26 kg, and the corresponding dry weights of the roots were 0.61, 0.54, 0.44 and 0.38 kg within the plant biomass collected from the four segments of 0–25, 25–50, 50–75 and 75–100 cm of the planted wetland W1, respectively. The mean total As concentration was measured as 3.15 ± 0.22,

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Table 2 Summary of average pH and redox (Eh ) values (mean ± SD) and ranges of the values (from minimum to maximum) in each wetland at the inflow and outflow and also at the intermediate points of a 25 cm interval along the flow path (each sample was taken at a depth of 15 cm from the wetland surface). Wetland

Parameter

W1 (planted)

pH Eh (mV)

W2 (planted)

pH Eh (mV)

W3 (unplanted)

pH Eh (mV)

Distances from the inflow (cm) 0 (Inflow)

25

50

75

100 (Outflow)

7.0 ± 0.5 (6.1–7.8) 425 ± 35 (335–492)

6.2 ± 1.1 (3.6–7.7) 187 ± 91 (−31 to 587)

6.0 ± 1.1 (3.4–7.6) 177 ± 38 (−143 to 595)

5.9 ± 1.4 (3.0–7.2) 220 ± 25 (−134 to 630)

5.8 ± 1.4 (3.2–7.3) 234 ± 44 (−139 to 677)

7.0 ± 0.4 (6.5–8.2) 424 ± 33 (340–490)

6.1 ± 1.2 (3.4–7.8) 140 ± 59 (−177 to 450)

5.9 ± 1.1 (3.2–7.7) 137 ± 73 (−110 to 484)

5.8 ± 1.0 (3.2–7.6) 147 ± 41 (−162 to 573)

5.7 ± 1.3 (3.4–7.7) 216 ± 35 (−157 to 695)

7.0 ± 0.5 (6.0–7.6) 424 ± 34 (335–470)

6.9 ± 0.7 (5.7–7.6) 130 ± 16 (−173 to 550)

7.0 ± 0.5 (5.2–7.6) 88 ± 32 (−188 to 435)

7.1 ± 0.4 (6.3–7.7) 71 ± 23 (−180 to 370)

7.1 ± 0.3 (6.4–7.6) 72 ± 29 (−190 to 375)

4.25 ± 0.21, 2.1 ± 0.14 and 1.1 ± 0.14 mg As kg−1 (dry wt.) in the shoots and was remarkably high in the roots of the respective four segments as 90 ± 4, 79 ± 6, 72 ± 4 and 51.5 ± 3.5 mg As kg−1 (dry wt.) in the planted wetland W1. Furthermore, the mean As concentration values were found to be 0.5 ± 0.14 and 2.25 ± 0.2 mg As kg−1 (dry wt.) in the control shoot and root samples of J. effusus, respectively. Similarly, in the other planted wetland, W2, the total dry weight of the shoots reached 0.16, 0.27, 0.18, 0.13 kg and the corresponding dry weights of the roots reached 0.29, 0.53, 0.28, and 0.24 kg within the segments of 0–25, 25–50, 50–75 and 75– 100 cm, respectively. The mean total As concentrations were analysed to be 6.4 ± 0.57, 3.15 ± 0.50, 2.17 ± 0.11 and 1.1 ± 0.14 mg As kg−1 (dry wt.) in the shoots and 91.5 ± 6.4, 161.5 ± 9.1, 108.5 ± 9.2 and 76 ± 8.5 mg As kg−1 (dry wt.) in the roots of the respective four segments along the flow path in the planted wetland W2. Within the total wetland (0–100 cm), the mean total As concentration was found to be 2.65 ± 0.2 and 3.2 ± 0.1 mg As kg−1 (dry wt.) within the shoots of W1 and W2, respectively. Similarly, mean values of 73 ± 4 and 109 ± 8 mg As kg−1 (dry wt.) were observed within the roots of W1 and W2, respectively (Fig. 3). 3.4. Concentration of the total As in the sediment After the termination of experiments, sludge was also collected from each segment (at distances of 25 cm along the flow path) of wetlands and the analytical results showed a mean total As concentration of 928 ± 9, 364 ± 11, 156 ± 8 and 86 ± 4 mg As kg−1 (dry wt.) in the respective four segments of 0–25, 25–50, 50–75 and 75–100 cm in the planted wetland W1 (Fig. 3). Similarly, a total As concentration with a mean value of 423 ± 11, 373 ± 9, 308 ± 6 and 175 ± 8 mg As kg−1 (dry wt.) was determined within the sediment collected from the respective four segments of the other planted wetland W2. In comparison to the planted wetlands, respective mean As concentration values of 571 ± 11, 283 ± 16, 213 ± 7 and 153 ± 15 mg As kg−1 (dry wt.) were found within the sediments collected from the corresponding four segments of the unplanted wetland W3. The mean As concentrations within the first 0–25 cm segment of the wetlands were 3–11 times higher than the As level within the sediment of the last 75–100 cm segment and were measured to be nearly 130–290 times higher than the control sediment samples. The mean As concentration was calculated as 383 ± 8, 320 ± 4 and 305 ± 7 mg As kg−1 (dry wt.) within the sediment of the respective total 0–100 cm wetlands W1, W2 and W3 (Fig. 3).

3.5. Distribution and mass balance of As Table 3 presents a summary of the total As-mass distribution in the different wetland compartments (the shoots, roots, etc.) and a complete mass balance calculation within the different wetland segments along the flow path over the whole operation period of 491 days. When considering the inflowing As mass as 100%, the percentile value of the outflowing As mass and retention were estimated. For example, a total mass of 335.2 mg As (100%) was loaded as an inflow, and a total mass of 248.8 mg As (74.2%) was flushed out at a 25 cm distance from the inflow in the planted wetland W1. The masses of As recovered in the pore water, the shoots, the roots and the sediments within this 0–25 cm segment were measured as 2.8, 0.4, 54.8 and 22.0 mg As, which resulted in 0.9%, 0.1%, 16.3% and 6.6% of the total inflowing As-mass retention, respectively. The remaining 6.3 mg As (nearly 1.9% of the total inflowing As mass) was considered to be unaccountable or retained in uncounted sinks. The outflowing As mass after 25 cm was considered as the inflowing As mass for the next 25–50 cm segment. Similar calculations were carried out for the other segments along the flow path of the planted wetland W1, and a complete mass balance of the total As was calculated within the wetlands W1, W2 and W3. Interestingly, it was observed that a range of 12–16% of the total inflowing As mass was recovered via the roots within the four segments of the planted wetland W1. The highest value of 34.9% of the inflowing As mass was recovered via the roots within 25–50 cm segment of in the planted wetland W2. On the contrary, in the unplanted wetland W3, a high amount of As mass (of 19.1 mg As) was trapped within the sediment of the first 0–25 cm segment and also a higher amount (of 32.5 mg As) of As mass was estimated to be unaccountable in the first segment, as compared to the other three segments. Based on this distribution and mass balance calculation in each segment, it was observed that a total of 47% and 52% of the total As mass were retained within the first 50 cm (the first half of the wetlands) along the flow path of the planted wetlands W1 and W2, respectively (Fig. 4). In contrast, only 28% of the total inflowing As mass was captured or retained within the first 50 cm in the unplanted wetland W3. Nearly 32% of the total As mass was retained within the second half of the planted wetlands, as compared to 25% in the unplanted wetland W3. After calculating the total As mass in the plant shoots, roots and sediments of all four segments of both planted wetlands, it was observed that nearly 44% and 49% of the total inflowing As mass were accumulated/recovered/concentrated within the roots and only 10% and 7% were entrapped or deposited within the gravel bed (as sediments) of the planted wetlands W1 and W2, respectively

243.9 (75.5%) 178.2 (73.1%) 159.1 (89.3%) 127.9 (80.4%) 127.3 (39.4%)

281.4 (83.8%) 248.5 (88.3%) 222.1 (89.4%) 190.6 (85.8%) 189.7 (56.5%)

38.1 (11.8%) −28.8 (−11.8%) −17.3 (−9.7%) 9.1 (5.7%) 1.7 (0.5%)

32.5 (9.7%) 22.5 (8.0%) 19.3 (7.8%) 26.6 (12.0%) 101.9 (30.3%)

Outflow (mg As)

248.8 (74.2%) 197.5 (79.4%) 154.6 (78.3%) 139.5 (90.2%) 138.8 (41.4%)

Unaccountable (mg As)

2.0 (0.6%) 3.2 (1.3%) 2.1 (1.2%) 1.2 (0.7%) 8.5 (2.6%)

2.8 (0.8%) 2.4 (0.8%) 1.8 (0.7%) 1.0 (0.4%) 8.0 (2.4%)

11.5 (3.6%) 5.3 (2.2%) 3.3 (1.9%) 2.7 (1.7%) 22.8 (7.1%)

19.1 (5.7%) 8.0 (2.8%) 5.3 (2.1%) 3.9 (1.7%) 36.3 (10.8%)

26.6 (8.2%) 85.2 (34.9%) 30.6 (17.1%) 18.1 (11.4%) 160.5 (49.6%)

– – – – – 335.8 (100%) 281.4 (100%) 248.5 (100%) 222.1 (100%) 335.8 (100%)

– – – – –

1.0 (0.3%) 0.8 (0.3%) 0.4 (0.2%) 0.2 (0.1%) 2.4 (0.7%) 323.2 (100%) 243.9 (100%) 178.2 (100%) 159.1 (100%) 323.2 (100%)

0–25 25–50 50–75 75–100 0–100

0.4 (0.1%) 1.1 (0.4%) 0.6 (0.3%) 0.3 (0.2%) 2.4 (0.7%)

0–25 25–50 50–75 75–100 0–100

Shoots (mg As) Total inflow (mg As)

0–25 25–50 50–75 75–100 0–100

W3 (unplanted)

W2 (planted)

Segment (cm)

Table 4 shows the removal efficiency of SO4 2− –S and TOC in different experimental phases within the corresponding wetlands W1, W2 and W3. A comparatively higher SO4 2− –S removal (79%) was observed in the unplanted wetland W3 in comparison to the two planted wetlands W1 and W2. It is important to note that only a small amount (32%) of TOC was removed from the planted wetland W2 in experimental phase I, whereas a remarkably higher TOC removal efficiency was observed in the planted W1 and the unplanted wetland W3. Interestingly, in the other experimental phases (III and IV), the unplanted wetland W3 indicated a higher TOC removal efficiency than the planted wetlands (Table 4). 4. Discussion 4.1. Comparison of As-retention capacities The results obtained for the laboratory-scale wetland systems clearly indicated that the planted wetlands had a (nearly 15%) Table 4 Removal efficiency of SO4 2− –S and TOC in different experimental phases (I, II, III and IV) within the model wetlands W1, W2 and W3.

W1 (planted)

Wet-land

(Fig. 5). Consequently, the total As accumulation in the plant shoots provided only a small contribution to the mass balance. Nearly 1% of the total As mass was recovered via the plant shoots in both planted wetlands. 2–3% of the total As mass were retained in the pore water and 1–2% were considered to be unaccountable. Nearly 39% and 41% of the total inflowing As mass were flushed out of the respective planted wetland systems W1 and W2. In contrast, it should be noted that in the unplanted wetland W3, only 13% of the total inflowing As mass was captured in the collected sediment and pore water and nearly 30% of the total inflowing As mass was unaccountable. The remaining 57% of the As mass was flushed out of the wetland bed system as an outflow. 3.6. Removal efficiency of SO4 2− –S and TOC

335.2 (100%) 248.8 (100%) 197.5 (100%) 154.6 (100%) 335.2 (100%)

2.8 (0.9%) 2.0 (0.8%) 1.4 (0.7%) 1.1 (0.7%) 7.3 (2.2%)

Sediment (mg As)

22.0 (6.6%) 5.2 (2.1%) 2.5 (1.3%) 2.8 (1.8%) 32.5 (9.7%) 54.8 (16.3%) 43.0 (17.3%) 31.9 (16.2%) 19.4 (12.6%) 149.1 (44.5%)

Fig. 4. As-retention capacity estimated as a percentage of the inflowing As mass in different segments of the model wetlands W1, W2 and W3 (0–100 cm represents the total As-retention capacity of the three corresponding wetlands).

Roots (mg As)

Pore water (mg As)

6.3 (1.9%) 0.1 (0.02%) 6.4 (3.2%) −8.4 (−5.5%) 5.2 (1.6%)

K.Z. Rahman et al. / Ecological Engineering 37 (2011) 1214–1224 Table 3 Distribution of total As in each wetland segment, which includes the total inflow and outflow mass of As and the distribution in the different wetland compartments (shoots, roots, sediment, etc.) over the whole operation period of 491 days. Data are given in mg As, and the values in parentheses are percentile amounts of total inflowing As.

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Wetland

Parameter

Removal efficiency (%) Phase I

Phase II

Phase III

Phase IV

W1 (planted)

SO4 2− –S TOC

62 86

– –

– 32

– 58

W2 (planted)

SO4 2− –S TOC

71 32

– –

– 38

– 49

W3 (unplanted)

SO4 2− –S TOC

79 71

– –

– 56

– 65

K.Z. Rahman et al. / Ecological Engineering 37 (2011) 1214–1224

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systems, most likely by As2 S3 precipitation with low solubility (Kim et al., 1999). The dynamic redox conditions within the wetlands presumably provided suitable conditions for the growth of more microbial biomass which indeed played a key role for the higher As retention in planted wetlands. Despite the probable re-oxidation of sulphur and the release of As, the plant-root activity (providing O2 to the rhizosphere) presumably contributed to a potentially strong immobilisation of As in the planted wetlands, as compared to the unplanted. Therefore, the plants play an important role for a better stabilisation of the immobilised As despite a probable re-oxidation of reduced compounds. Moreover, no particular trend of As(V) reduction and As(III) formation was monitored within the unplanted bed throughout the whole operation period (Rahman et al., 2008a). However, doubling the organic C dosage in the planted wetland W2 was not accompanied by drastic changes in the overall As-mass retention efficiencies, as compared to the planted wetland W1. Only 2% more As mass was retained in the planted W2 in comparison to the other planted wetland W1. TOC removal efficiency was remarkably low (only 32%) in the planted wetland W2, which clearly demonstrated a stressful condition for the plant biomass and presumably also for the microbial community. A slightly higher dissimilatory sulphate reduction (of 71%) in the planted wetland W2 compared to the other planted wetland W1 (62%) probably resulted in a slightly higher As-mass retention in the extremely overdosed organic C-enriched wetland (W2). The planted wetland W1 exhibited a slightly higher TOC removal efficiency (86%) than the unplanted wetland W3 (71%), which clearly indicated that more microbial activities, more redox reactions and consequently more As-mass retention occurred presumably in the planted wetlands, as compared to the unplanted ones. 4.2. Correlation of redox (Eh ), pH and As-mass retention

Fig. 5. Mass balance of As estimated as a percentage of inflowing total As mass in the model subsurface horizontal-flow constructed wetlands W1, W2 and W3.

higher As-mass retention capacity than the unplanted wetland (Fig. 4). This was also similar to the other study by Singhakant et al. (2009), where the wetlands planted with vetiver grass showed higher As-removal efficiencies than the unplanted wetland. This may be due to the fact that the wetland rhizosphere offers specific redox gradient (both micro- and macro-) conditions enabling the development of a highly diverse microbial consortia capable of different beneficial redox reactions (Bezbaruah and Zhang, 2004; Wiessner et al., 2005a). Moreover, slightly higher oxidised conditions due to the plant root-mediated oxygen release and the re-oxidation of reduced As species probably influenced higher sorption and precipitation reactions in the planted wetlands. All of these fundamental processes presumably contributed to both an efficient As-mass retention in terms of adsorption by microorganisms, plant roots, organic substrates and/or adsorption onto oxide minerals and a concomitant co-precipitation, specifically with Fe(III) oxyhydroxides. The reasons for the immobilisation of arsenic under oxic conditions were similarly explained by other authors (Bednar et al., 2005). Under reducing conditions, a constant SO4 2− –S inflow concentration probably initiated an intensive sulphate reduction by the dissimilatory sulphate-reducing bacteria and contributed to a highly efficient As-mass removal from the corresponding wetland

It is known that the reducing conditions lead to the mobilisation of arsenic as As(III) entering the liquid phase (Mok and Wai, 1994). This was clearly demonstrated within the wetlands in this study, where the reducing conditions resulted in an increase in the concentration of dissolved As. The addition of organic carbon remarkably changed the redox conditions due to a depletion of dissolved oxygen by microbial degradation of organic matter and facilitated a rapid decrease in the redox values to a minimum of −143, −177 and −190 mV within the different segments along the flow path in the respective wetlands W1, W2 and W3 (Table 2). Particular changes in redox dynamics from oxic to anoxic probably accelerated the reductive dissolution of Fe(III) oxyhydroxides and also the chemically and microbially mediated reduction of As(V) to more soluble As(III), which resulted in a probable release of arsenic within the wetlands (data not shown). Strictly reduced conditions could also have favoured the dissimilatory sulphate reduction to form dissolved sulphide which might have contributed towards arsenic-sulphide precipitation (most likely as As2 S3 ) under Csurplus conditions. It was evident that the microbial sulphate-reduction rate was higher in the unplanted wetland W3 compared to the planted ones, probably due to more reducing conditions (Eh ∼−170 to −190 mV) in the system (Table 2). Despite an efficient sulphate reduction and a consequently higher As-mass retention in the unplanted wetland, a clear re-mobilisation of As demonstrated the instability of the retained As mass after changing the dynamic redox conditions (Rahman et al., 2008a). During the stoppage period of As dosage and organic C loading (phase II), where all the systems were changing from reducing conditions to more rapid oxidising conditions, a remobilisation of As was clearly observed in the unplanted wetland (Fig. 2).

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In contrast, the already retained As mass was showing more stability and strong As bonding capacity within the planted wetlands W1 and W2, probably due to the presence of plant roots (phases II and IV, Fig. 2). Compared to the unplanted wetland, substantially higher mean redox potential values were observed within the four segments of both planted wetlands W1 and W2 (Table 2). These elevated Eh values clearly indicated that the planted wetlands exhibited more oxidized conditions than the unplanted wetland since the plants were directly involved in the transport of oxygen through their roots into the rhizosphere of wetland systems (Bezbaruah and Zhang, 2004; Stottmeister et al., 2003; Wiessner et al., 2005b). Oxic conditions in both planted wetlands presumably resulted in a better immobilisation of arsenic compared to the unplanted wetland. Several researchers studied the influence of the relationship of pH and redox conditions on As solubility and species transformations (Dzombak and Morel, 1990; Smedley and Kinniburgh, 2002). The lower pH level in both planted wetlands compared to the unplanted was probably caused by better ammonium oxidation, the release of organic acids (H+ ) from the root decomposition and sulphide oxidation, the plant uptake of K+ and the exchange for H+ , etc. These lowered pH values probably contributed to a highly favoured predominant As(V) adsorption and co-precipitation with Fe(III) oxyhydroxides (Waychunas et al., 1993). Hence, a better As immobilisation was observed in the planted wetlands, as compared to the unplanted wetland in this study. 4.3. Distribution of As in plant biomass The mean As concentrations (dry weight) within the plant shoots in both wetlands were nearly 6-fold higher than the As in the control shoot samples. However, the mean As concentrations within the shoots exhibited a decreasing tendency along the flow path (Fig. 3). For example, the mean As concentration within the shoots of the 0–25 cm inflow feeding zone was nearly 4–6fold higher than that of the shoots of the outflow zone in both planted wetlands. Therefore, it was observed that a higher amount of accumulated As was translocated into the plant shoots within the first half of the wetlands (0–50 cm), as compared to the second half (50–100 cm). Interestingly, different plant growth rates and shoot distributions were observed along the flow path of both planted wetlands during the whole experimental period. The mean shoot densities (of green and healthy shoots) were higher (data not shown) within the first half of the wetlands (0–50 cm) than in the second half (50–100 cm). Therefore, the plant growth rate might have a positive impact on the As uptake of the plant, the translocation or distribution into their biomass. In general, a nearly 12–30 times higher As concentration was exhibited by the plant roots compared to the shoots in both planted wetlands W1 and W2. This was also reported by other authors elsewhere (Buddhawong, 2005). This clearly indicated that the mobility of As in J. effusus was low. In this study, a vast majority of As appeared to be fixed in/on the roots and only a limited amount was translocated to the shoots. The concentrations of As within the roots of the inflow feeding zone were 25–30 times higher than those in the shoots (Fig. 3), suggesting that these inflow feeding zone roots were the main accumulation organ. Comparatively higher As concentrations were also obtained within the roots of the first half of the wetland compared to the second half in the planted wetland W1. A similar trend was also observed within the roots of the planted wetland W2, but a remarkably higher As concentration was observed in wetland W2 compared to the roots of W1. Doubling of the organic carbon (TOC) loading might be a probable reason for the higher As concentration

in the shoots and roots of the planted wetland W2 as compared to the other planted wetland bed W1. The highest mean As concentration of 161.5 ± 9.1 mg As kg−1 was observed within the roots of the 25–50 cm segment in the planted wetland W2, which was nearly twice the As concentration within the roots of the last 75–100 cm segment and nearly 72-fold higher than the control root sample (Fig. 3). Higher plant growth rates and dense root structure in the middle section of the wetland (data not shown) together with the doubling of the organic carbon (TOC) loading and an intensive dissimilatory sulphate reduction are probable reasons for the high As concentrations in the planted wetland W2, as compared to the other planted wetland W1. Accumulation and precipitation can occur on the surface of plant roots in the form of iron oxyhydroxides, which form the ironplaque attached around the plant root surface and in turn strongly adsorbs arsenic (Taggart et al., 2009; Rahman et al., 2008b). This iron plaque on the macrophyte roots has a high affinity for As, tending to have a higher affinity for As(V) than As(III) (Chen et al., 2005), and might be a probable reason for the higher As concentration within the roots in this study. Another mechanism might be that the roots may simply have been in increased contact with reduced forms of As in the sludge sediment, which have not been removed from the roots despite intensive washing. Or alternatively, the prevailing anoxic conditions around the less active/dead roots may have allowed arsenate [As(V)] associated with plaque to be reduced to arsenite [As(III)], a process perhaps driven by microbial activity (Jones et al., 2003; Weiss et al., 2003). Therefore, the microorganisms occurring in the rhizosphere, and/or the excretion of substances by the roots could cause a reduction of As(V) to As(III) in the direct root vicinity and an uptake by the plants or a reduction of As(V) to As(III) within the plant roots, which might contribute to the higher As concentration. 4.4. Distribution of As in sediment A wide variation of the total As level was observed within the sediments collected from different segments of all the three wetlands. A sharp decreasing tendency of the total As concentrations within the collected sediments along the flow path of the wetlands indicated the dynamics of As retention in the model wetlands (Fig. 3). Clear evidence of higher As adsorption and/or precipitation within the sediment of the first 25 cm (of the inflow feeding zone) was observed, as compared to the last 25 cm of the 75–100 cm outflow zone in all the corresponding wetlands. Therefore, the concentrations of total As were in an elevated level close to the inlet area and decreased along the wetland gradient more rapidly. Intensive dissimilatory SO4 2− reduction (data not shown) clearly indicated higher microbial activity within the first 25 cm and thereby a presumably higher arsenic–sulphide precipitation might increase the total As concentration within the sediment of the first segment, as compared to other segments. Vymazal and Krása (2003) also reported a decrease in concentration of metals in the sediment of a constructed wetland treating municipal wastewater. Sulphides were also observed in an elevated level within the first 25 cm (of the inflow feeding zone) of the wetlands, which was similar to the recent investigation by Wiessner et al. (2010), and presumably precipitation of arsenic as arsenic–sulphide complexes was thought to be an important removal process for As within the first 25 cm of the wetlands. Within the sediments of the total (overall) planted wetlands (0–100 cm), the mean As concentrations were slightly higher than in the unplanted wetland. This might be attributable to the fact that a higher adsorption and/or precipitation of As, high microbial activity, including other different beneficial redox reactions for the

K.Z. Rahman et al. / Ecological Engineering 37 (2011) 1214–1224

transformations of As species, were taking place within the planted wetlands (W1 and W2), as compared to the unplanted (W3) ones. 4.5. As-mass balance within the model wetlands There might be four possible fates for As removal from the model subsurface horizontal-flow wetlands in this study: (1) chelation or complexation by organic matter, biofilm and thus sequester or entrap into the gravel bed sediments, (2) accumulation or sequestration into plant biomass (shoots, roots, etc.), (3) retention in the pore water, and (4) volatilisation into the atmosphere. The remaining of the As mass was supposed to be flushing out of the wetlands through the outlet where it was collected. Based on the distribution of the total As-mass in the different wetland compartments, this study demonstrated the minor role of the plant shoots (nearly 1% by J. effusus) in terms of As uptake and a very low translocation efficiency from roots-to-shoots (amounting to 1.5–2%) in both planted wetlands, which is a common finding in studies of treatment wetlands (Batty and Younger, 2002). Total As content in underground parts was found to be much higher than that in above-ground parts. This finding was also in agreement with Singhakant et al. (2009). About 2% As accumulation in the plants from mass balance was reported in a study of a constructed wetland employing fourteen species of wetland plants in the removal of metals from electric utility wastewater (Ye et al., 2003). A substantially higher amount of As mass was accumulated in the roots and sediments of the inflow feeding zone (0–25 cm) in the planted wetland W1. This trend was also seen in the other planted wetland W2 and also within the sediment of the unplanted wetland W3. Nearly 50% of the total As-mass retention within the first half of the planted wetlands, as compared to only 28% in the unplanted one, clearly indicated the role of plant biomass in the planted wetlands. Microorganisms attached to the plant roots and biofilms can probably absorb higher amounts of As and, hence, a high concentration of As was found near the inflow feeding zone, where a high intensity of the microbial and chemical processes can be assumed. The other coincident mechanisms, such as precipitation, co-precipitation, complexion and adsorption, might also occur to a greater extent near the inflow. Substantially higher TOC removal efficiencies might also be indicative for higher chemical and microbial activities and, consequently, for a higher As-mass retention within the first half of the planted wetland W1, as compared to the unplanted wetland. The larger amounts (57% of the total inflow As mass) of As were flushed out through the outlet of the unplanted wetland, as compared to the planted (amounting to 39–41%) ones, which also pointed to a higher As-bonding capacity in the planted wetlands, presumably due to the presence of plant roots. Nearly 2% of the total As mass in both planted wetlands was unaccountable, which resembled a better mass balance of As in W1 and W2. On the contrary, in the unplanted wetland W3, a substantially higher amount of As (30% of total inflow As mass) was unaccountable or accumulated into an uncounted sink. This might be due to the fact that a substantial amount of As mass was presumably volatilised and released into the environment from the wetland surface. But only traces (2–3 ␮g As l−1 ) of inorganic volatile arsenic species [gaseous arsine (AsH3 )] were measured a few times within the pore water at different locations in both planted wetlands in this study. Nevertheless, no such volatile As species were detected within the water phase of the unplanted wetland. Another probable reason for this unaccountable As mass in the unplanted wetland might be that traces of sediments were coming out and precipitated in the outflow tank on several occasions during the experimental period. As-mass content within these small amounts of flushed-out sediments was neither analysed nor taken

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into account for the mass balance calculation in this study. Such a higher percentage (30% of total inflow) of unaccountable As mass raises various interesting questions and, therefore, more research is needed in order to investigate the fate of the unaccountable As mass in the unplanted wetland. From the previous results (Buddhawong, 2005), it was found that the gravel material could not absorb As in substantial amounts from the solution. Therefore, this kind of gravel itself had no remarkable impact on the mass balance of As in constructed wetland systems. Fractional analysis of deposited sediment might also be necessary to investigate different forms of As within the wetland beds. Singhakant et al. (2009) reported about the forms of the retained As by using sequential fractionation, which could indicate As complexation with Fe and Mn on the media surface of 31–38% and As trapping into the media of 42–52% of the total As. For long-term operation, special attention must be paid to the fate and As-accumulation dynamics within different compartments of constructed wetlands in order to avoid potential toxic effects. The accumulated arsenic could pose a serious threat to the wetland plants and sediments, nearby the water bodies and the environment. Suitable and cost effective technique should be developed and implemented for a highly efficient post-treatment method that includes a better handling and management strategy for the hazardous biosolids.

5. Conclusions A detailed study on the mass balance of As indicated that wetland plants play an important role by enhancing the Asretention capacities in subsurface horizontal-flow laboratory-scale constructed wetlands. The 491-day experimental results revealed that the planted wetlands have a higher As-mass retention capacity than wetlands without plants. Re-mobilisation of As is predominant while changing the redox conditions in the unplanted wetland, as compared to the minor risks for re-mobilisation and volatilisation of As in planted wetlands. Plant shoots (J. effusus) play a minor role in terms of As uptake or recovery in comparison with the plant roots. Therefore, these plant shoots can be commercially used since they do not accumulate total As mass to a greater extent. In general, higher As concentrations were exhibited by the plant roots compared to the shoots. Although the sediments showed the highest As concentrations among all the wetland compartments, remarkably higher amounts of As mass were retained or recovered by the plant roots within the planted wetlands in this study. Therefore, plant root biomass is the most important compartment for As retention. It was also shown that there was a substantial increase in the arsenic level within the inflow feeding regions (the first half of the wetlands) and decreased step by step along the wetland gradient. However, extremely overdosed organic carbon showed no remarkable impact on the As-mass retention capacities in the planted wetlands. Based on the As-mass balance calculation, it was evident that a substantially higher portion (nearly 30%) of total inflowing As mass was unaccountable in the unplanted bed, which might be released as volatile arsenic species or lost from the system due to various unknown reasons. The obtained results demonstrated the feasibility and a major step forward for applying subsurface horizontal-flow constructed wetlands for the treatment of secondary domestic As-contaminated effluent prior to a disposal to the receiving water bodies (rivers, lakes, etc.) or an application to agricultural fields for irrigation purposes. Designers of constructed wetlands need to be aware of the probable fate, impact, distribution and mass balance of trace metals present in applied wastewaters and sediments

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