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Heavy metal removal in an UASB-CW system treating municipal wastewater
Chemosphere 93 (2013) 1317–1323
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Heavy metal removal in an UASB-CW system treating municipal wastewater D. de la Varga, M.A. Díaz, I. Ruiz, M. Soto ⇑ Dept. of Physical Chemistry and Chemical Engineering I, University of A Coruña, Rúa da Fraga n°1, 15008 A Coruña, Galiza, Spain
h i g h l i g h t s We report for the first time the long-term removal of HMs in an UASB-CW system. Reduced conditions in the UASB-CW system treating sewage allow a high HM removal. Overall removals were in the following order: 94% > Sn > Cr > Cu > Pb > Zn > Fe > 63%. Removals of Sn, Cr, Cu, Zn and Ni were maintained at least for the first 5 years.
a r t i c l e
i n f o
Article history: Received 2 June 2013 Received in revised form 13 July 2013 Accepted 15 July 2013 Available online 12 August 2013 Keywords: Heavy metals Anaerobic digesters Constructed wetlands Sewage treatment Long-term efficiency
a b s t r a c t s The objective of the present study was to investigate for the first time the long-term removal of heavy metals (HMs) in a combined UASB-CW system treating municipal wastewater. The research was carried out in a field pilot plant constituted for an up-flow anaerobic sludge bed (UASB) digester as a pretreatment, followed by a surface flow constructed wetland (CW) and finally by a subsurface flow CW. While the UASB showed (pseudo) steady state operational conditions and generated a periodical purge of sludge, CWs were characterised by the progressive accumulation and mineralisation of retained solids. This paper analyses the evolution of HM removal from the water stream over time (over a period of 4.7 year of operation) and the accumulation of HMs in UASB sludge and CW sediments at two horizons of 2.7 and 4.0 year of operation. High removal efficiencies were found for some metals in the following order: Sn > Cr > Cu > Pb > Zn > Fe (63–94%). Medium removal efficiencies were registered for Ni (49%), Hg (42%), and Ag (40%), and finally Mn and As showed negative percentage removals. Removal efficiencies of total HMs were higher in UASB and SF units and lower in the last SSF unit. Ó 2013 Elsevier Ltd. All rights reserved.
1. Introduction Anaerobic digesters, in particular up-flow anaerobic sludge bed (UASB) reactors, have been applied in the last few years for the treatment of municipal wastewater, achieving good removal efficiency of suspended solids and biodegradable organic matter in warmer climate countries (Foresti, 2001). Conversely, in temperate climate conditions where wastewater temperature ranges from 5 to 20 °C, UASB digesters show lower efficiency and anaerobic digesters have been proposed as a wastewater pretreatment (Barros et al., 2008; Álvarez et al., 2008a,b). Indeed, a posttreatment of anaerobic digester effluents could be necessary in both cases when a high effluent quality is required (Álvarez et al., 2008b). Although there are several options for the post-treatment of UASB effluents, extensive systems like constructed wetlands
⇑ Corresponding author. Tel.: +34 981 167 050; fax: +34 981 167 065. E-mail address: [email protected] (M. Soto). 0045-6535/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2013.07.043
(CWs) are able to reach a high effluent quality. Furthermore, combined UASB-CW systems maintain low cost and sustainability characteristics, including simplicity of operation and maintenance. CWs are commonly used for the removal of organic matter, suspended solids, fecal microorganisms and nutrients. These systems also remove heavy metals (HMs) and other specific pollutants (Calijuri et al., 2011). However, there is little information about the removal and fate of HMs in CWs treating municipal wastewater (Vymazal et al., 2010). To our knowledge, no study on the removal of HMs in CWs treating anaerobic effluents exists nor is there one on anaerobic digesters treating municipal wastewater. Under aerobic conditions, HM retention and accumulation in wetland substrate is mainly due to the formation of metal hydroxides (i.e. Fe and Mn hydroxides; Singer and Stumm, 1970). Under anoxic or anaerobic conditions, the precipitation of metal sulphides is the main process contributing to the removal of HMs in CWs. In anaerobic conditions, sulphate reduction leads to the formation of hydrogen sulphide and most HMs react with sulphide to form highly insoluble precipitates (Stumm and Morgan 1981).
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However, in oxidant conditions, metal sulphides become soluble and the metal could be released again (Vymazal et al., 2010). Furthermore, anaerobic conditions favour the resolution of Fe and Mn hydroxides (Green et al., 2003; Mansfeldt, 2004). Thus, metal retention and accumulation in CW sediments could be influenced by spatial and temporal variations in redox conditions. Horizontal flow CWs are usually dominated by anaerobic conditions. Pedescoll et al. (2011a) and Pedescoll et al. (2011b) evaluated the clogging risk and contaminant removal efficiency of CWs when an anaerobic digester was used as a primary treatment alternative to conventional settler. Due to the reduced nature of its effluents, the anaerobic digester generated a more reduced environment of the wetlands which, in turn, did not improve contaminant removal efficiency in wetlands (Pedescoll et al., 2011b). However, the anaerobic digester significantly reduced the amount of suspended solids entering CWs and therefore helped prevent or delay clogging processes. In fact, a low solids accumulation rate in CWs was observed when an anaerobic digester was used as the pretreatment system (Pedescoll et al., 2011a; de la Varga et al., 2013). The observed effects of anaerobic digester pre-treatment lowering solids content and redox potential of the effluents entering CWs probably affects HM removal in CWs, but no experimental studies exist on this subject. The objective of the present study was to investigate for the first time the long-term removal of HMs in a combined UASB-CW system treating municipal wastewater. The research was carried out in a pilot plant constituted for an UASB digester as a pretreatment, followed by a surface flow (SF) CW and finally by a subsurface flow (SSF) CW. This paper analyses the evolution of HM removal from the water stream over time (over a period of 4.7 year of operation) and the accumulation of HMs in UASB sludge and CW sediments at two horizons of 2.7 and 4.0 year of operation.
wetland was designed to work with a water layer of 20 cm above the gravel surface) and 0.5 m for the SSF CW. The average hydraulic retention time (HRT) was 9.3 h in the UASB and 55 h in the CWs. Other details of the pilot plant have been reported by Ruiz et al. (2010). 2.2. Sampling strategy Integrated water samples (every 4 h for a 24 h period) were collected for 3 week in winter 2008 (February), for 3 week in summer 2009 (June–July), and for 2 week in spring 2010 (March–April) from the influent, the UASB reactor effluent, the SF CW effluent and the SSF CW effluent at a rate of 2 sampling campaigns weekly. Successive sampling campaigns corresponded to 2.7, 4 and 4.8 year of plant operation. Samples from 2008 to 2009 were analysed for total and soluble metals (48 wastewater samples in total) while samples from 2010 were only analysed for total metals (16 additional wastewater samples). Samples were collected in 1 L polyethylene bottles and transported in a refrigerated form to the laboratory, where they were stored at 4 °C until analysis. Sludge samples from the UASB purged sludge were obtained for the 2008 and 2009 campaigns (2.7 and 4 year of plant operation), yielding a total of 2 and 3 composite samples per campaign, respectively. Samples from solids accumulated in the gravel bed of CWs over time were obtained for the 2008 and 2009 campaigns, producing one composite sample per wetland unit and campaign. For each wetland unit, composite solid samples were obtained from four punctual samples collected both near the inlet and near the outlet, as indicated in Fig. 1. Sludge and solids samples were collected in 1 L polyethylene bottles and frozen at 18 °C until analysis for total metals. 2.3. Analysis
2. Materials and methods 2.1. Pilot plant description The experimental pilot plant is located in the outskirts of Silvouta (Santiago de Compostela, Galicia, Spain). It was constructed in Spring 2005 on the municipal WWTP grounds. The plant (Fig. 1) consists of an upflow anaerobic sludge bed (UASB) digester and two CW systems. Following sand and grease removal, the UASB was fed with raw municipal wastewater at a hydraulic loading rate of 50–120 m3 d 1. The active volume of the reactor was 25.5 m3, and its total height and diameter were 7.1 and 2.5 m, respectively. The excess sludge generated was removed through a lateral port, located 4.5 m from the bottom, once a week at the most, until the sludge bed fell to the level of this port. A fraction of the UASB effluent (17–20 m3 d 1) was diverted to a SF CW followed by a SSF CW. Both CWs (75 m2 each) were operated in series and were planted with bulrushes (Juncus effusus) at 2 plants m 2. The average size of the gravel in CWs was 6–12 mm, the initial porosity was 0.45 and the density was 1180 kg m 3. The water depth was 0.5 m, while the gravel depths were 0.3 m for the SF CW (so SF
(1) Influent
UASB reactor
Sludge purge
An aliquot part of wastewater samples was filtered through 0.45 mm membrane filters and acidified with 1% concentrated nitric acid before perform determination of dissolved metals. For total metals, an aliquot of 50 mL of raw wastewater was digested with aqua regia (1 mL HNO3 + 3 mL HCl) using a graphite digestion block (DigiPrep of SCP SCIENCE) programmed to reach 95 °C in 80 min and then maintained at 95 °C for 180 min. After digestion, samples were replenished with MilliQ de-ionised water to 50 mL and then filtered through a Millex HN de 0.45 lm filter. An aliquot of sludge samples was air-dried and the solids were finely shredded and mixed. 0.5 g of solids sample was subjected to acid digestion (10 mL HNO3 cc + 1 mL H2O2 cc) in a closed microwave apparatus (ETHOS PLUS de Millestone). The oven temperature was programmed to reach 175 °C in 5.5 min, being held for 4.5 min. After digestion, samples were replenished with Milli-Q de-ionised water to 50 mL and then filtered through a Millex HN de 0.45 lm filter. All metal concentrations were determined using inductively coupled plasma mass spectrometry (ICP-MS Element XR or Element2 from Thermo Electron). Detection limits (DL) for metal SSF -CW
SF -CW (2) UASB Eff.
(3) SF Eff.
(4) SSF Eff.
(water flow excces)
Fig. 1. Schematic set-up of the pilot plant (plant views). Numbers and (d) indicate water sampling points and (s) represents the sample points for UASB sludge and for the solids in the wetlands.
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analysis in wastewater samples were in the following ranges (in lg L 1): 0.03–0.1 for Cd and Hg; 0.1–1.0 for Ag, Sn, As and Pb, and 1 for Cr, Mn, Fe, Ni, Cu and Zn. DLs for sludge samples ranged from 0.05 to 1 mg kg 1. General water parameters were also measured in all wastewater samples following analytical methods described in Standard Methods (APHA, 1995). 3. Results and discussion 3.1. General parameters of the system operation Table 1 reveals the data of general water quality parameters at different sampling points of the hybrid anaerobic digester-CW system. Wastewater temperature ranged from 6 to 22 °C and pH from 5.8 to 7.9. Higher dissolved oxygen (DO) values were registered at the influent point (ranging from 4.4 to 7.6 mg L 1), while DO in the effluents of the three units was indistinctly lower (in the range of 0–2.8 mg L 1). In the same way, redox potential values (ranging from 188 to 325 mV) were indicative of reduced effluents through the three system units. Organic load, as indicated by the content in TSS, COD and BOD, was low, corresponding to diluted municipal wastewater (Henze et al., 2002). Applied surface organic loading rates (SLR) were in the range of 1200–6500 g BOD m 2 d 1 (UASB) and 5–21 g BOD m 2 d 1 (10 g BOD m 2 d 1 on average for the two CW units) (de la Varga et al., 2013). In these conditions, the overall system removed 88% of BOD on average, 86% of COD and 96% of TSS on average. Final effluent concentrations averaged 21 mg BOD L 1, 48 mg COD L 1 and 8 mg SS L 1. The WWTP was operated in agreement with Directive 91/271/EEC concerning urban wastewater treatment.
The average metal content was congruent with the dilute character of this municipal wastewater. The ratio of concentrations found in our study to those concentrations indicated by Henze et al. (2002) for diluted municipal wastewater resulted in 1.34 on average for the several HMs considered (or 0.81 when the HM concentrations for medium concentration municipal wastewater indicated by these authors were considered). However, the ratio for individual metals showed a wider range. Considering the ratios for diluted municipal wastewater, lower values were found for Cd (0.04), Hg (0.18), Pb (0.32) and Ag (0.36) and higher values for Zn (2.0), Cu (3.0) and Cr (3.5). Other metals (i.e. As, Fe, Mn and Ni) showed concentrations near to that indicated by Henze et al. (2002) for diluted municipal wastewater. Similar or slightly lower concentrations of metals were reported for wastewater from small populations treated in CWs (Lesage et al., 2007; Kröpfelová et al., 2009; Arroyo et al., 2010), although the contribution of individual metals varies from case to case. The ratio of concentrations found in these studies to concentrations indicated by Henze et al. (2002) for diluted municipal wastewater resulted in the range of 0.53 to 1.41, on average for the several HMs considered. High values of As, Cd and Hg were reported by Arroyo et al. (2010), while Kröpfelová et al. (2009) indicated high values for Fe, Mn and Ni. The percent of each metal found in a dissolved form in raw wastewater is also indicated in Table 2. The content of metals in a dissolved form ranged from very low values of 5% (Fe) to high values up to 74% (Ni), with an average of 37%. Similar average values of 33% and 35% of total metals in a dissolved form were obtained by Ligero (2001) and Lesage et al. (2007), respectively.
3.3. HM removal from wastewater 3.2. HM concentration in influent wastewater Table 2 shows the concentration of metals in the wastewater influent to the system obtained from the campaigns carried out.
Fig. 2A shows the percentage removal of total and dissolved HMs for the overall UASB-CW system as an average for the monitoring campaigns carried out. Total HM removal through the
Table 1 Minimum, maximum (in brackets) and average concentrations of the water quality parameters in the influent wastewater, UASB, SFCW and SSFCW effluents. Influent
UASB effluent
SF effluent
SSF effluent
T (°C) Winter 2008 Summer 2009 Spring 2010
n/a (19.1–21.6) 20.3 (8.1–15.3) 11.8
n/a (18.6–19.1) 18.9 (7.3–14.7) 10.7
n/a (18.7–19.1) 18.9 (6.5–14.2) 10.1
n/a (18.7–18.9) 18.8 (6.4–12.7) 9.2
OD (mg L 1) Winter 2008 Summer 2009 Spring 2010
(4.4–7.6) 6.4 n/a (5.1–7.6) 6.2
(0–2.4) 1.1 n/a (0.2–1.6) 0.8
(0–1.4) 0.7 n/a (0–0.5) 0.2
(0–2.8) 1.2 n/a (0–0.1) 0.0
pH Winter 2008 Summer 2009 Spring 2010
(5.8–7.3) (6.8–7.1) (7.0–7.9)
(6.3–7.2) (6.8–7.3) (6.6–6.9)
(6.8–7.1) (6.5–7.7) (6.6–6.9)
(6.7–7.1) (6.5–7.2) (6.6–6.8)
Eh (mV) Winter 2008 Summer 2009 Spring 2010
( 99/ 125) 110 (+79/ 101) 25 (+126/ 116) 3
( 290/ 332) ( 202/ 224) ( 129/ 217)
BOD (mg L 1) Winter 2008 Summer 2009 Spring 2010
(69–235) 158 (129–200) 169 (46–180) 87
(32–145) 90 (72–121) 99 (28–75) 47
(0–52) 27 (16–90) 40 (10–22) 16
(0–30) 14 (7–55) 18 (7–13) 10
COD (mg L 1) Winter 2008 Summer 2009 Spring 2010
(222–438) 321 (314–465) 368 (82–230) 166
(92–218) 155 (119–246) 181 (56–136) 87
(<2–66) 41 (50–130) 86 (14–56) 36
(<2–54) 29 (26–65) 53 (10–35) 22
TSS (mg L 1) Winter 2008 Summer 2009 Spring 2010
(130–246) 187 (145–316) 228 (62–134) 87
(32–56) 42 (21–79) 48 (22–39) 34
(5–16) 9 (6–45) 17 (5–18) 11
(<2–5) 4 (4–8) 5 (3–14) 7
309 205 188
( 322/ 328) ( 210/ 254) ( 223/ 238)
325 228 232
( 307/ 324) ( 236/ 260) ( 232/ 247)
318 247 240
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Table 2 Concentration of metal in the wastewater influent to the system.
a
n Cr Mn Fe Ni Cu Zn Ag Cd Sn Hg Pb As
Winter 2008
Summer 2009
Spring 2010
Average (C.V.)b
Fraction soluble (%)c
6 104 ± 75 118 ± 13 1585 ± 1095 9.4 ± 5.0 126 ± 13 410 ± 319 2.4 ± 0. 8 <0.05 41 ± 27 0.24 ± 0.07 19 ± 6 2.6 ± 1.1
6 17 ± 13 80 ± 16 376.4 ± 232.2 4.2 ± 2.2 42 ± 16 114 ± 58 0. 8 ± 0.6 0.14 ± 0.03 8.1 ± 7.9 0.08 ± 0.02 3.9 ± 1.4 2.1 ± 0.3
4 37 ± 17 90 ± 12 598.2 ± 83.8 6.4 ± 4.7 187 ± 63 255 ± 138 1.1 ± 0.2 <0.06 34 ± 29 0.2 ± 0.1 5.4 ± 2.0 1.0 ± 0.0
16 53 (86) 96 (21) 853.3 (75.4) 6.7 (40) 118 (62) 260 (57) 1.4 (60) n/a 28 (63) 0.2 (46) 9.5 (89) 1.9 (43)
12 41 59 5 74 44 34 n/a n/a 8 33 8 68
Concentrations in lg L 1. a Number of composite samples. b Average from the data for the three campaigns followed by coefficient of variation (%) in brackets. c Obtained from the average values for the two first campaigns.
successive monitoring campaigns is presented in Fig. 2B, while the percentage of total and dissolved HM removal at each treatment unit is presented in Fig. 3A and B, respectively. Cd, Ag and Hg were not included in Figs. 2 and 3 because most values of effluent concentration appeared below the DL. In Figs. 2 and 3, HMs are ordered from higher to lower total HM removal for the overall system. According to Fig. 2A, HMs could be classified in the following ranges of total HM removal (overall system):
(A)
Sn Cr Cu Pb Zn
– High removal efficiencies: Sn > Cr > Cu > Pb > Zn > Fe (ranging from 71% to 94%). – Medium removal efficiencies: 50% > Ni > Hg > Ag (ranging from 40% to 49%). – Null or negative removal efficiencies: Mn, As (ranging from 23 to 59%). The removal of dissolved HMs followed the general pattern of total HMs with some exceptions. Similar removals of total and dissolved forms were registered for Cr, Cu, Pn, Zn and Ni, while Sn, Hg and Ag showed lower removal percentages of dissolved forms. A singular case was the behaviour of total and dissolved forms of Fe, as substantial removal of total Fe was achieved in all campaigns for the overall system (and for the two first units of the treatment system, Fig. 3A), but dissolved Fe showed negative percentage removals at the overall treatment system (Fig. 2A) and in particular at UASB and SSF units (Fig. 3B). The behaviour of Fe could be attributed to the high solubility of iron oxides in the anoxi/anaerobic conditions which prevailed in all the units of the treatment system. Furthermore, we must take into account that Fe appeared in the influent mostly in non-dissolved forms (95% of total Fe). Mn and As showed negative removal percentages for both total and dissolved forms. Significant differences between total influent and total effluent concentrations of Mn were not found (p = 0.78), while dissolved concentrations of Mn were significantly higher in the system effluent than in influent (p = 0.000). Conversely, no significant differences between influent and effluent concentrations were found for As, for both total (p = 0.44) and dissolved forms (p = 0.082). The higher solubility of Mn oxides and As species in anaerobic conditions and the fact that Mn does not readily form an insoluble sulphide phase may explain the wash out of this metals from the system (Kröpfelová et al., 2009). Apart from the higher effluent concentrations of Mn, As and dissolved Fe, the UASB-CW showed a high efficiency in removal of most HMs, in particular of Sn, Cr, Cu, Pb, Zn, Fe and Ni. The removal of Cd is unknown as the concentration of this metal for all the system units was below DL. Removals of total Ag (40%) and total Hg
Fe Ni TOTAL
As
DISSOLVED
Mn -60
-40
-20
0
20
40
60
80
100
Sn
(B)
Cr Cu Pb Zn Fe Ni
-60
-40
-20
As
WINTER 08
Mn
SPRING 10
SUMMER 09
0
20
40
60
80
100
HM REMOVAL (%) Fig. 2. HM removal for the overall treatment system: (A) average removal of total and dissolved metals, and (B) removal of total metals for the successive sampling campaigns.
(42%) for the overall system could be underestimated because final concentrations were in most cases below DL. The UASB unit showed more accurately total Ag removal of 48% and total Hg removal of 28%.
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removal of dissolved forms of Sn, Cu, Pb and Ni, whilst strongly increasing the effluent concentration of As and Mn. The amount of total Cu, dissolved Sn and both forms of Fe and As clearly increased after the movement of wastewater through the last treatment unit (SSF). In the rest of cases, SSF unit showed a negligible effect on HM concentration.
Sn
(A)
Cr Cu Pb Zn
3.4. Comparison of HM removal in several CW systems
Fe
The results of our study have been compared with those previously reported by Lesage et al. (2007), Kröpfelová et al. (2009) and Arroyo et al. (2010) for domestic sewage treated in a horizontal flow or hybrid CWs. As we have indicated above (Section 3.2), influent concentrations of these systems were similar to that of the present study. Fig. 4 compares the removal efficiencies found in our study with those of afore-mentioned bibliography. Cd, Ag, Hg and Sn were not included, because some values for these metals were below DLs in our study, or there is no data available in some of the cited studies. The research of Arroyo et al. (2010) was carried out in an extensive (overall HRT of 10.5 d) hybrid CW system which treated domestic sewage from a town of 1600 population equivalent (pe). The system consisted of a settler, a pond (2 m deep), a SF CW unit and a SSF CW unit. Kröpfelová et al. (2009) reported the removal of HMs in three SSF CW systems treating the wastewater from different towns ranging from 100 to 700 pe. These SSF CW units were operated at about 5 m2 pe 1 (SLR ranging from 2.7 to 9.6 g BOD m 2 d 1) after pretreatment in septic or imhoff tanks. Lesage et al. (2007) described the operation of a SSF system treating pre-settled domestic wastewater from a 350 pe at HRT of 6.7 d and 3.7 m2 pe 1. Arroyo et al. (2010) indicated that their system was less successful in removing metals than other CWs, as those described by Kröpfelová et al. (2009). A wide range of removal efficiencies was observed, ranging from 55% for chromium and 73% for manganese. These authors considered that negative average removal of iron and nickel were unexpected results for their system. The results of Kröpfelová et al. (2009) indicated that SSF CWs could be a very useful tool for the removal of trace elements such as aluminium, zinc, copper, lead and chromium. For the system reported by Lesage et al. (2007), removal efficiencies varied between 49% for Ni and 93% for Al, including mean removals of Cd, Cu, Pb, Zn and Cr in this range.
Ni
UASB SF
As
SSF
Mn -60
-40
-20
0
20
40
60
80
100
Sn
(B)
Cr Cu Pb Zn
75%
Fe Ni
UASB
As
SF SSF
Mn -60
-40
-20
0
20
40
60
80
100
HM REMOVAL (%) Fig. 3. HM removals for each treatment unit (% of HM influent to each unit): (A) total HMs, and (B) dissolved HMs.
The removal efficiency of Sn, Cr, Cu, Zn and Ni was maintained from the first to the third campaigns (Fig. 2B) and then along the operation time from 2.7 to 4.8 year of operation. The removal efficiency of these metals resulted slightly lower in the second campaign, probably due to the lower influent concentration (see Table 2). On the other hand, a continuous decrease in removal efficiency along time was observed for Pb and Fe. As shown in Fig. 3, the removal efficiency of total HMs was very similar in UASB and SF units, and clearly lower (even negative) in the SSF unit. Slightly higher percentage removals of Sn, Cu and Zn were obtained in UASB while percentage removals of Cr, Pb, Fe and Ni were higher in SF CW. However, as the SF unit received the effluent from the UASB, the amount of most of the HMs (i.e. Sn, Cr, Cu, Pb, Zn, Fe and Ni) retained in the SF unit was lower (by a factor ranging from 1.4 to 4) than the amount of these metals removed in UASB. This suggests that the sludge from the UASB should be the main sink for these HMs. A correlation between percentage removal in UASB digester and influent concentration (R2 > 0.98) was observed for As, Cr, Mn and Zn and even for Fe (R2 = 0.69), Ni (0.60), Pb (0.60), Cu (0.52) and Sn (0.41). Thus, the percentage removal of HMs in UASB appeared partly influenced by influent concentration. For the other two treatment units, correlations between percentage removal and influent concentration to each unit were not observed. Removal of HMs in a dissolved form took place mainly in the UASB, except for Fe and Mn. The SF unit also contributed to the
110
TOTAL HM REMOVAL (%)
90 70 50 30 10 -10
Fe
Mn
As
Ni
Zn
Pb
Cu
Cr
-30 -50 -70 Reviewed bibliography -90
This study
-110
Fig. 4. Removal efficiencies of several HMs in this study (n = 3 campaigns) compared to the average removal efficiencies (n = 5 different systems) obtained from Arroyo et al. (2010), Kröpfelová et al. (2009) and Lesage et al. (2007).
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Table 3 Average metal content in UASB sludge and accumulated solids in CW units. Campaign I (2.7 year of plant operation)
Cr Mn Fe Ni Cu Zn Ag Cd Sn Hg Pb As
Campaign II (4 year of plant operation)
UASB
SF
SSF
UASB
SF
SSF
720 ± 19 162 ± 10 11 766 ± 684 38 ± 3 707 ± 37 3048 ± 853 14 ± 0.1 0.74 ± 0.02 179 ± 5 2.1 ± 0.1 110 ± 1 11 ± 1
365 228 23 956 60 424 1065 16 0.37 151 1.2 96 15
101 385 32 846 35 95 271 3.0 0.12 45 0.1 47 11
690 ± 40 180 ± 34 10 535 ± 953 59 ± 3 857 ± 48 2135 ± 107 1.1 ± 0.3 1.21 ± 0.05 3.3 ± 1.2 2.3 ± 0.7 150 ± 9 14 ± 2
451 236 18 954 52 562 1446 1.0 0.78 5.8 1.3 106 16
120 343 51 320 59 207 1264 1.4 1.39 10.9 0.9 53 20
There are two elements (Mn and As) for which average outflow concentrations in all studies were higher than inflow concentrations. The same behaviour was found in the present study. As indicated by Lesage et al. (2007) and Kröpfelová et al. (2009), reduced manganese compounds are very soluble and therefore they are washed out under anaerobic conditions. Lesage et al. (2007) also reported exportation of particulate Fe and Arroyo et al. (2010) of total Fe. Conversely, our study showed overall removal of total Fe of 71%, while only Fe export from the last unit was registered. Also Kröpfelová et al. (2009) found removal efficiencies of Fe ranging from 50% to 56% in two of their systems, while the third, (with a lower SLR), showed negative iron removals. Kröpfelová et al. (2009)’s system also showed distinctly lower removals of HMs than the other two. With the exception of Fe, as discussed above, all other metals listed in Fig. 4 showed a similar behaviour. Furthermore, in general, the UASB-CW hybrid system showed a higher HM removal than previous reported studies. This could be attributable to the more reduced conditions in the units of the UASB-CW systems because of the anaerobic processes undergone in the anaerobic digester (Pedescoll et al., 2011b). Low redox potential and oxygen concentration in the effluents from the three units of the system (Table 1) confirm the existence of highly reduced conditions in this system. Conversely, Arroyo et al. (2010) reported higher redox potential and oxygen concentration in their system and lower HM removal. Reduced conditions were partly favoured by the application of higher organic SLR to the present system, which ranged from 8 to 21 g BOD m 2 d 1. So, in agreement with our results and those of Kröpfelová et al. (2009), the application of high organic SLR in CW systems will favour the overall removal of HMs. 3.5. The content of HMs in sludges Table 3 shows the HM content in UASB sludge and solids accumulated in CW units for each sampling campaign. HM concentration in UASB sludge in the same campaign showed little variability, with coefficients of variation (CV) below 10%. Variations between campaigns I and II were also reduced (CV < 35%), except for Ag and Sn which shown anomalous high values for campaign I. Apart from Ag and Sn, the ratio of concentrations for UASB sludge in campaign II to campaign I ranged from 0.7 to 1.7 and showed an average of 1.18. In CW units, apart from Ag and Sn, the average ratio of concentrations in campaign II to campaign I resulted of 1.2 in SF (ranging from 0.9 to 2.0) and 3.3 in SSF (ranging from 0.9 to 11.6). Therefore, HM concentrations in the first unit of CWs showed a reduced variation between 2.7 and 4.0 year of operation, while they suffered a strong increase in the second CW unit for the same period of operation.
In general, CW units showed an increase of HM concentration in accumulated solids. This increase agrees well with the net sedimentary characteristics of CW ecosystems, the predominance of precipitation reactions of inorganic compounds and the mineralisation of organic matter in CW sediments. However, an increase of accumulated solids (from approximately 3–7 kg m 2) between 2.7 and 4.7 year of operation in these CW units has been previously reported (Ruiz et al., 2010; de la Varga et al., 2013). This progressive increase in total accumulated solids could help explain the maintenance of the high removal efficiencies of HMs (see Fig. 2) in combination with moderate increases in HM content of accumulated solids (as indicated in Table 3). As a general rule for most of the analysed metals (Zn, Cr, Cu, Pb, Cd and Hg), metal concentration decreased along the treatment path. For these metals, the concentrations in CW sediments also increased with the operation time. Conversely, Fe and Mn increased their concentration along the treatment path, showing the high concentration in the last CW unit. Furthermore, these metals did not show a defined evolution in time, while Ni and As did not show a defined spatial or temporal evolution. Similar HM profiles along the treatment path were reported by Lesage et al. (2007), except for Fe, as this metal in the plant of Lesage et al. (2007) also presented decreasing concentrations along the treatment path. HM concentrations in SF sediments are in the same order or somewhat higher than those reported by Lesage et al. (2007). 4. Conclusions The UASB-CW hybrid system treating a typical municipal wastewater showed a higher HM removal than reported in some previous studies. This behaviour was attributed to the more reduced conditions in the units of the UASB-CW systems due to the anaerobic processes undergone in the UASB and to the application of a high organic SLR in CW units. High removal efficiencies were found for some metals in the following order: 94% > Sn > Cr > Cu > Pb > Zn > Fe > 63%. Medium removal efficiencies were registered for Ni (49%), Hg (42%), and Ag (40%), although the figures for Hg and Ag could be underestimated because of several effluent values below the DL. Finally, Mn and As showed negative percentage removals although only dissolved Mn showed significant differences between influent and effluent concentrations. Removal efficiencies of total HMs were very similar in UASB and SF units, when expressed on the base of the influent to the respective treatment unit and were clearly lower (even negative) in the SSF unit. A correlation between the percentage removal and influent concentration was observed for most HMs in UASB digester but not in SF and SSF CW units. Removal of HMs in a dissolved form took place mainly in the UASB, except for Fe and Mn. The SF CW
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unit also contributed to the removal of dissolved forms of Sn, Cu, Pb and Ni. The removal efficiency of Sn, Cr, Cu, Zn and Ni was maintained along the operation time (from 2.7 to 4.8 year) while a decrease was observed for Pb, Fe, As and Mn. Acknowledgements This research was supported by the Spanish Department of Education and Science (project CTM2008-06676-C05). The authors thank the firm Aquagest for helping in the plant operation and maintenance. References Álvarez, J.A., Armstrong, E., Gómez, M., Soto, M., 2008a. Anaerobic treatment of low strength municipal wastewater by a two-stage pilot plant under psychrophilic conditions. Bioresour. Technol. 99, 7051–7062. Álvarez, J.A., Ruíz, I., Soto, M., 2008b. Anaerobic digesters as a pretreatment for constructed wetlands. Ecol. Eng. 33, 54–67. APHA, 1995. Standard Methods. Standard Methods for the Examination of Water and Wastewater, 19th ed. American Public Health Association, Washington, DC. Arroyo, P., Ansola, G., de Luis, E., 2010. Effectiveness of a full-scale constructed wetland for the removal of metals from domestic wastewater. Water Air Soil Pollut. 210, 473–481. Barros, P., Ruiz, I., Soto, M., 2008. Performance of an anaerobic digester-wetland system for a small community. Ecol. Eng. 33, 142–149. Calijuri, M.L., da Fonseca Santiago, A., Fernandes Moreira Neto, R., de Castro Carvalho, I., 2011. Evaluation of the ability of a natural wetland to remove heavy metals generated by runways and other paved areas from an airport complex in Brazil. Water Air Soil Pollut. 219, 319–327. de la Varga, D., Díaz, M.A., Ruiz, I., Soto, M., 2013. Avoiding clogging in constructed wetlands by using anaerobic digesters as pretreatment. Ecol. Eng. 52, 262–269.
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Foresti, E., 2001. Perspectives on anaerobic treatment in developing countries. Water Sci. Technol. 44 (8), 141–148. Green, C.H., Heil, D.M., Cardon, G.E., Butters, G.L., Kelly, E.F., 2003. Solubilization of manganese and trace metals in soils affected by acid mine runoff. J. Environ. Qual. 32, 1323–1334. Henze, M., Harremoes, P., Jansens, J., Arvin, E., 2002. Wastewater Treatment. Springer-Verlag, New York. Kröpfelová, L., Vymazal, J., Svehla, J., Stíchová, J., 2009. Removal of trace elements in three horizontal sub-surface flow constructed wetlands in the Czech Republic. Environ. Pollut. 157, 1186–1194. Lesage, E., Rousseau, D.P.L., Meers, E., Tack, F.M.G., De Pauw, N., 2007. Accumulation of metals in a horizontal subsurface flow constructed wetland treating domestic wastewater in Flanders, Belgium. Sci. Total Environ. 380, 102–115. Ligero, P., 2001. Hydrolisis and Granulation in Anaerobic Treatment of Urban Wastewater. PhD thesis. Faculty of Science, University of A Coruña, A Coruña. Mansfeldt, T., 2004. Redox potential of bulk soil and soil solution concentration of nitrate, manganese, iron, and sulfate in two Gleysols. J. Plant Nutr. Soil Sci. 167, 7–16. Pedescoll, A., Corzo, A., Álvarez, E., García, J., Puigagut, J., 2011a. The effect of primary treatment and flow regime on clogging development in horizontal subsurface flow constructed wetlands: an experimental evaluation. Water Res. 45, 3579–3589. Pedescoll, A., Corzo, A., Álvarez, E., Puigagut, J., García, J., 2011b. Contaminant removal efficiency depending on primary treatment and operational strategy in horizontal subsurface flow treatment wetlands. Ecol. Eng. 37, 372–380. Ruiz, I., Díaz, M.A., Crujeiras, B., García, J., Soto, M., 2010. Solids hydrolysis and accumulation in a hybrid anaerobic digester-constructed wetlands system. Ecol. Eng. 36, 1007–1016. Singer, P.C., Stumm, W., 1970. Acid mine drainage—the rate limiting step. Science 167, 1121–1123. Stumm, W., Morgan, J.J., 1981. Aquatic Chemistry. An Introduction Emphasizing Chemical Equilibria in Natural Waters, 2nd ed. Wiley-Interscience, New York. Vymazal, J., Jaroslav, S., Kröpfelová, L., Nemcová, J., Suchy, V., 2010. Heavy metals in sediments from constructed wetlands treating municipal wastewater. Biogeochemistry 101, 335–356.
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Heavy metal removal in an UASB-CW system treating municipal wastewater D. de la Varga, M.A. Díaz, I. Ruiz, M. Soto ⇑ Dept. of Physical Chemistry and Chemical Engineering I, University of A Coruña, Rúa da Fraga n°1, 15008 A Coruña, Galiza, Spain
h i g h l i g h t s We report for the first time the long-term removal of HMs in an UASB-CW system. Reduced conditions in the UASB-CW system treating sewage allow a high HM removal. Overall removals were in the following order: 94% > Sn > Cr > Cu > Pb > Zn > Fe > 63%. Removals of Sn, Cr, Cu, Zn and Ni were maintained at least for the first 5 years.
a r t i c l e
i n f o
Article history: Received 2 June 2013 Received in revised form 13 July 2013 Accepted 15 July 2013 Available online 12 August 2013 Keywords: Heavy metals Anaerobic digesters Constructed wetlands Sewage treatment Long-term efficiency
a b s t r a c t s The objective of the present study was to investigate for the first time the long-term removal of heavy metals (HMs) in a combined UASB-CW system treating municipal wastewater. The research was carried out in a field pilot plant constituted for an up-flow anaerobic sludge bed (UASB) digester as a pretreatment, followed by a surface flow constructed wetland (CW) and finally by a subsurface flow CW. While the UASB showed (pseudo) steady state operational conditions and generated a periodical purge of sludge, CWs were characterised by the progressive accumulation and mineralisation of retained solids. This paper analyses the evolution of HM removal from the water stream over time (over a period of 4.7 year of operation) and the accumulation of HMs in UASB sludge and CW sediments at two horizons of 2.7 and 4.0 year of operation. High removal efficiencies were found for some metals in the following order: Sn > Cr > Cu > Pb > Zn > Fe (63–94%). Medium removal efficiencies were registered for Ni (49%), Hg (42%), and Ag (40%), and finally Mn and As showed negative percentage removals. Removal efficiencies of total HMs were higher in UASB and SF units and lower in the last SSF unit. Ó 2013 Elsevier Ltd. All rights reserved.
1. Introduction Anaerobic digesters, in particular up-flow anaerobic sludge bed (UASB) reactors, have been applied in the last few years for the treatment of municipal wastewater, achieving good removal efficiency of suspended solids and biodegradable organic matter in warmer climate countries (Foresti, 2001). Conversely, in temperate climate conditions where wastewater temperature ranges from 5 to 20 °C, UASB digesters show lower efficiency and anaerobic digesters have been proposed as a wastewater pretreatment (Barros et al., 2008; Álvarez et al., 2008a,b). Indeed, a posttreatment of anaerobic digester effluents could be necessary in both cases when a high effluent quality is required (Álvarez et al., 2008b). Although there are several options for the post-treatment of UASB effluents, extensive systems like constructed wetlands
⇑ Corresponding author. Tel.: +34 981 167 050; fax: +34 981 167 065. E-mail address: [email protected] (M. Soto). 0045-6535/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2013.07.043
(CWs) are able to reach a high effluent quality. Furthermore, combined UASB-CW systems maintain low cost and sustainability characteristics, including simplicity of operation and maintenance. CWs are commonly used for the removal of organic matter, suspended solids, fecal microorganisms and nutrients. These systems also remove heavy metals (HMs) and other specific pollutants (Calijuri et al., 2011). However, there is little information about the removal and fate of HMs in CWs treating municipal wastewater (Vymazal et al., 2010). To our knowledge, no study on the removal of HMs in CWs treating anaerobic effluents exists nor is there one on anaerobic digesters treating municipal wastewater. Under aerobic conditions, HM retention and accumulation in wetland substrate is mainly due to the formation of metal hydroxides (i.e. Fe and Mn hydroxides; Singer and Stumm, 1970). Under anoxic or anaerobic conditions, the precipitation of metal sulphides is the main process contributing to the removal of HMs in CWs. In anaerobic conditions, sulphate reduction leads to the formation of hydrogen sulphide and most HMs react with sulphide to form highly insoluble precipitates (Stumm and Morgan 1981).
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However, in oxidant conditions, metal sulphides become soluble and the metal could be released again (Vymazal et al., 2010). Furthermore, anaerobic conditions favour the resolution of Fe and Mn hydroxides (Green et al., 2003; Mansfeldt, 2004). Thus, metal retention and accumulation in CW sediments could be influenced by spatial and temporal variations in redox conditions. Horizontal flow CWs are usually dominated by anaerobic conditions. Pedescoll et al. (2011a) and Pedescoll et al. (2011b) evaluated the clogging risk and contaminant removal efficiency of CWs when an anaerobic digester was used as a primary treatment alternative to conventional settler. Due to the reduced nature of its effluents, the anaerobic digester generated a more reduced environment of the wetlands which, in turn, did not improve contaminant removal efficiency in wetlands (Pedescoll et al., 2011b). However, the anaerobic digester significantly reduced the amount of suspended solids entering CWs and therefore helped prevent or delay clogging processes. In fact, a low solids accumulation rate in CWs was observed when an anaerobic digester was used as the pretreatment system (Pedescoll et al., 2011a; de la Varga et al., 2013). The observed effects of anaerobic digester pre-treatment lowering solids content and redox potential of the effluents entering CWs probably affects HM removal in CWs, but no experimental studies exist on this subject. The objective of the present study was to investigate for the first time the long-term removal of HMs in a combined UASB-CW system treating municipal wastewater. The research was carried out in a pilot plant constituted for an UASB digester as a pretreatment, followed by a surface flow (SF) CW and finally by a subsurface flow (SSF) CW. This paper analyses the evolution of HM removal from the water stream over time (over a period of 4.7 year of operation) and the accumulation of HMs in UASB sludge and CW sediments at two horizons of 2.7 and 4.0 year of operation.
wetland was designed to work with a water layer of 20 cm above the gravel surface) and 0.5 m for the SSF CW. The average hydraulic retention time (HRT) was 9.3 h in the UASB and 55 h in the CWs. Other details of the pilot plant have been reported by Ruiz et al. (2010). 2.2. Sampling strategy Integrated water samples (every 4 h for a 24 h period) were collected for 3 week in winter 2008 (February), for 3 week in summer 2009 (June–July), and for 2 week in spring 2010 (March–April) from the influent, the UASB reactor effluent, the SF CW effluent and the SSF CW effluent at a rate of 2 sampling campaigns weekly. Successive sampling campaigns corresponded to 2.7, 4 and 4.8 year of plant operation. Samples from 2008 to 2009 were analysed for total and soluble metals (48 wastewater samples in total) while samples from 2010 were only analysed for total metals (16 additional wastewater samples). Samples were collected in 1 L polyethylene bottles and transported in a refrigerated form to the laboratory, where they were stored at 4 °C until analysis. Sludge samples from the UASB purged sludge were obtained for the 2008 and 2009 campaigns (2.7 and 4 year of plant operation), yielding a total of 2 and 3 composite samples per campaign, respectively. Samples from solids accumulated in the gravel bed of CWs over time were obtained for the 2008 and 2009 campaigns, producing one composite sample per wetland unit and campaign. For each wetland unit, composite solid samples were obtained from four punctual samples collected both near the inlet and near the outlet, as indicated in Fig. 1. Sludge and solids samples were collected in 1 L polyethylene bottles and frozen at 18 °C until analysis for total metals. 2.3. Analysis
2. Materials and methods 2.1. Pilot plant description The experimental pilot plant is located in the outskirts of Silvouta (Santiago de Compostela, Galicia, Spain). It was constructed in Spring 2005 on the municipal WWTP grounds. The plant (Fig. 1) consists of an upflow anaerobic sludge bed (UASB) digester and two CW systems. Following sand and grease removal, the UASB was fed with raw municipal wastewater at a hydraulic loading rate of 50–120 m3 d 1. The active volume of the reactor was 25.5 m3, and its total height and diameter were 7.1 and 2.5 m, respectively. The excess sludge generated was removed through a lateral port, located 4.5 m from the bottom, once a week at the most, until the sludge bed fell to the level of this port. A fraction of the UASB effluent (17–20 m3 d 1) was diverted to a SF CW followed by a SSF CW. Both CWs (75 m2 each) were operated in series and were planted with bulrushes (Juncus effusus) at 2 plants m 2. The average size of the gravel in CWs was 6–12 mm, the initial porosity was 0.45 and the density was 1180 kg m 3. The water depth was 0.5 m, while the gravel depths were 0.3 m for the SF CW (so SF
(1) Influent
UASB reactor
Sludge purge
An aliquot part of wastewater samples was filtered through 0.45 mm membrane filters and acidified with 1% concentrated nitric acid before perform determination of dissolved metals. For total metals, an aliquot of 50 mL of raw wastewater was digested with aqua regia (1 mL HNO3 + 3 mL HCl) using a graphite digestion block (DigiPrep of SCP SCIENCE) programmed to reach 95 °C in 80 min and then maintained at 95 °C for 180 min. After digestion, samples were replenished with MilliQ de-ionised water to 50 mL and then filtered through a Millex HN de 0.45 lm filter. An aliquot of sludge samples was air-dried and the solids were finely shredded and mixed. 0.5 g of solids sample was subjected to acid digestion (10 mL HNO3 cc + 1 mL H2O2 cc) in a closed microwave apparatus (ETHOS PLUS de Millestone). The oven temperature was programmed to reach 175 °C in 5.5 min, being held for 4.5 min. After digestion, samples were replenished with Milli-Q de-ionised water to 50 mL and then filtered through a Millex HN de 0.45 lm filter. All metal concentrations were determined using inductively coupled plasma mass spectrometry (ICP-MS Element XR or Element2 from Thermo Electron). Detection limits (DL) for metal SSF -CW
SF -CW (2) UASB Eff.
(3) SF Eff.
(4) SSF Eff.
(water flow excces)
Fig. 1. Schematic set-up of the pilot plant (plant views). Numbers and (d) indicate water sampling points and (s) represents the sample points for UASB sludge and for the solids in the wetlands.
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analysis in wastewater samples were in the following ranges (in lg L 1): 0.03–0.1 for Cd and Hg; 0.1–1.0 for Ag, Sn, As and Pb, and 1 for Cr, Mn, Fe, Ni, Cu and Zn. DLs for sludge samples ranged from 0.05 to 1 mg kg 1. General water parameters were also measured in all wastewater samples following analytical methods described in Standard Methods (APHA, 1995). 3. Results and discussion 3.1. General parameters of the system operation Table 1 reveals the data of general water quality parameters at different sampling points of the hybrid anaerobic digester-CW system. Wastewater temperature ranged from 6 to 22 °C and pH from 5.8 to 7.9. Higher dissolved oxygen (DO) values were registered at the influent point (ranging from 4.4 to 7.6 mg L 1), while DO in the effluents of the three units was indistinctly lower (in the range of 0–2.8 mg L 1). In the same way, redox potential values (ranging from 188 to 325 mV) were indicative of reduced effluents through the three system units. Organic load, as indicated by the content in TSS, COD and BOD, was low, corresponding to diluted municipal wastewater (Henze et al., 2002). Applied surface organic loading rates (SLR) were in the range of 1200–6500 g BOD m 2 d 1 (UASB) and 5–21 g BOD m 2 d 1 (10 g BOD m 2 d 1 on average for the two CW units) (de la Varga et al., 2013). In these conditions, the overall system removed 88% of BOD on average, 86% of COD and 96% of TSS on average. Final effluent concentrations averaged 21 mg BOD L 1, 48 mg COD L 1 and 8 mg SS L 1. The WWTP was operated in agreement with Directive 91/271/EEC concerning urban wastewater treatment.
The average metal content was congruent with the dilute character of this municipal wastewater. The ratio of concentrations found in our study to those concentrations indicated by Henze et al. (2002) for diluted municipal wastewater resulted in 1.34 on average for the several HMs considered (or 0.81 when the HM concentrations for medium concentration municipal wastewater indicated by these authors were considered). However, the ratio for individual metals showed a wider range. Considering the ratios for diluted municipal wastewater, lower values were found for Cd (0.04), Hg (0.18), Pb (0.32) and Ag (0.36) and higher values for Zn (2.0), Cu (3.0) and Cr (3.5). Other metals (i.e. As, Fe, Mn and Ni) showed concentrations near to that indicated by Henze et al. (2002) for diluted municipal wastewater. Similar or slightly lower concentrations of metals were reported for wastewater from small populations treated in CWs (Lesage et al., 2007; Kröpfelová et al., 2009; Arroyo et al., 2010), although the contribution of individual metals varies from case to case. The ratio of concentrations found in these studies to concentrations indicated by Henze et al. (2002) for diluted municipal wastewater resulted in the range of 0.53 to 1.41, on average for the several HMs considered. High values of As, Cd and Hg were reported by Arroyo et al. (2010), while Kröpfelová et al. (2009) indicated high values for Fe, Mn and Ni. The percent of each metal found in a dissolved form in raw wastewater is also indicated in Table 2. The content of metals in a dissolved form ranged from very low values of 5% (Fe) to high values up to 74% (Ni), with an average of 37%. Similar average values of 33% and 35% of total metals in a dissolved form were obtained by Ligero (2001) and Lesage et al. (2007), respectively.
3.3. HM removal from wastewater 3.2. HM concentration in influent wastewater Table 2 shows the concentration of metals in the wastewater influent to the system obtained from the campaigns carried out.
Fig. 2A shows the percentage removal of total and dissolved HMs for the overall UASB-CW system as an average for the monitoring campaigns carried out. Total HM removal through the
Table 1 Minimum, maximum (in brackets) and average concentrations of the water quality parameters in the influent wastewater, UASB, SFCW and SSFCW effluents. Influent
UASB effluent
SF effluent
SSF effluent
T (°C) Winter 2008 Summer 2009 Spring 2010
n/a (19.1–21.6) 20.3 (8.1–15.3) 11.8
n/a (18.6–19.1) 18.9 (7.3–14.7) 10.7
n/a (18.7–19.1) 18.9 (6.5–14.2) 10.1
n/a (18.7–18.9) 18.8 (6.4–12.7) 9.2
OD (mg L 1) Winter 2008 Summer 2009 Spring 2010
(4.4–7.6) 6.4 n/a (5.1–7.6) 6.2
(0–2.4) 1.1 n/a (0.2–1.6) 0.8
(0–1.4) 0.7 n/a (0–0.5) 0.2
(0–2.8) 1.2 n/a (0–0.1) 0.0
pH Winter 2008 Summer 2009 Spring 2010
(5.8–7.3) (6.8–7.1) (7.0–7.9)
(6.3–7.2) (6.8–7.3) (6.6–6.9)
(6.8–7.1) (6.5–7.7) (6.6–6.9)
(6.7–7.1) (6.5–7.2) (6.6–6.8)
Eh (mV) Winter 2008 Summer 2009 Spring 2010
( 99/ 125) 110 (+79/ 101) 25 (+126/ 116) 3
( 290/ 332) ( 202/ 224) ( 129/ 217)
BOD (mg L 1) Winter 2008 Summer 2009 Spring 2010
(69–235) 158 (129–200) 169 (46–180) 87
(32–145) 90 (72–121) 99 (28–75) 47
(0–52) 27 (16–90) 40 (10–22) 16
(0–30) 14 (7–55) 18 (7–13) 10
COD (mg L 1) Winter 2008 Summer 2009 Spring 2010
(222–438) 321 (314–465) 368 (82–230) 166
(92–218) 155 (119–246) 181 (56–136) 87
(<2–66) 41 (50–130) 86 (14–56) 36
(<2–54) 29 (26–65) 53 (10–35) 22
TSS (mg L 1) Winter 2008 Summer 2009 Spring 2010
(130–246) 187 (145–316) 228 (62–134) 87
(32–56) 42 (21–79) 48 (22–39) 34
(5–16) 9 (6–45) 17 (5–18) 11
(<2–5) 4 (4–8) 5 (3–14) 7
309 205 188
( 322/ 328) ( 210/ 254) ( 223/ 238)
325 228 232
( 307/ 324) ( 236/ 260) ( 232/ 247)
318 247 240
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Table 2 Concentration of metal in the wastewater influent to the system.
a
n Cr Mn Fe Ni Cu Zn Ag Cd Sn Hg Pb As
Winter 2008
Summer 2009
Spring 2010
Average (C.V.)b
Fraction soluble (%)c
6 104 ± 75 118 ± 13 1585 ± 1095 9.4 ± 5.0 126 ± 13 410 ± 319 2.4 ± 0. 8 <0.05 41 ± 27 0.24 ± 0.07 19 ± 6 2.6 ± 1.1
6 17 ± 13 80 ± 16 376.4 ± 232.2 4.2 ± 2.2 42 ± 16 114 ± 58 0. 8 ± 0.6 0.14 ± 0.03 8.1 ± 7.9 0.08 ± 0.02 3.9 ± 1.4 2.1 ± 0.3
4 37 ± 17 90 ± 12 598.2 ± 83.8 6.4 ± 4.7 187 ± 63 255 ± 138 1.1 ± 0.2 <0.06 34 ± 29 0.2 ± 0.1 5.4 ± 2.0 1.0 ± 0.0
16 53 (86) 96 (21) 853.3 (75.4) 6.7 (40) 118 (62) 260 (57) 1.4 (60) n/a 28 (63) 0.2 (46) 9.5 (89) 1.9 (43)
12 41 59 5 74 44 34 n/a n/a 8 33 8 68
Concentrations in lg L 1. a Number of composite samples. b Average from the data for the three campaigns followed by coefficient of variation (%) in brackets. c Obtained from the average values for the two first campaigns.
successive monitoring campaigns is presented in Fig. 2B, while the percentage of total and dissolved HM removal at each treatment unit is presented in Fig. 3A and B, respectively. Cd, Ag and Hg were not included in Figs. 2 and 3 because most values of effluent concentration appeared below the DL. In Figs. 2 and 3, HMs are ordered from higher to lower total HM removal for the overall system. According to Fig. 2A, HMs could be classified in the following ranges of total HM removal (overall system):
(A)
Sn Cr Cu Pb Zn
– High removal efficiencies: Sn > Cr > Cu > Pb > Zn > Fe (ranging from 71% to 94%). – Medium removal efficiencies: 50% > Ni > Hg > Ag (ranging from 40% to 49%). – Null or negative removal efficiencies: Mn, As (ranging from 23 to 59%). The removal of dissolved HMs followed the general pattern of total HMs with some exceptions. Similar removals of total and dissolved forms were registered for Cr, Cu, Pn, Zn and Ni, while Sn, Hg and Ag showed lower removal percentages of dissolved forms. A singular case was the behaviour of total and dissolved forms of Fe, as substantial removal of total Fe was achieved in all campaigns for the overall system (and for the two first units of the treatment system, Fig. 3A), but dissolved Fe showed negative percentage removals at the overall treatment system (Fig. 2A) and in particular at UASB and SSF units (Fig. 3B). The behaviour of Fe could be attributed to the high solubility of iron oxides in the anoxi/anaerobic conditions which prevailed in all the units of the treatment system. Furthermore, we must take into account that Fe appeared in the influent mostly in non-dissolved forms (95% of total Fe). Mn and As showed negative removal percentages for both total and dissolved forms. Significant differences between total influent and total effluent concentrations of Mn were not found (p = 0.78), while dissolved concentrations of Mn were significantly higher in the system effluent than in influent (p = 0.000). Conversely, no significant differences between influent and effluent concentrations were found for As, for both total (p = 0.44) and dissolved forms (p = 0.082). The higher solubility of Mn oxides and As species in anaerobic conditions and the fact that Mn does not readily form an insoluble sulphide phase may explain the wash out of this metals from the system (Kröpfelová et al., 2009). Apart from the higher effluent concentrations of Mn, As and dissolved Fe, the UASB-CW showed a high efficiency in removal of most HMs, in particular of Sn, Cr, Cu, Pb, Zn, Fe and Ni. The removal of Cd is unknown as the concentration of this metal for all the system units was below DL. Removals of total Ag (40%) and total Hg
Fe Ni TOTAL
As
DISSOLVED
Mn -60
-40
-20
0
20
40
60
80
100
Sn
(B)
Cr Cu Pb Zn Fe Ni
-60
-40
-20
As
WINTER 08
Mn
SPRING 10
SUMMER 09
0
20
40
60
80
100
HM REMOVAL (%) Fig. 2. HM removal for the overall treatment system: (A) average removal of total and dissolved metals, and (B) removal of total metals for the successive sampling campaigns.
(42%) for the overall system could be underestimated because final concentrations were in most cases below DL. The UASB unit showed more accurately total Ag removal of 48% and total Hg removal of 28%.
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removal of dissolved forms of Sn, Cu, Pb and Ni, whilst strongly increasing the effluent concentration of As and Mn. The amount of total Cu, dissolved Sn and both forms of Fe and As clearly increased after the movement of wastewater through the last treatment unit (SSF). In the rest of cases, SSF unit showed a negligible effect on HM concentration.
Sn
(A)
Cr Cu Pb Zn
3.4. Comparison of HM removal in several CW systems
Fe
The results of our study have been compared with those previously reported by Lesage et al. (2007), Kröpfelová et al. (2009) and Arroyo et al. (2010) for domestic sewage treated in a horizontal flow or hybrid CWs. As we have indicated above (Section 3.2), influent concentrations of these systems were similar to that of the present study. Fig. 4 compares the removal efficiencies found in our study with those of afore-mentioned bibliography. Cd, Ag, Hg and Sn were not included, because some values for these metals were below DLs in our study, or there is no data available in some of the cited studies. The research of Arroyo et al. (2010) was carried out in an extensive (overall HRT of 10.5 d) hybrid CW system which treated domestic sewage from a town of 1600 population equivalent (pe). The system consisted of a settler, a pond (2 m deep), a SF CW unit and a SSF CW unit. Kröpfelová et al. (2009) reported the removal of HMs in three SSF CW systems treating the wastewater from different towns ranging from 100 to 700 pe. These SSF CW units were operated at about 5 m2 pe 1 (SLR ranging from 2.7 to 9.6 g BOD m 2 d 1) after pretreatment in septic or imhoff tanks. Lesage et al. (2007) described the operation of a SSF system treating pre-settled domestic wastewater from a 350 pe at HRT of 6.7 d and 3.7 m2 pe 1. Arroyo et al. (2010) indicated that their system was less successful in removing metals than other CWs, as those described by Kröpfelová et al. (2009). A wide range of removal efficiencies was observed, ranging from 55% for chromium and 73% for manganese. These authors considered that negative average removal of iron and nickel were unexpected results for their system. The results of Kröpfelová et al. (2009) indicated that SSF CWs could be a very useful tool for the removal of trace elements such as aluminium, zinc, copper, lead and chromium. For the system reported by Lesage et al. (2007), removal efficiencies varied between 49% for Ni and 93% for Al, including mean removals of Cd, Cu, Pb, Zn and Cr in this range.
Ni
UASB SF
As
SSF
Mn -60
-40
-20
0
20
40
60
80
100
Sn
(B)
Cr Cu Pb Zn
75%
Fe Ni
UASB
As
SF SSF
Mn -60
-40
-20
0
20
40
60
80
100
HM REMOVAL (%) Fig. 3. HM removals for each treatment unit (% of HM influent to each unit): (A) total HMs, and (B) dissolved HMs.
The removal efficiency of Sn, Cr, Cu, Zn and Ni was maintained from the first to the third campaigns (Fig. 2B) and then along the operation time from 2.7 to 4.8 year of operation. The removal efficiency of these metals resulted slightly lower in the second campaign, probably due to the lower influent concentration (see Table 2). On the other hand, a continuous decrease in removal efficiency along time was observed for Pb and Fe. As shown in Fig. 3, the removal efficiency of total HMs was very similar in UASB and SF units, and clearly lower (even negative) in the SSF unit. Slightly higher percentage removals of Sn, Cu and Zn were obtained in UASB while percentage removals of Cr, Pb, Fe and Ni were higher in SF CW. However, as the SF unit received the effluent from the UASB, the amount of most of the HMs (i.e. Sn, Cr, Cu, Pb, Zn, Fe and Ni) retained in the SF unit was lower (by a factor ranging from 1.4 to 4) than the amount of these metals removed in UASB. This suggests that the sludge from the UASB should be the main sink for these HMs. A correlation between percentage removal in UASB digester and influent concentration (R2 > 0.98) was observed for As, Cr, Mn and Zn and even for Fe (R2 = 0.69), Ni (0.60), Pb (0.60), Cu (0.52) and Sn (0.41). Thus, the percentage removal of HMs in UASB appeared partly influenced by influent concentration. For the other two treatment units, correlations between percentage removal and influent concentration to each unit were not observed. Removal of HMs in a dissolved form took place mainly in the UASB, except for Fe and Mn. The SF unit also contributed to the
110
TOTAL HM REMOVAL (%)
90 70 50 30 10 -10
Fe
Mn
As
Ni
Zn
Pb
Cu
Cr
-30 -50 -70 Reviewed bibliography -90
This study
-110
Fig. 4. Removal efficiencies of several HMs in this study (n = 3 campaigns) compared to the average removal efficiencies (n = 5 different systems) obtained from Arroyo et al. (2010), Kröpfelová et al. (2009) and Lesage et al. (2007).
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Table 3 Average metal content in UASB sludge and accumulated solids in CW units. Campaign I (2.7 year of plant operation)
Cr Mn Fe Ni Cu Zn Ag Cd Sn Hg Pb As
Campaign II (4 year of plant operation)
UASB
SF
SSF
UASB
SF
SSF
720 ± 19 162 ± 10 11 766 ± 684 38 ± 3 707 ± 37 3048 ± 853 14 ± 0.1 0.74 ± 0.02 179 ± 5 2.1 ± 0.1 110 ± 1 11 ± 1
365 228 23 956 60 424 1065 16 0.37 151 1.2 96 15
101 385 32 846 35 95 271 3.0 0.12 45 0.1 47 11
690 ± 40 180 ± 34 10 535 ± 953 59 ± 3 857 ± 48 2135 ± 107 1.1 ± 0.3 1.21 ± 0.05 3.3 ± 1.2 2.3 ± 0.7 150 ± 9 14 ± 2
451 236 18 954 52 562 1446 1.0 0.78 5.8 1.3 106 16
120 343 51 320 59 207 1264 1.4 1.39 10.9 0.9 53 20
There are two elements (Mn and As) for which average outflow concentrations in all studies were higher than inflow concentrations. The same behaviour was found in the present study. As indicated by Lesage et al. (2007) and Kröpfelová et al. (2009), reduced manganese compounds are very soluble and therefore they are washed out under anaerobic conditions. Lesage et al. (2007) also reported exportation of particulate Fe and Arroyo et al. (2010) of total Fe. Conversely, our study showed overall removal of total Fe of 71%, while only Fe export from the last unit was registered. Also Kröpfelová et al. (2009) found removal efficiencies of Fe ranging from 50% to 56% in two of their systems, while the third, (with a lower SLR), showed negative iron removals. Kröpfelová et al. (2009)’s system also showed distinctly lower removals of HMs than the other two. With the exception of Fe, as discussed above, all other metals listed in Fig. 4 showed a similar behaviour. Furthermore, in general, the UASB-CW hybrid system showed a higher HM removal than previous reported studies. This could be attributable to the more reduced conditions in the units of the UASB-CW systems because of the anaerobic processes undergone in the anaerobic digester (Pedescoll et al., 2011b). Low redox potential and oxygen concentration in the effluents from the three units of the system (Table 1) confirm the existence of highly reduced conditions in this system. Conversely, Arroyo et al. (2010) reported higher redox potential and oxygen concentration in their system and lower HM removal. Reduced conditions were partly favoured by the application of higher organic SLR to the present system, which ranged from 8 to 21 g BOD m 2 d 1. So, in agreement with our results and those of Kröpfelová et al. (2009), the application of high organic SLR in CW systems will favour the overall removal of HMs. 3.5. The content of HMs in sludges Table 3 shows the HM content in UASB sludge and solids accumulated in CW units for each sampling campaign. HM concentration in UASB sludge in the same campaign showed little variability, with coefficients of variation (CV) below 10%. Variations between campaigns I and II were also reduced (CV < 35%), except for Ag and Sn which shown anomalous high values for campaign I. Apart from Ag and Sn, the ratio of concentrations for UASB sludge in campaign II to campaign I ranged from 0.7 to 1.7 and showed an average of 1.18. In CW units, apart from Ag and Sn, the average ratio of concentrations in campaign II to campaign I resulted of 1.2 in SF (ranging from 0.9 to 2.0) and 3.3 in SSF (ranging from 0.9 to 11.6). Therefore, HM concentrations in the first unit of CWs showed a reduced variation between 2.7 and 4.0 year of operation, while they suffered a strong increase in the second CW unit for the same period of operation.
In general, CW units showed an increase of HM concentration in accumulated solids. This increase agrees well with the net sedimentary characteristics of CW ecosystems, the predominance of precipitation reactions of inorganic compounds and the mineralisation of organic matter in CW sediments. However, an increase of accumulated solids (from approximately 3–7 kg m 2) between 2.7 and 4.7 year of operation in these CW units has been previously reported (Ruiz et al., 2010; de la Varga et al., 2013). This progressive increase in total accumulated solids could help explain the maintenance of the high removal efficiencies of HMs (see Fig. 2) in combination with moderate increases in HM content of accumulated solids (as indicated in Table 3). As a general rule for most of the analysed metals (Zn, Cr, Cu, Pb, Cd and Hg), metal concentration decreased along the treatment path. For these metals, the concentrations in CW sediments also increased with the operation time. Conversely, Fe and Mn increased their concentration along the treatment path, showing the high concentration in the last CW unit. Furthermore, these metals did not show a defined evolution in time, while Ni and As did not show a defined spatial or temporal evolution. Similar HM profiles along the treatment path were reported by Lesage et al. (2007), except for Fe, as this metal in the plant of Lesage et al. (2007) also presented decreasing concentrations along the treatment path. HM concentrations in SF sediments are in the same order or somewhat higher than those reported by Lesage et al. (2007). 4. Conclusions The UASB-CW hybrid system treating a typical municipal wastewater showed a higher HM removal than reported in some previous studies. This behaviour was attributed to the more reduced conditions in the units of the UASB-CW systems due to the anaerobic processes undergone in the UASB and to the application of a high organic SLR in CW units. High removal efficiencies were found for some metals in the following order: 94% > Sn > Cr > Cu > Pb > Zn > Fe > 63%. Medium removal efficiencies were registered for Ni (49%), Hg (42%), and Ag (40%), although the figures for Hg and Ag could be underestimated because of several effluent values below the DL. Finally, Mn and As showed negative percentage removals although only dissolved Mn showed significant differences between influent and effluent concentrations. Removal efficiencies of total HMs were very similar in UASB and SF units, when expressed on the base of the influent to the respective treatment unit and were clearly lower (even negative) in the SSF unit. A correlation between the percentage removal and influent concentration was observed for most HMs in UASB digester but not in SF and SSF CW units. Removal of HMs in a dissolved form took place mainly in the UASB, except for Fe and Mn. The SF CW
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unit also contributed to the removal of dissolved forms of Sn, Cu, Pb and Ni. The removal efficiency of Sn, Cr, Cu, Zn and Ni was maintained along the operation time (from 2.7 to 4.8 year) while a decrease was observed for Pb, Fe, As and Mn. Acknowledgements This research was supported by the Spanish Department of Education and Science (project CTM2008-06676-C05). The authors thank the firm Aquagest for helping in the plant operation and maintenance. References Álvarez, J.A., Armstrong, E., Gómez, M., Soto, M., 2008a. Anaerobic treatment of low strength municipal wastewater by a two-stage pilot plant under psychrophilic conditions. Bioresour. Technol. 99, 7051–7062. Álvarez, J.A., Ruíz, I., Soto, M., 2008b. Anaerobic digesters as a pretreatment for constructed wetlands. Ecol. Eng. 33, 54–67. APHA, 1995. Standard Methods. Standard Methods for the Examination of Water and Wastewater, 19th ed. American Public Health Association, Washington, DC. Arroyo, P., Ansola, G., de Luis, E., 2010. Effectiveness of a full-scale constructed wetland for the removal of metals from domestic wastewater. Water Air Soil Pollut. 210, 473–481. Barros, P., Ruiz, I., Soto, M., 2008. Performance of an anaerobic digester-wetland system for a small community. Ecol. Eng. 33, 142–149. Calijuri, M.L., da Fonseca Santiago, A., Fernandes Moreira Neto, R., de Castro Carvalho, I., 2011. Evaluation of the ability of a natural wetland to remove heavy metals generated by runways and other paved areas from an airport complex in Brazil. Water Air Soil Pollut. 219, 319–327. de la Varga, D., Díaz, M.A., Ruiz, I., Soto, M., 2013. Avoiding clogging in constructed wetlands by using anaerobic digesters as pretreatment. Ecol. Eng. 52, 262–269.
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