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Development of a sulfidogenic bioreactor system for removal of co-existent selenium, iron and nitrate from drinking water sources
Journal of Environmental Management 254 (2020) 109757
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Research article
Development of a sulfidogenic bioreactor system for removal of co-existent selenium, iron and nitrate from drinking water sources Himanshu Sonkeshariya a, Arvind Kumar Shakya a, b, *, Pranab Kumar Ghosh a a b
Department of Civil Engineering, IIT Guwahati, Guwahati, India Department of Biotechnology, Rama University Uttar Pradesh, Kanpur, India
A R T I C L E I N F O
A B S T R A C T
Keywords: Selenium Iron Nitrate Selenium sulfide Achavalite Drinking water
The present study showed for the first time that selenium, iron, and nitrate could be simultaneously removed in a sulfidogenic bioreactor to meet drinking water standards. A bioreactor inoculated with mixed bacterial con sortium was operated for around 330 days in anoxic environment at 30 � C under varying combination of influent selenate (200–1000 μg/L as selenium), and iron (3–10 mg/L) in presence of 50 mg/L of nitrate. Required amount of acetic acid (as carbon source) and sulfate were supplied and the reactor was operated at different empty bed contact time (EBCT) of 45–120 min. Along with complete removal of nitrate, the reactor removed both selenium and iron to meet the drinking water standards. Field emission transmission electron microscopy (FETEM) and Xray diffraction (XRD) analyses confirmed the formation of selenium sulfide (SeS), achavalite (FeSe) and pyrite (FeS2 ), which were the possible removal mechanisms of selenium and iron. Thus, this study exhibited that se lenium, iron, and nitrate can be simultaneously removed to meet the drinking water standards in a sulfidogenic bioreactor.
1. Introduction Selenium (Se) is a non-metal and has been reported as a metalloid by many researchers (Fordyce, 2007; Tan et al., 2016). Selenium occurs in water as selenate (SeO24 ), selenite (SeO23 ), insoluble elemental sele nium (Se0) and selenide (Se2 ) (Vesper et al., 2008). Selenium is required as a nutrient in the range of 30–70 μg Se/d (Lenz and Lens, 2009), its deficiency causes “Keshan” disease in humans (Fordyce, 2007). Consumption of high selenium-containing water leads to chronic or acute selenosis (Dhillon and Dhillon, 2003), cancer, renal, neuro logical and dermatological problems (Sun et al., 2014). Due to toxic nature of Se, different permissible limits of 10, 40 and 50 μg/L, are set by BIS, WHO and U.S. EPA, respectively, in drinking water (BIS:10500, 2012; USEPA, 2018; WHO, 2011). Moreover, some countries have imposed even stringer limits of 1 μg/L (Russia) and 8 μg/L (Germany) (Dhillon and Dhillon, 2016). Sulfuric acid production, metal refinery, coal combustion, electronic industry, mining activities and usage of fertilizer contributes Se in water (Nancharaiah and Lens, 2015; Pyr � ska, 2002). In groundwater, selenium occurrence of 341–1100 μg/L zyn in India (Bajaj et al., 2011; Raghunath et al., 2002), 1000 μg/L in Finland (Dhillon and Dhillon, 2016) and 4200 μg/L in California
(McNeal and Balistrieri, 1989) have been reported. Furthermore, co-occurrence of Se and Fe in groundwater of India (Se ¼ 175 μg/L, Fe ¼ 6.46 mg/L) and northern united states (Se ¼ 223 μg/L and Fe ¼ 3.81 mg/L) have been reported (Bhalla et al., 2011; USGS, 2009). Moreover, some reports specified the co-occurrence of NO3 (~80–247 mg/L) along with Se (~140–740 μg/L) and Fe (~5 mg/L) (Al Kuisi and Abdel-Fattah, 2010; Bhalla et al., 2011; CGWB, 2014). Reverse osmosis, ion-exchange, and adsorption are some of the common practices for Se removal alone (Tan et al., 2016). Due to several limitations of these methodologies on simultaneous removal of multiple contaminants (Shakya et al., 2019), biological processes have gained significant attention in recent past on not only concurrent removal of co-contaminants, but also generation of low and stable sludge as waste (Shakya et al., 2018; Shakya and Ghosh, 2019a, 2019b). Recently, the efficiency of sulfidogenic bioprocess on simultaneous removal of arse nate, nitrate, and iron from groundwater has been reported (Shakya and Ghosh, 2019a, 2019b). However, simultaneous removal of Se, Fe, and NO3 from contaminated groundwater has not yet been reported so far. The objective of this investigation is to survey the possibility and applicability of sulfidogenic bioprocess on the simultaneous selenium, iron, and nitrate removal from polluted groundwater to meet the
* Corresponding author. Department of Civil Engineering, IIT Guwahati, Guwahati, India. E-mail addresses: [email protected], [email protected] (A.K. Shakya). https://doi.org/10.1016/j.jenvman.2019.109757 Received 6 February 2019; Received in revised form 7 October 2019; Accepted 21 October 2019 Available online 14 November 2019 0301-4797/© 2019 Elsevier Ltd. All rights reserved.
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Journal of Environmental Management 254 (2020) 109757
permissible limits for drinking purposes.
(II) solution was filtered through 0.2 μm filter papers and supplied to the bioreactor with the help of a syringe pump (SP 10, Miclins, India) to avoid its pre-precipitation through reactions with the other feed com ponents. Two nos. of annular screens (1 mm diameter openings, 70% perforation) were provided in the AGBR, one at the top to avoid the expulsion of WAC while backwashing, and another at the bottom to avoid loss of WAC granules with the effluent. Additionally, the top screen helps in better distribution of feed over the entire cross-section of the bioreactor.
2. Materials and methods 2.1. Sources of biomass and supporting material used in the bioreactor The anaerobic sludge, which was collected from the anoxic zone of IIT Guwahati sewage treatment plant, was acclimatized in an anoxic environment to grow sulfate-reducing bacteria (SRB). The sewage treatment plant is a facultative lagoon type constructed for the treatment of domestic wastewater generated in the hostels and employee quarters of IIT Guwahati campus. The bioreactor was inoculated with the accli matized bacterial consortium. Waste activated carbon (WAC) granules were utilized as the supporting material for growing the biofilm in the bioreactor. Exhausted WAC was collected from domestic water purifiers installed at different locations of IIT Guwahati. After the collection, it was rinsed twice with the Millipore™ water and oven-dried at 70 � C for 24 h, before its packing in lab-scale attached growth bioreactor (AGBR). WAC was chosen for utilization of a waste material as a resource based on its applicability as packing material (Shakya and Ghosh, 2018a), and its ability in the removal of contaminants (after suitable pre-treatments) such as atrazine (Ghosh et al., 2005), and hexavalent chromium (Ghosh, 2009). Prior to use as the packing material in AGBR, batch adsorption study was conducted with WAC (2 g/L) at 30 � C, pH~7 and 150 rpm to assess its selenium adsorption capacity (at 1000 μg/L of initial selenium).
2.3. Synthetic groundwater Throughout the experiments Se, Fe and NO3 amended synthetic groundwater was used as feed for the AGBR containing (in mg/L)Na2SO4 (as SO24 ): 25–75, NaNO3 (as NO3 ): 50, Na2SeO4 (as Se): 0.2–1.0, FeCl2.4H2O (as Fe): 3–10, CH3COOH (as C): 35–48, NaHCO3 (as HCO3 ): 45–50, NaCl (as Cl ): 13, CaCl2 (as Cl ): 13, MgCl2.6H2O (as Cl ): 13, K2CO3 (as CO23 ): 1.5 and H3PO4 (as P): 0.5 (Shakya and Ghosh, 2018a). Synthetic groundwater was prepared every day and flushed with nitrogen gas for 30 min to lower down the dissolved oxygen to below 1 mg/L. Amount of acetic acid was supplied to meet the requirement of carbon necessity for removing all electron acceptors (residual DO, nitrate, sulfate, and selenate), the mean net yield of 0.4 g biomass/g acetic acid as COD plus a surplus of 1.5. 2.4. AGBR operating conditions The acclimatized microbial consortium (0.3 L) of 3570 mg/L as MLSS (MLVSS ¼ 2480 mg/L) was inoculated to the AGBR. AGBR was wrapped (full depth) with silicone tubes, through which warm water (~30 � C) was circulated from a thermostat water-bath (RA-8, LAUDA, Germany). To maintain the EBCT of 45, 60, 90, and 120 min, the flow rate was kept at 7.4, 5.6, 3.7, 2.8 ml/min, respectively using a peristaltic pump. Backwashing of the AGBR bed was carried out routinely (10 days in terval) using deoxygenated-deionized (DI) water for around 3–4 min commixed with nitrogen gas to achieve complete fluidization of WAC bed (Shakya and Ghosh, 2019a, 2019b). The effluent samples were taken at 24 h interval from the effluent port and intermittently from all sampling ports, whenever port profile sampling was planned. Operating schedule and details of port profile sampling of the AGBR is shown in
2.2. AGBR setup A lab-scale AGBR was fabricated using a plexiglass column having an inner diameter of 5 cm, and height of 32 cm, which was filled up to 17 cm with 210 g of WAC granules (size ~ 1.5–2 mm). The AGBR setup is shown in Fig. 1. The empty bed contact time (EBCT) were measured based on the total bed volume of 333 cm3 (height of 17 cm). A total of seven ports were designed along the depth of the bioreactor. The top and bottom most ports were designated as the inlet and effluent ports respectively, and the remaining five ports (namely, S1 to S5 with c/c spacing of 3 cm) were used for profile sampling. Two nos. peristaltic pumps (PP 10 and PP 20EX, Miclins, India) were used for regulating the flow of synthetic groundwater and backwash water in the bioreactor. Fe
Fig. 1. Schematic diagram of lab scale set-up of AGBR. 2
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Journal of Environmental Management 254 (2020) 109757
Table 1. Initially, the AGBR was fed with 25 mg/L of sulfate, 50 mg/L of nitrate, 105 mg/L of COD and run at the EBCT of 120 min. Fe (3 mg/L) and Se (200 μg/L) were introduced in the system on the 84th day, after attaining a steady-state with respect to sulfate removal. Once the start-up period was over, the effects of initial Fe (3, 4, 5, 7, and 10 mg/L) were studied. From 264th day onwards, the effects of Se (300, 500, 750 and 1000 μg/L) on the AGBR performance were analyzed. During this experimental period, the AGBR was fed with 10 mg/L of Fe and 75 mg/L of sulfate at the EBCT of 60 min. To optimize Fe and Se removal, sulfate concentration was also varied between 25 and 75 mg/L. To fulfill the requirement of additional carbon at an enhanced influent sulfate (50 and 75 mg/L), influent COD was accordingly increased in the feed.
outfitted with a monochromator. X-ray diffraction (XRD) patterns were recorded for a freeze-dried sample from 5� to 85� in the 2θ range using Cu Kα radiation (λ ¼ 1.54 Å) at 40 mA and 40 kV with 1-sec step duration and 0.05� step size. The XRD results were concluded using the Joint Committee on Powder Diffraction and Standards (JCPDS) from the X’Pert Highscore (Match version 3) software. Samples were further analyzed on field emission transmission elec tron microscopy (FETEM) (JEOL JEM-2100F) to confirm the existence of compound obtained from XRD analysis. Samples for FETEM were pul verized, dispersed in acetone, and then subjected to sonication for 30 min (Kirk et al., 2010). The obtained solution was then diluted and drop-casted on the carbon-coated copper (TEM) grid using a Pasteur pipette. The grid was left to dry in a desiccator for overnight and then used for FETEM analysis.
2.5. Analytical approach
3. Results and discussion
Sulfate, COD, and NO3 were measured by nephelo turbidity meter (4500 SO24 - E), closed refluxed method (5220-C), and UV–Visible spectrophotometer (4500 NO3 -B), respectively (APHA, 2005). Before being analyzed, total selenium and iron samples were acidified with 2 ml 1N HCl and was made up to a final volume of 100 ml to dissolve any precipitates. The acidified samples were finally preserved at 4 � C until further analysis. The total selenium was analyzed (λ ¼ 196 nm) in atomic absorption spectrophotometer (SpectrAA 55B, Varian) assem bled with vapor generation assembly (VGA- 77). For the evaluation of Fe, phenanthroline method (3500-Fe B) and/or AAS (flame mode) was used.
3.1. Start-up phase and effects of initial iron on the performance of AGBR Initially, the AGBR was run for the start-up phase to get fair colo nization of mixed microbial consortium on WAC granules. Although selenium adsorption on activated carbon has been reported (Dhillon and Dhillon, 2016), poor selenium adsorption (less than 10% in 24 h) was observed on WAC which inferred biogenic sulphides played the primary role in selenium removal. Complete NO3 removal was observed from the 1st day of AGBR run, which is in line with a study by Shakya and Ghosh (2019a, 2019b). Sulfate reduction was poor during the first few days (day 0–25) of start-up and only 4–14 mg/L of it was removed across the AGBR. This could be attributed to the prolonged adaptation period required for SRB in biofilm, comprising of a wide array of active bac terial genera (Frunzo et al., 2012). Sulfate reduction was improved slowly, and only 10–12 mg/L of it was left out in the effluent during the last few days (day 61–83) of start-up phase (Fig. A1). With the short ening of EBCT, sulfate reduction was hampered only for first few days of operation. (Shakya and Ghosh, 2018a). Similarly, COD in the effluent was also stabilized slowly, and only 12–15 mg/L remained during the last few days (day 61–83). Once steady state in terms of sulfate removal was attained, EBCT was reduced stepwise to 90 min and 60 min on 61st and 74th day, respectively. After attaining stabilized sulfate reduction and EBCT optimization, Se (200 μg/L) and Fe (3 mg/L) were introduced into the AGBR on 84th day. The effects of varying initial Fe on the performance of AGBR is shown
2.6. Mineralogical analysis The collection of biosolids from the backwash suspension was done in a nitrogen gas-filled container. The collected backwash suspension was centrifuged for 10 min at 5000 rpm to get a semi-liquid paste, which was freeze-dried and kept in sealed Falcon™ tubes at 4 � C for preser vation. To limit the contact between ambient air and biosolids, tubes were stored in an airtight double zip lock pouches and opened only in anoxic environment. The elemental confirmation in the backwash bio solids was carried out by field emission scanning electron microscopy (FESEM) (Sigma, Zeiss, Germany) and energy dispersive X-ray (EDX) (Oxford, INCA 300, UK). Before analysis, the freeze-dried samples were spread on the FESEM stub (stuck with carbon tape) and then gold coated. To check the presence of any crystalline phase, biosolids were characterized using the PANalytical X-pert PRO MPD diffractometer Table 1 Operating schedule of AGBR. Purpose
Start-up phase Effect of initial iron
Effect of initial selenium
Operating Days
Concentration in Influent Feed Selenium (μg/ L)
Iron (mg/ L)
Sulfate (mg/ L)
Nitrate (mg/ L)
COD (mg/ L)
1–60 61–73 74–83 84–103 104–123 124–142 143–172 173–180
– – – 200 200 200 200 300
– – – 3 4 4 5 5
25 25 25 25 25 25 25 25
50 50 50 50 50 50 50 50
105 105 105 105 105 105 105 105
120 90 60 60 60 45 45 45
181–198
300
5
25
50
105
60
199–227
300
5
50
50
121
60
228–248 249–263 264–272 264–272 273–292 293–312 313–332
300 300 300 300 500 750 1000
7 7 10 10 10 10 10
50 75 75 75 75 75 75
50 50 50 50 50 50 50
121 134 134 134 134 134 134
60 60 60 60 60 60 60
3
EBCT (minutes)
Remarks
Profile sampling on day 180 Profile sampling on day 181 Profile sampling on day 227
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in Fig. 2. Iron removal (initial 4 mg/L) met the permissible limit in drinking water in presence of sulfate (initial 25 mg/L) and 45 min EBCT. Iron between 0.42 and 0.91 mg/L was observed in the effluent from day 143–198 (initial Fe of 5 mg/L). As Fe removal in sulfidogenic systems is mainly due to the formation of biogenic iron sulfides, the plausible explanation behind elevated Fe in the effluent could be due to insuffi cient amount of sulfides that is necessary for the precipitation as iron sulfides (Shakya and Ghosh, 2019a, 2019b). As further sulfate reduction was not anticipated in AGBR, EBCT was extended to 60 min (from 45 min) on 181th day to facilitate longer reaction and contact time be tween biogenic sulphides and iron (for iron sulfides precipitation). This resulted in better sulfate removal and eventually improved Fe removal. However, Fe (0.42 mg/L) in the effluent was still higher than the BIS permissible limit. Hypothesizing insufficient sulfides for iron sulfides precipitation, influent sulfate was raised to 50 mg/L (on 199th day), which boosted Fe removal to well below the permissible limit. At initial Fe of 7 mg/L (day 228–248), 1.5 mg/L of Fe appeared once more in the effluent, necessitating more sulphides requirement. Thus initial sulfate was further raised to 75 mg/L on day 249. On increasing initial sulfate, Fe in effluent was reduced below the permissible limit, with around 11–15 mg/L of sulfate remaining. Higher EBCT and sulfate are docu mented to ameliorate a bioreactor performance (Shakya and Ghosh, 2018a), whereas lower EBCT may result in poor sulfate removal and/or limited contact time for biogenic sulfides formation, which in turn left high iron in the effluent (Upadhyaya et al., 2010). Further increase in initial iron to 10 mg/L (day 264–272) did not hampered iron removal and effluent iron remained below drinking standards during subsequent AGBR run. Effluent Se remained well below the BIS permissible limit for most of the time. During the entire phase, effluent pH was always in the range of 7.2–7.6, which was slightly more than the raw water. Increase in pH could be attributed to the formation of alkalinity during oxyanions (NO3 and sulfate) reduction (Brahmacharimayum and Ghosh, 2014). Effluent COD was in the range of 10–13 mg/L except for few data, which were comparatively better as described by Shakya and Ghosh (2018a). Higher COD of 22 and 25 mg/L was observed on day 199 and 249, respectively and was possibly due to extra carbon added to meet carbon
requirement for elevated sulfate levels (Shakya and Ghosh, 2018a). 3.2. Effects of selenium concentration on the performance of AGBR The effects of initial Se concentrations of 300, 500, 750, and 1000 μg/L on the performance of AGBR are shown in Fig. 3. During this experimental period, AGBR was operated at initial Fe, NO3 , sulfate, and COD of 10 mg/L, 50 mg/L, 75 mg/L, and 134 mg/L, respectively at 60 min EBCT. The above-mentioned feeding and operating conditions were chosen as per the effective performance of AGBR at 10 mg/L of initial Fe concentration. Despite increment in the initial Se concentra tions, Se and Fe in the effluent remained well below the BIS permissible limit. Furthermore, with the increase in influent Se, increment in Fe removal was observed (day 273–332), which could be attributed to the formation of achavalite (Fe–Se) in the AGBR (Gerhardt et al., 1991). During the last few days (day 313–332) of this study, both Fe and Se removal were found to be around 99%, whereas NO3 was completely removed from the bioreactor. The plausible cause for better selenium and iron removal could be due to availability of sufficient sulphides and/or reaction time for precipitation of selenium and/or iron as their sulphides. In the AGBR, selenium was likely eliminated from the liquid phase through direct selenium sulfide precipitation and/or surface precipitation on iron sulfide solids (Mitchell et al., 2013). Adsorption on iron sulfides might have provided additional selenium removal. Previ ous reports also suggested iron compounds as good adsorbents of sele nium (Dhillon and Dhillon, 2016). However, an overdose of nitrate and sulfate should be avoided, that would otherwise raise carbonate alka linity (Altun et al., 2014) and sulphide levels (Luo et al., 2008). Although precipitation of metal sulfides is favored at high alkalinity (Hall, 1986), it can dissolute precipitates of metal sulfides (Henke, 2009) and may cause an increase in selenium and iron in the effluent. Effluent sulfate and COD were observed to be around 6.5 and 11.5 mg/L, respectively during this phase. Effluent pH remained between 7.40 and 7.65, which was in the range of permissible limit for drinking water. These results showed that the Se increment of up to 1000 μg/L did not alter the performance of the AGBR.
Fig. 2. Effects of initial iron concentration on the performance of AGBR. 4
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Fig. 3. Effects of initial selenium concentration on the performance of AGBR.
3.3. Port profile analysis
standards.
To illustrate the variations in the concentration of contaminants along the AGBR bed depth, port profile sampling were performed on 180th, 181th and 227th day of the reactor operation (Fig. 4). It also depicted the effects of EBCT and sulfate concentrations on Fe and Se removal efficiencies. The samples were taken anaerobically from each sampling port (S1 to S5 and effluent) and analyzed for Se, Fe, NO3 , sulfate, and COD. As shown in Fig. 4, complete removal of NO3 (50 mg/ L) was observed in port S1, and sulfate removal occurs after that. This is mainly because sulfate is thermodynamically an inferior electron acceptor than NO3 . Similar results were also reported by Shakya and Ghosh (2019a, 2019b) during simultaneous removal of nitrate and sul fate using mixed microbial consortium. It was observed (Fig. 4) that Fe and Se removal followed sulfate removal in AGBR. Although some fluctuations in Fe concentration was observed, Se in the effluent was always below the permissible limit until it reached the port S5. The results of port profile sampling performed on day 180 and 181 (45 min EBCT) are shown in Fig. 4 (a, b). As observed in Fig. 4 (a) sulfate removal across the system was around 18 mg/L, Fe removal followed sulfate removal, and 0.85 mg/L of Fe was left out in port S5. Most of the sulfate was removed in port S4, and no further sulfate reduction was observed. On day 181, no considerable improvement in effluent sulfate and Fe removal was observed (Fig. 4 (b)) when EBCT was increased to 60 min. This implies that poor Fe removal in the system was mainly not associated with insufficient contact time. Considering the insufficient sulfide availability the cause, it was decided to increase the influent sulfate keeping the EBCT constant at 60 min. Fig. 4 (c) (day 227th) indicated that around 37.5 mg/L (till port S4), sulfate was removed across the system, which in turn enhanced Fe removal. This inferred that increased influent sulfate provided enough sulfides to precipitate as biogenic iron sulphides (Shakya and Ghosh, 2018b). Thus, port profile sampling results illustrated that sufficient contact time, and sulfate dose were necessary for the removal of Fe in the AGBR to meet the drinking
3.4. Overall performance of AGBR The outcome of the experiments conducted on performance assess ment of AGBR on simultaneous removal of Se (200–1000 μg/L) and Fe (3–10 mg/L) depicted that the removal of both the pollutants met their permissible limits in drinking water, along with complete removal of NO3 (50 mg/L). An increment in effluent pH was observed but was al ways within the range of drinking water standards. Selenium removal from a sulfidogenic bioreactor occured either through selenium sulfide precipitation (Jung et al., 2016) or adsorption by iron sulfides and/or co-precipitation (Charlet et al., 2012; Diener et al., 2012). Biogenic sulfides formed due to the reduction of sulfate in the bioreactor com bined with the reduced selenium and/or iron to precipitate as their respective sulfides. Although no negative effect on Se removal was observed, Fe in the effluent was found to be more than the permissible limit (at initial Fe of 5 and 7 mg/L) during some experimental condi tions. As an outcome of the sulfidogenic activity, iron removal in the present system took place mostly through the formation of iron sulfides. Therefore, the appearance of Fe (more than 0.3 mg/L) in the effluent could be due to the inadequate amount of sulfides and/or short reaction time (Shakya and Ghosh, 2018b), which was overcome by providing more amount of sulfate and/or longer EBCT. Increasing EBCT alone to 60 min (from 45 min) improved Fe removal (Fig. 2; day 181–198) but could not meet the drinking water standards (as effluent Fe was 0.5 mg/L). However, increase in influent sulfate (25–50 mg/L) along with the increase in EBCT (45 min–60 min) helps in reducing iron in the effluent to below 0.3 mg/L (Fig. 2; day 199–227). Similarly, sulfate dose of 75 mg/L was required for initial Fe or 7 mg/L (Fig. 2; day 249–263). Thus, it can be inferred that proper sulfate dose and sufficient contact time are important parameters to get desirable performance.
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Fig. 4. Port profile sampling of AGBR on day 180 (a) and day 181 (b), showing the effects of EBCT and sampling on day 181 (b) and day 227 (c), showing the effects of sulfate concentrations on selenium and iron removal efficiencies.
3.5. Biogenic precipitate characterization
and iron compared to selenium are probably due to the higher sulfate and iron in the feed, which led to the formation of more iron sulfides in the AGBR. FESEM/EDX results were further validated by XRD analysis. XRD analysis depicted the existence of selenium sulfide (SeS; JCPDS #98-002-1203), achavalite (FeSe; JCPDS #98-011-6498) and pyrite (FeS2 ; JCPDS #98-004-9014) in the biosolids (Fig. 5b). The FESEM and XRD analyses confirmed that the precipitation as iron persulfide (FeS2 )
FESEM micrograph with EDX mapping (Fig. 5a) showed uniformly distributed sulfur, iron and selenium in the biosolids. Presence of rela tively high amounts of sulfur {Fig. A2 (a)} and iron {Fig. A2 (b)} in comparison to selenium (Fig. A2 (c)) were observed in the elemental mapping of an individual element. Presence of higher amounts of sulfur 6
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Fig. 5. (a) FESEM/EDX elemental mapping and (b) XRD pattern of biosolids formed in AGBR.
and selenium sulfide (SeS) were the removal mechanism of Fe and Se in the bioreactor (Jung et al., 2016). Furthermore, Selenium removal was also ascribed to adsorption on iron sulfides (FeS/FeS2 ) and/or co-precipitation as iron selenide (FeSe on FeS/FeS2 ) (Charlet et al., 2012; Diener et al., 2012). Therefore, in the present study, Fe addition enhanced Se removal. The peak intensities showed that the compound was of semi-crystalline in nature, which was further supported by TEM analysis. Fig. 6 showed FETEM images {Fig. 6 (a,b)}, SAED pattern {Fig. 6 (c)}, high resolution transmission electron microscopy (HRTEM) image (Fig. 6 (d)) and corresponding Inverse Fast Fourier Transform (IFFT) image (Fig. 6, e) of the collected biosolids. ImageJ and Gatan Digital Micrograph software (version 3.4) were used to determine the d-spacing from selected area electron diffraction (SAED) pattern and HRTEM image, respectively. The belt-structure showed distinctive clear
lattice fringes with a d-spacing of 0.27 nm {Fig. 6 (e)} which corre sponds to the compound FeS2. The d-spacing values obtained from XRD analysis (sharp peak at 33� 2θ value) were also in line with the TEM analysis. The SAED pattern indicated the precipitated compound exhibiting both crystallinity, and amorphicity in nature, which was also observed in the XRD pattern. Crystalline nature of the obtained pre cipitates suggests the stable nature of biosolids (de Matos et al., 2018), which ensures their safe disposal in a landfill. Moreover, the reduced biosolids from the anaerobic bioreactor offers three-fold benefits: (1) low volume of biosludge generation; (2) stable in anaerobic environ ments, and (3) stable after exposure to air (first aerobic environment) due to formation of oxides of iron, which are also potential adsorbents of Se (Mitchell et al., 2013).
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Fig. 6. (a, b) TEM images, (c) SAED pattern, (d) HRTEM image and (e) IFFT of the backwash biosolids collected from AGBR.
4. Conclusions
References
This study demonstrated the simultaneous removal of Se (200–1000 μg/L) and Fe (3–10 mg/L) to meet drinking water standards. Nitrate (50 mg/L) was completely removed from synthetic groundwater. At initial Fe of 5 and 7 mg/L, more than 0.3 mg/L was left out in the effluent due to an inadequate amount of sulfides or insufficient contact time. Addition of more sulfate (50 and 75 mg/L) in the synthetic water and/or increase in EBCT (60 min) helped in reducing effluent iron concentration. Due to the selection of suitable EBCT and proper assessment of carbon dose, soluble COD in the effluent was only around 11 mg/L. The characterization of biosoilds confirmed the bioprecipitation as SeS, FeSe, and FeS2 were the removal mechanisms of Se and Fe. FETEM analysis suggested the formation of stable bio-sludge in the bioreactor. Being low volume (anaerobic/anoxic process) and stable nature, sludge-handling problem can be expected to be low. This outperforms the physicochemical and aerobic processes of Se removal in view of practical utility. Overall, sulfidogenic bioreactor system can simultaneously remove selenium, iron, and nitrate from polluted groundwater to meet drinking water guidelines at an optimum dose of sulfate and sufficient EBCT.
Al Kuisi, M., Abdel-Fattah, A., 2010. Groundwater vulnerability to selenium in semi-arid environments: amman Zarqa Basin, Jordan. Environ. Geochem. Health 32 (2), 107–128. Altun, M., Sahinkaya, E., Durukan, I., Bektas, S., Komnitsas, K., 2014. Arsenic removal in a sulfidogenic fixed-bed column bioreactor. J. Hazard Mater. 269, 31–37. APHA, 2005. Standard Methods for the Examination of Water & Wastewater, twenty-first ed. American Public Health Association, Washington DC. Bajaj, M., Eiche, E., Neumann, T., Winter, J., Gallert, C., 2011. Hazardous concentrations of selenium in soil and groundwater in North-West India. J. Hazard Mater. 189 (3), 640–646. Bhalla, A., Singh, G., Kumar, S., Shahi, J., Mehta, D., 2011. Elemental analysis of ground water from different regions of Punjab state (India) using EDXRF technique and the sources of water contamination. Int Conf Environ Comput Sci 156–164. BIS:10500, 2012. Indian Standard: Drinking Water - Specification. Bureau of Indian Standard, New Delhi, India (Second Revision). Brahmacharimayum, B., Ghosh, P.K., 2014. Sulfate bioreduction and elemental sulfur formation in a packed bed reactor. J. Environ. Chem. Eng. 2 (3), 1287–1293. CGWB, P., 2014. Water Quality Issues and Challenges in Punjab, Central Ground Water Board. Ministry of Water Resources, Government of India. March-2014. Charlet, L., Kang, M., Bardelli, F., Kirsch, R., G�ehin, A., Gren� eche, J.-M., Chen, F., 2012. Nanocomposite pyrite–greigite reactivity toward Se (IV)/Se (VI). Environ. Sci. Technol. 46 (9), 4869–4876. de Matos, L.P., Costa, P.F., Moreira, M., Gomes, P.C.S., de Queiroz Silva, S., Gurgel, L.V. A., Teixeira, M.C., 2018. Simultaneous removal of sulfate and arsenic using immobilized non-traditional SRB mixed culture and alternative low-cost carbon sources. Chem. Eng. J. 334, 1630–1641. Dhillon, K.S., Dhillon, S.K., 2003. Quality of underground water and its contribution towards selenium enrichment of the soil-plant system for a seleniferous region of northwest India. J. Hydrol. 272, 120–130. Dhillon, K.S., Dhillon, S.K., 2016. Selenium in groundwater and its contribution towards daily dietary Se intake under different hydrogeological zones of Punjab, India. J. Hydrol. 533, 615–626. Diener, A., Neumann, T., Kramar, U., Schild, D., 2012. Structure of selenium incorporated in pyrite and mackinawite as determined by XAFS analyses. J. Contam. Hydrol. 133, 30–39. Fordyce, F., 2007. Selenium geochemistry and health. AMBIO A J. Hum. Environ. 36 (1), 94–97. Frunzo, L., Esposito, G., Pirozzi, F., Lens, P., 2012. Dynamic mathematical modeling of sulfate reducing gas-lift reactors. Process Biochem. 47 (12), 2172–2181. Gerhardt, M., Green, F., Newman, D., Lundquist, R.J., Blake Tresan, R., Oswald, J.W, 1991. Removal of Selenium Using a Novel Algal-Bacterial Process, pp. 799–805. Ghosh, P.K., 2009. Hexavalent chromium [Cr (VI)] removal by acid modified waste activated carbons. J. Hazard Mater. 171 (1–3), 116–122. Ghosh, P.K., Philip, L., Bandyopadhyay, M., 2005. Management of Atrazine bearing wastewater using an upflow anaerobic sludge blanket reactor–adsorption system. Pract. Period. Hazard. Toxic, Radioact. Waste Manag. 9 (2), 112–121. Hall, A., 1986. Pyrite-pyrrhotine redox reactions in nature. Mineral. Mag. 50 (6), 223–229. Henke, K., 2009. Arsenic: Environmental Chemistry, Health Threats and Waste Treatment. John Wiley & Sons, Chichester, UK. https://doi.org/10.1002/ 9780470741122.ch2. Ltd.
Declaration of interest statement The authors declare that we have no conflict of interest. The help in any form received from any party have been acknowledged. Acknowledgments The authors would like to appreciate the assistance from the Central Instrument Facility, Department of Civil Engineering, Physics and Bio sciences and Bioengineering of IIT Guwahati, for giving different instrumental services, manpower and so forth are very recognized. Au thors would also acknowledge Mr. Niranjan Kumar, Mr. Siddhartha Paul, and Mr. Gajjela Rajashekhar, IIT Guwahati, for their help in various required analysis during this project. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.jenvman.2019.109757. 8
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Journal of Environmental Management 254 (2020) 109757 Shakya, A.K., Ghosh, P.K., 2019. Concurrent removal of nitrate, arsenic and ironfrom simulated and real-life groundwater to meet drinking water standards: effects ofoperational and environmental parameters. J. Environ. Manag. 9–18. Shakya, A.K., Ghosh, P.K., 2019. Effects of backwashing strategy and dissolved oxygen on arsenic removal to meet drinking water standards in a sulfidogenic attached growth reactor. J. Hazard Mater. 369, 309–317. Shakya, A.K., Paul, S., Ghosh, P.K., 2019. Bio-attenuation of arsenic and iron coupledwith nitrate remediation in multi-oxyanionic system: batch and column studies. J. Hazard Mater. 375, 182–190. Sun, H.-J., Rathinasabapathi, B., Wu, B., Luo, J., Pu, L.-P., Ma, L.Q., 2014. Arsenic and selenium toxicity and their interactive effects in humans. Environ. Int. 69, 148–158. Tan, L.C., Nancharaiah, Y.V., van Hullebusch, E.D., Lens, P.N.L., 2016. Selenium: environmental significance, pollution, and biological treatment technologies. Biotechnol. Adv. 34 (5), 886–907. Upadhyaya, G., Jackson, J., Clancy, T.M., Hyun, S.P., Brown, J., Hayes, K.F., Raskin, L., 2010. Simultaneous removal of nitrate and arsenic from drinking water sources utilizing a fixed-bed bioreactor system. Water Res. 44 (17), 4958–4969. USEPA, 2018. United States environmental protection agency. “National Primary Drinking Water Regulations” URL: https://www.epa.gov/ground-wateranddrinking-water/national-primary-drinking-water-regulations#Inorganic. (Accessed 17 July 2018). USGS, 2009. Occurrence and Distribution of Iron, Manganese, and Selected Trace Elements in Ground Water in the Glacial Aquifer System of the Northern United States. Scientific Investigations Report, pp. 2009–5006. Vesper, D.J., Roy, M., Rhoads, C.J., 2008. Selenium distribution and mode of occurrence in the Kanawha Formation, southern West Virginia, USA. Int. J. Coal Geol. 73 (3–4), 237–249. WHO, 2011. Guidelines for Drinking-Water Quality World Health Organisation, vol. 4. Geneva, 315e318.
Jung, B., Safan, A., Batchelor, B., Abdel-Wahab, A., 2016. Spectroscopic study of Se(IV) removal from water by reductive precipitation using sulfide. Chemosphere 163, 351–358. Kirk, M.F., Roden, E.E., Crossey, L.J., Brealey, A.J., Spilde, M.N., 2010. Experimental analysis of arsenic precipitation during microbial sulfate and iron reduction in model aquifer sediment reactors. Geochem. Cosmochim. Acta 74, 2538–2555. Lenz, M., Lens, P.N., 2009. The essential toxin: the changing perception of selenium in environmental sciences. Sci. Total Environ. 407 (12), 3620–3633. Luo, Q., Tsukamoto, T.K., Zamzow, K.L., Miller, G.C., 2008. Arsenic, Selenium, and Sulfate Removal Using an Ethanol-Enhanced Sulfate-Reducing Bioreactor, vol. 27. Springer, pp. 100–108. McNeal, J.M., Balistrieri, L.S., 1989. Geochemistry and occurrence of selenium: an overview. Selenium Agric. Environ.(seleniuminagric) 1–13. Mitchell, K., Couture, R.-M., Johnson, T.M., Mason, P.R., Van Cappellen, P., 2013. Selenium sorption and isotope fractionation: iron (III) oxides versus iron (II) sulfides. Chem. Geol. 342, 21–28. Nancharaiah, Y.V., Lens, P.N., 2015. Selenium biomineralization for biotechnological applications. Trends Biotechnol. 33 (6), 323–330. Pyrzy� nska, K., 2002. Determination of selenium species in environmental samples. Microchimica Acta 140 (1–2), 55–62. Raghunath, R., Tripathi, R., Mahapatra, S., Sadasivan, S., 2002. Selenium levels in biological matrices in adult population of Mumbai, India. Sci. Total Environ. 285 (1–3), 21–27. Shakya, A.K., Ghosh, P.K., 2018. Simultaneous removal of arsenic and nitrate in absence of iron in an attached growth bioreactor to meet drinking water standards: importance of sulphate and empty bed contact time. J. Clean. Prod. 186, 304–312. Shakya, A.K., Ghosh, P.K., 2018. Simultaneous removal of arsenic, iron and nitrate in an attached growth bioreactor to meet drinking water standards: importance of sulphate and empty bed contact time. J. Clean. Prod. 186, 1011–1020. Shakya, A.K., Rajput, P., Ghosh, P.K., 2018. Investigation on stability and leachingcharacteristics of mixtures of biogenic arsenosulphides and iron sulphides formed under reduced conditions. J. Hazard Mater. 353, 320–328.
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Research article
Development of a sulfidogenic bioreactor system for removal of co-existent selenium, iron and nitrate from drinking water sources Himanshu Sonkeshariya a, Arvind Kumar Shakya a, b, *, Pranab Kumar Ghosh a a b
Department of Civil Engineering, IIT Guwahati, Guwahati, India Department of Biotechnology, Rama University Uttar Pradesh, Kanpur, India
A R T I C L E I N F O
A B S T R A C T
Keywords: Selenium Iron Nitrate Selenium sulfide Achavalite Drinking water
The present study showed for the first time that selenium, iron, and nitrate could be simultaneously removed in a sulfidogenic bioreactor to meet drinking water standards. A bioreactor inoculated with mixed bacterial con sortium was operated for around 330 days in anoxic environment at 30 � C under varying combination of influent selenate (200–1000 μg/L as selenium), and iron (3–10 mg/L) in presence of 50 mg/L of nitrate. Required amount of acetic acid (as carbon source) and sulfate were supplied and the reactor was operated at different empty bed contact time (EBCT) of 45–120 min. Along with complete removal of nitrate, the reactor removed both selenium and iron to meet the drinking water standards. Field emission transmission electron microscopy (FETEM) and Xray diffraction (XRD) analyses confirmed the formation of selenium sulfide (SeS), achavalite (FeSe) and pyrite (FeS2 ), which were the possible removal mechanisms of selenium and iron. Thus, this study exhibited that se lenium, iron, and nitrate can be simultaneously removed to meet the drinking water standards in a sulfidogenic bioreactor.
1. Introduction Selenium (Se) is a non-metal and has been reported as a metalloid by many researchers (Fordyce, 2007; Tan et al., 2016). Selenium occurs in water as selenate (SeO24 ), selenite (SeO23 ), insoluble elemental sele nium (Se0) and selenide (Se2 ) (Vesper et al., 2008). Selenium is required as a nutrient in the range of 30–70 μg Se/d (Lenz and Lens, 2009), its deficiency causes “Keshan” disease in humans (Fordyce, 2007). Consumption of high selenium-containing water leads to chronic or acute selenosis (Dhillon and Dhillon, 2003), cancer, renal, neuro logical and dermatological problems (Sun et al., 2014). Due to toxic nature of Se, different permissible limits of 10, 40 and 50 μg/L, are set by BIS, WHO and U.S. EPA, respectively, in drinking water (BIS:10500, 2012; USEPA, 2018; WHO, 2011). Moreover, some countries have imposed even stringer limits of 1 μg/L (Russia) and 8 μg/L (Germany) (Dhillon and Dhillon, 2016). Sulfuric acid production, metal refinery, coal combustion, electronic industry, mining activities and usage of fertilizer contributes Se in water (Nancharaiah and Lens, 2015; Pyr � ska, 2002). In groundwater, selenium occurrence of 341–1100 μg/L zyn in India (Bajaj et al., 2011; Raghunath et al., 2002), 1000 μg/L in Finland (Dhillon and Dhillon, 2016) and 4200 μg/L in California
(McNeal and Balistrieri, 1989) have been reported. Furthermore, co-occurrence of Se and Fe in groundwater of India (Se ¼ 175 μg/L, Fe ¼ 6.46 mg/L) and northern united states (Se ¼ 223 μg/L and Fe ¼ 3.81 mg/L) have been reported (Bhalla et al., 2011; USGS, 2009). Moreover, some reports specified the co-occurrence of NO3 (~80–247 mg/L) along with Se (~140–740 μg/L) and Fe (~5 mg/L) (Al Kuisi and Abdel-Fattah, 2010; Bhalla et al., 2011; CGWB, 2014). Reverse osmosis, ion-exchange, and adsorption are some of the common practices for Se removal alone (Tan et al., 2016). Due to several limitations of these methodologies on simultaneous removal of multiple contaminants (Shakya et al., 2019), biological processes have gained significant attention in recent past on not only concurrent removal of co-contaminants, but also generation of low and stable sludge as waste (Shakya et al., 2018; Shakya and Ghosh, 2019a, 2019b). Recently, the efficiency of sulfidogenic bioprocess on simultaneous removal of arse nate, nitrate, and iron from groundwater has been reported (Shakya and Ghosh, 2019a, 2019b). However, simultaneous removal of Se, Fe, and NO3 from contaminated groundwater has not yet been reported so far. The objective of this investigation is to survey the possibility and applicability of sulfidogenic bioprocess on the simultaneous selenium, iron, and nitrate removal from polluted groundwater to meet the
* Corresponding author. Department of Civil Engineering, IIT Guwahati, Guwahati, India. E-mail addresses: [email protected], [email protected] (A.K. Shakya). https://doi.org/10.1016/j.jenvman.2019.109757 Received 6 February 2019; Received in revised form 7 October 2019; Accepted 21 October 2019 Available online 14 November 2019 0301-4797/© 2019 Elsevier Ltd. All rights reserved.
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permissible limits for drinking purposes.
(II) solution was filtered through 0.2 μm filter papers and supplied to the bioreactor with the help of a syringe pump (SP 10, Miclins, India) to avoid its pre-precipitation through reactions with the other feed com ponents. Two nos. of annular screens (1 mm diameter openings, 70% perforation) were provided in the AGBR, one at the top to avoid the expulsion of WAC while backwashing, and another at the bottom to avoid loss of WAC granules with the effluent. Additionally, the top screen helps in better distribution of feed over the entire cross-section of the bioreactor.
2. Materials and methods 2.1. Sources of biomass and supporting material used in the bioreactor The anaerobic sludge, which was collected from the anoxic zone of IIT Guwahati sewage treatment plant, was acclimatized in an anoxic environment to grow sulfate-reducing bacteria (SRB). The sewage treatment plant is a facultative lagoon type constructed for the treatment of domestic wastewater generated in the hostels and employee quarters of IIT Guwahati campus. The bioreactor was inoculated with the accli matized bacterial consortium. Waste activated carbon (WAC) granules were utilized as the supporting material for growing the biofilm in the bioreactor. Exhausted WAC was collected from domestic water purifiers installed at different locations of IIT Guwahati. After the collection, it was rinsed twice with the Millipore™ water and oven-dried at 70 � C for 24 h, before its packing in lab-scale attached growth bioreactor (AGBR). WAC was chosen for utilization of a waste material as a resource based on its applicability as packing material (Shakya and Ghosh, 2018a), and its ability in the removal of contaminants (after suitable pre-treatments) such as atrazine (Ghosh et al., 2005), and hexavalent chromium (Ghosh, 2009). Prior to use as the packing material in AGBR, batch adsorption study was conducted with WAC (2 g/L) at 30 � C, pH~7 and 150 rpm to assess its selenium adsorption capacity (at 1000 μg/L of initial selenium).
2.3. Synthetic groundwater Throughout the experiments Se, Fe and NO3 amended synthetic groundwater was used as feed for the AGBR containing (in mg/L)Na2SO4 (as SO24 ): 25–75, NaNO3 (as NO3 ): 50, Na2SeO4 (as Se): 0.2–1.0, FeCl2.4H2O (as Fe): 3–10, CH3COOH (as C): 35–48, NaHCO3 (as HCO3 ): 45–50, NaCl (as Cl ): 13, CaCl2 (as Cl ): 13, MgCl2.6H2O (as Cl ): 13, K2CO3 (as CO23 ): 1.5 and H3PO4 (as P): 0.5 (Shakya and Ghosh, 2018a). Synthetic groundwater was prepared every day and flushed with nitrogen gas for 30 min to lower down the dissolved oxygen to below 1 mg/L. Amount of acetic acid was supplied to meet the requirement of carbon necessity for removing all electron acceptors (residual DO, nitrate, sulfate, and selenate), the mean net yield of 0.4 g biomass/g acetic acid as COD plus a surplus of 1.5. 2.4. AGBR operating conditions The acclimatized microbial consortium (0.3 L) of 3570 mg/L as MLSS (MLVSS ¼ 2480 mg/L) was inoculated to the AGBR. AGBR was wrapped (full depth) with silicone tubes, through which warm water (~30 � C) was circulated from a thermostat water-bath (RA-8, LAUDA, Germany). To maintain the EBCT of 45, 60, 90, and 120 min, the flow rate was kept at 7.4, 5.6, 3.7, 2.8 ml/min, respectively using a peristaltic pump. Backwashing of the AGBR bed was carried out routinely (10 days in terval) using deoxygenated-deionized (DI) water for around 3–4 min commixed with nitrogen gas to achieve complete fluidization of WAC bed (Shakya and Ghosh, 2019a, 2019b). The effluent samples were taken at 24 h interval from the effluent port and intermittently from all sampling ports, whenever port profile sampling was planned. Operating schedule and details of port profile sampling of the AGBR is shown in
2.2. AGBR setup A lab-scale AGBR was fabricated using a plexiglass column having an inner diameter of 5 cm, and height of 32 cm, which was filled up to 17 cm with 210 g of WAC granules (size ~ 1.5–2 mm). The AGBR setup is shown in Fig. 1. The empty bed contact time (EBCT) were measured based on the total bed volume of 333 cm3 (height of 17 cm). A total of seven ports were designed along the depth of the bioreactor. The top and bottom most ports were designated as the inlet and effluent ports respectively, and the remaining five ports (namely, S1 to S5 with c/c spacing of 3 cm) were used for profile sampling. Two nos. peristaltic pumps (PP 10 and PP 20EX, Miclins, India) were used for regulating the flow of synthetic groundwater and backwash water in the bioreactor. Fe
Fig. 1. Schematic diagram of lab scale set-up of AGBR. 2
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Table 1. Initially, the AGBR was fed with 25 mg/L of sulfate, 50 mg/L of nitrate, 105 mg/L of COD and run at the EBCT of 120 min. Fe (3 mg/L) and Se (200 μg/L) were introduced in the system on the 84th day, after attaining a steady-state with respect to sulfate removal. Once the start-up period was over, the effects of initial Fe (3, 4, 5, 7, and 10 mg/L) were studied. From 264th day onwards, the effects of Se (300, 500, 750 and 1000 μg/L) on the AGBR performance were analyzed. During this experimental period, the AGBR was fed with 10 mg/L of Fe and 75 mg/L of sulfate at the EBCT of 60 min. To optimize Fe and Se removal, sulfate concentration was also varied between 25 and 75 mg/L. To fulfill the requirement of additional carbon at an enhanced influent sulfate (50 and 75 mg/L), influent COD was accordingly increased in the feed.
outfitted with a monochromator. X-ray diffraction (XRD) patterns were recorded for a freeze-dried sample from 5� to 85� in the 2θ range using Cu Kα radiation (λ ¼ 1.54 Å) at 40 mA and 40 kV with 1-sec step duration and 0.05� step size. The XRD results were concluded using the Joint Committee on Powder Diffraction and Standards (JCPDS) from the X’Pert Highscore (Match version 3) software. Samples were further analyzed on field emission transmission elec tron microscopy (FETEM) (JEOL JEM-2100F) to confirm the existence of compound obtained from XRD analysis. Samples for FETEM were pul verized, dispersed in acetone, and then subjected to sonication for 30 min (Kirk et al., 2010). The obtained solution was then diluted and drop-casted on the carbon-coated copper (TEM) grid using a Pasteur pipette. The grid was left to dry in a desiccator for overnight and then used for FETEM analysis.
2.5. Analytical approach
3. Results and discussion
Sulfate, COD, and NO3 were measured by nephelo turbidity meter (4500 SO24 - E), closed refluxed method (5220-C), and UV–Visible spectrophotometer (4500 NO3 -B), respectively (APHA, 2005). Before being analyzed, total selenium and iron samples were acidified with 2 ml 1N HCl and was made up to a final volume of 100 ml to dissolve any precipitates. The acidified samples were finally preserved at 4 � C until further analysis. The total selenium was analyzed (λ ¼ 196 nm) in atomic absorption spectrophotometer (SpectrAA 55B, Varian) assem bled with vapor generation assembly (VGA- 77). For the evaluation of Fe, phenanthroline method (3500-Fe B) and/or AAS (flame mode) was used.
3.1. Start-up phase and effects of initial iron on the performance of AGBR Initially, the AGBR was run for the start-up phase to get fair colo nization of mixed microbial consortium on WAC granules. Although selenium adsorption on activated carbon has been reported (Dhillon and Dhillon, 2016), poor selenium adsorption (less than 10% in 24 h) was observed on WAC which inferred biogenic sulphides played the primary role in selenium removal. Complete NO3 removal was observed from the 1st day of AGBR run, which is in line with a study by Shakya and Ghosh (2019a, 2019b). Sulfate reduction was poor during the first few days (day 0–25) of start-up and only 4–14 mg/L of it was removed across the AGBR. This could be attributed to the prolonged adaptation period required for SRB in biofilm, comprising of a wide array of active bac terial genera (Frunzo et al., 2012). Sulfate reduction was improved slowly, and only 10–12 mg/L of it was left out in the effluent during the last few days (day 61–83) of start-up phase (Fig. A1). With the short ening of EBCT, sulfate reduction was hampered only for first few days of operation. (Shakya and Ghosh, 2018a). Similarly, COD in the effluent was also stabilized slowly, and only 12–15 mg/L remained during the last few days (day 61–83). Once steady state in terms of sulfate removal was attained, EBCT was reduced stepwise to 90 min and 60 min on 61st and 74th day, respectively. After attaining stabilized sulfate reduction and EBCT optimization, Se (200 μg/L) and Fe (3 mg/L) were introduced into the AGBR on 84th day. The effects of varying initial Fe on the performance of AGBR is shown
2.6. Mineralogical analysis The collection of biosolids from the backwash suspension was done in a nitrogen gas-filled container. The collected backwash suspension was centrifuged for 10 min at 5000 rpm to get a semi-liquid paste, which was freeze-dried and kept in sealed Falcon™ tubes at 4 � C for preser vation. To limit the contact between ambient air and biosolids, tubes were stored in an airtight double zip lock pouches and opened only in anoxic environment. The elemental confirmation in the backwash bio solids was carried out by field emission scanning electron microscopy (FESEM) (Sigma, Zeiss, Germany) and energy dispersive X-ray (EDX) (Oxford, INCA 300, UK). Before analysis, the freeze-dried samples were spread on the FESEM stub (stuck with carbon tape) and then gold coated. To check the presence of any crystalline phase, biosolids were characterized using the PANalytical X-pert PRO MPD diffractometer Table 1 Operating schedule of AGBR. Purpose
Start-up phase Effect of initial iron
Effect of initial selenium
Operating Days
Concentration in Influent Feed Selenium (μg/ L)
Iron (mg/ L)
Sulfate (mg/ L)
Nitrate (mg/ L)
COD (mg/ L)
1–60 61–73 74–83 84–103 104–123 124–142 143–172 173–180
– – – 200 200 200 200 300
– – – 3 4 4 5 5
25 25 25 25 25 25 25 25
50 50 50 50 50 50 50 50
105 105 105 105 105 105 105 105
120 90 60 60 60 45 45 45
181–198
300
5
25
50
105
60
199–227
300
5
50
50
121
60
228–248 249–263 264–272 264–272 273–292 293–312 313–332
300 300 300 300 500 750 1000
7 7 10 10 10 10 10
50 75 75 75 75 75 75
50 50 50 50 50 50 50
121 134 134 134 134 134 134
60 60 60 60 60 60 60
3
EBCT (minutes)
Remarks
Profile sampling on day 180 Profile sampling on day 181 Profile sampling on day 227
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in Fig. 2. Iron removal (initial 4 mg/L) met the permissible limit in drinking water in presence of sulfate (initial 25 mg/L) and 45 min EBCT. Iron between 0.42 and 0.91 mg/L was observed in the effluent from day 143–198 (initial Fe of 5 mg/L). As Fe removal in sulfidogenic systems is mainly due to the formation of biogenic iron sulfides, the plausible explanation behind elevated Fe in the effluent could be due to insuffi cient amount of sulfides that is necessary for the precipitation as iron sulfides (Shakya and Ghosh, 2019a, 2019b). As further sulfate reduction was not anticipated in AGBR, EBCT was extended to 60 min (from 45 min) on 181th day to facilitate longer reaction and contact time be tween biogenic sulphides and iron (for iron sulfides precipitation). This resulted in better sulfate removal and eventually improved Fe removal. However, Fe (0.42 mg/L) in the effluent was still higher than the BIS permissible limit. Hypothesizing insufficient sulfides for iron sulfides precipitation, influent sulfate was raised to 50 mg/L (on 199th day), which boosted Fe removal to well below the permissible limit. At initial Fe of 7 mg/L (day 228–248), 1.5 mg/L of Fe appeared once more in the effluent, necessitating more sulphides requirement. Thus initial sulfate was further raised to 75 mg/L on day 249. On increasing initial sulfate, Fe in effluent was reduced below the permissible limit, with around 11–15 mg/L of sulfate remaining. Higher EBCT and sulfate are docu mented to ameliorate a bioreactor performance (Shakya and Ghosh, 2018a), whereas lower EBCT may result in poor sulfate removal and/or limited contact time for biogenic sulfides formation, which in turn left high iron in the effluent (Upadhyaya et al., 2010). Further increase in initial iron to 10 mg/L (day 264–272) did not hampered iron removal and effluent iron remained below drinking standards during subsequent AGBR run. Effluent Se remained well below the BIS permissible limit for most of the time. During the entire phase, effluent pH was always in the range of 7.2–7.6, which was slightly more than the raw water. Increase in pH could be attributed to the formation of alkalinity during oxyanions (NO3 and sulfate) reduction (Brahmacharimayum and Ghosh, 2014). Effluent COD was in the range of 10–13 mg/L except for few data, which were comparatively better as described by Shakya and Ghosh (2018a). Higher COD of 22 and 25 mg/L was observed on day 199 and 249, respectively and was possibly due to extra carbon added to meet carbon
requirement for elevated sulfate levels (Shakya and Ghosh, 2018a). 3.2. Effects of selenium concentration on the performance of AGBR The effects of initial Se concentrations of 300, 500, 750, and 1000 μg/L on the performance of AGBR are shown in Fig. 3. During this experimental period, AGBR was operated at initial Fe, NO3 , sulfate, and COD of 10 mg/L, 50 mg/L, 75 mg/L, and 134 mg/L, respectively at 60 min EBCT. The above-mentioned feeding and operating conditions were chosen as per the effective performance of AGBR at 10 mg/L of initial Fe concentration. Despite increment in the initial Se concentra tions, Se and Fe in the effluent remained well below the BIS permissible limit. Furthermore, with the increase in influent Se, increment in Fe removal was observed (day 273–332), which could be attributed to the formation of achavalite (Fe–Se) in the AGBR (Gerhardt et al., 1991). During the last few days (day 313–332) of this study, both Fe and Se removal were found to be around 99%, whereas NO3 was completely removed from the bioreactor. The plausible cause for better selenium and iron removal could be due to availability of sufficient sulphides and/or reaction time for precipitation of selenium and/or iron as their sulphides. In the AGBR, selenium was likely eliminated from the liquid phase through direct selenium sulfide precipitation and/or surface precipitation on iron sulfide solids (Mitchell et al., 2013). Adsorption on iron sulfides might have provided additional selenium removal. Previ ous reports also suggested iron compounds as good adsorbents of sele nium (Dhillon and Dhillon, 2016). However, an overdose of nitrate and sulfate should be avoided, that would otherwise raise carbonate alka linity (Altun et al., 2014) and sulphide levels (Luo et al., 2008). Although precipitation of metal sulfides is favored at high alkalinity (Hall, 1986), it can dissolute precipitates of metal sulfides (Henke, 2009) and may cause an increase in selenium and iron in the effluent. Effluent sulfate and COD were observed to be around 6.5 and 11.5 mg/L, respectively during this phase. Effluent pH remained between 7.40 and 7.65, which was in the range of permissible limit for drinking water. These results showed that the Se increment of up to 1000 μg/L did not alter the performance of the AGBR.
Fig. 2. Effects of initial iron concentration on the performance of AGBR. 4
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Fig. 3. Effects of initial selenium concentration on the performance of AGBR.
3.3. Port profile analysis
standards.
To illustrate the variations in the concentration of contaminants along the AGBR bed depth, port profile sampling were performed on 180th, 181th and 227th day of the reactor operation (Fig. 4). It also depicted the effects of EBCT and sulfate concentrations on Fe and Se removal efficiencies. The samples were taken anaerobically from each sampling port (S1 to S5 and effluent) and analyzed for Se, Fe, NO3 , sulfate, and COD. As shown in Fig. 4, complete removal of NO3 (50 mg/ L) was observed in port S1, and sulfate removal occurs after that. This is mainly because sulfate is thermodynamically an inferior electron acceptor than NO3 . Similar results were also reported by Shakya and Ghosh (2019a, 2019b) during simultaneous removal of nitrate and sul fate using mixed microbial consortium. It was observed (Fig. 4) that Fe and Se removal followed sulfate removal in AGBR. Although some fluctuations in Fe concentration was observed, Se in the effluent was always below the permissible limit until it reached the port S5. The results of port profile sampling performed on day 180 and 181 (45 min EBCT) are shown in Fig. 4 (a, b). As observed in Fig. 4 (a) sulfate removal across the system was around 18 mg/L, Fe removal followed sulfate removal, and 0.85 mg/L of Fe was left out in port S5. Most of the sulfate was removed in port S4, and no further sulfate reduction was observed. On day 181, no considerable improvement in effluent sulfate and Fe removal was observed (Fig. 4 (b)) when EBCT was increased to 60 min. This implies that poor Fe removal in the system was mainly not associated with insufficient contact time. Considering the insufficient sulfide availability the cause, it was decided to increase the influent sulfate keeping the EBCT constant at 60 min. Fig. 4 (c) (day 227th) indicated that around 37.5 mg/L (till port S4), sulfate was removed across the system, which in turn enhanced Fe removal. This inferred that increased influent sulfate provided enough sulfides to precipitate as biogenic iron sulphides (Shakya and Ghosh, 2018b). Thus, port profile sampling results illustrated that sufficient contact time, and sulfate dose were necessary for the removal of Fe in the AGBR to meet the drinking
3.4. Overall performance of AGBR The outcome of the experiments conducted on performance assess ment of AGBR on simultaneous removal of Se (200–1000 μg/L) and Fe (3–10 mg/L) depicted that the removal of both the pollutants met their permissible limits in drinking water, along with complete removal of NO3 (50 mg/L). An increment in effluent pH was observed but was al ways within the range of drinking water standards. Selenium removal from a sulfidogenic bioreactor occured either through selenium sulfide precipitation (Jung et al., 2016) or adsorption by iron sulfides and/or co-precipitation (Charlet et al., 2012; Diener et al., 2012). Biogenic sulfides formed due to the reduction of sulfate in the bioreactor com bined with the reduced selenium and/or iron to precipitate as their respective sulfides. Although no negative effect on Se removal was observed, Fe in the effluent was found to be more than the permissible limit (at initial Fe of 5 and 7 mg/L) during some experimental condi tions. As an outcome of the sulfidogenic activity, iron removal in the present system took place mostly through the formation of iron sulfides. Therefore, the appearance of Fe (more than 0.3 mg/L) in the effluent could be due to the inadequate amount of sulfides and/or short reaction time (Shakya and Ghosh, 2018b), which was overcome by providing more amount of sulfate and/or longer EBCT. Increasing EBCT alone to 60 min (from 45 min) improved Fe removal (Fig. 2; day 181–198) but could not meet the drinking water standards (as effluent Fe was 0.5 mg/L). However, increase in influent sulfate (25–50 mg/L) along with the increase in EBCT (45 min–60 min) helps in reducing iron in the effluent to below 0.3 mg/L (Fig. 2; day 199–227). Similarly, sulfate dose of 75 mg/L was required for initial Fe or 7 mg/L (Fig. 2; day 249–263). Thus, it can be inferred that proper sulfate dose and sufficient contact time are important parameters to get desirable performance.
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Fig. 4. Port profile sampling of AGBR on day 180 (a) and day 181 (b), showing the effects of EBCT and sampling on day 181 (b) and day 227 (c), showing the effects of sulfate concentrations on selenium and iron removal efficiencies.
3.5. Biogenic precipitate characterization
and iron compared to selenium are probably due to the higher sulfate and iron in the feed, which led to the formation of more iron sulfides in the AGBR. FESEM/EDX results were further validated by XRD analysis. XRD analysis depicted the existence of selenium sulfide (SeS; JCPDS #98-002-1203), achavalite (FeSe; JCPDS #98-011-6498) and pyrite (FeS2 ; JCPDS #98-004-9014) in the biosolids (Fig. 5b). The FESEM and XRD analyses confirmed that the precipitation as iron persulfide (FeS2 )
FESEM micrograph with EDX mapping (Fig. 5a) showed uniformly distributed sulfur, iron and selenium in the biosolids. Presence of rela tively high amounts of sulfur {Fig. A2 (a)} and iron {Fig. A2 (b)} in comparison to selenium (Fig. A2 (c)) were observed in the elemental mapping of an individual element. Presence of higher amounts of sulfur 6
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Fig. 5. (a) FESEM/EDX elemental mapping and (b) XRD pattern of biosolids formed in AGBR.
and selenium sulfide (SeS) were the removal mechanism of Fe and Se in the bioreactor (Jung et al., 2016). Furthermore, Selenium removal was also ascribed to adsorption on iron sulfides (FeS/FeS2 ) and/or co-precipitation as iron selenide (FeSe on FeS/FeS2 ) (Charlet et al., 2012; Diener et al., 2012). Therefore, in the present study, Fe addition enhanced Se removal. The peak intensities showed that the compound was of semi-crystalline in nature, which was further supported by TEM analysis. Fig. 6 showed FETEM images {Fig. 6 (a,b)}, SAED pattern {Fig. 6 (c)}, high resolution transmission electron microscopy (HRTEM) image (Fig. 6 (d)) and corresponding Inverse Fast Fourier Transform (IFFT) image (Fig. 6, e) of the collected biosolids. ImageJ and Gatan Digital Micrograph software (version 3.4) were used to determine the d-spacing from selected area electron diffraction (SAED) pattern and HRTEM image, respectively. The belt-structure showed distinctive clear
lattice fringes with a d-spacing of 0.27 nm {Fig. 6 (e)} which corre sponds to the compound FeS2. The d-spacing values obtained from XRD analysis (sharp peak at 33� 2θ value) were also in line with the TEM analysis. The SAED pattern indicated the precipitated compound exhibiting both crystallinity, and amorphicity in nature, which was also observed in the XRD pattern. Crystalline nature of the obtained pre cipitates suggests the stable nature of biosolids (de Matos et al., 2018), which ensures their safe disposal in a landfill. Moreover, the reduced biosolids from the anaerobic bioreactor offers three-fold benefits: (1) low volume of biosludge generation; (2) stable in anaerobic environ ments, and (3) stable after exposure to air (first aerobic environment) due to formation of oxides of iron, which are also potential adsorbents of Se (Mitchell et al., 2013).
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Fig. 6. (a, b) TEM images, (c) SAED pattern, (d) HRTEM image and (e) IFFT of the backwash biosolids collected from AGBR.
4. Conclusions
References
This study demonstrated the simultaneous removal of Se (200–1000 μg/L) and Fe (3–10 mg/L) to meet drinking water standards. Nitrate (50 mg/L) was completely removed from synthetic groundwater. At initial Fe of 5 and 7 mg/L, more than 0.3 mg/L was left out in the effluent due to an inadequate amount of sulfides or insufficient contact time. Addition of more sulfate (50 and 75 mg/L) in the synthetic water and/or increase in EBCT (60 min) helped in reducing effluent iron concentration. Due to the selection of suitable EBCT and proper assessment of carbon dose, soluble COD in the effluent was only around 11 mg/L. The characterization of biosoilds confirmed the bioprecipitation as SeS, FeSe, and FeS2 were the removal mechanisms of Se and Fe. FETEM analysis suggested the formation of stable bio-sludge in the bioreactor. Being low volume (anaerobic/anoxic process) and stable nature, sludge-handling problem can be expected to be low. This outperforms the physicochemical and aerobic processes of Se removal in view of practical utility. Overall, sulfidogenic bioreactor system can simultaneously remove selenium, iron, and nitrate from polluted groundwater to meet drinking water guidelines at an optimum dose of sulfate and sufficient EBCT.
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Declaration of interest statement The authors declare that we have no conflict of interest. The help in any form received from any party have been acknowledged. Acknowledgments The authors would like to appreciate the assistance from the Central Instrument Facility, Department of Civil Engineering, Physics and Bio sciences and Bioengineering of IIT Guwahati, for giving different instrumental services, manpower and so forth are very recognized. Au thors would also acknowledge Mr. Niranjan Kumar, Mr. Siddhartha Paul, and Mr. Gajjela Rajashekhar, IIT Guwahati, for their help in various required analysis during this project. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.jenvman.2019.109757. 8
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