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Effects of sulfate on heavy metal release from iron corrosion scales in drinking water distribution system.

Trace heavy metals accumulated in iron corrosion scales within a drinking water distribution system (DWDS) could potentially be released to bulk water and consequently deteriorate the tap water quality. The objective of this study was to identify and evaluate the release of trace heavy metals in DWDS under changing source water conditions. Experimental pipe loops with different iron corrosion scales were set up to simulate the actual DWDS. The effects of sulfate levels on heavy metal release were systemically investigated. Heavy metal releases of Mn, Ni, Cu, Pb, Cr and As could be rapidly triggered by sulfate addition but the releases slowly decreased over time. Heavy metal release was more severe in pipes transporting groundwater (GW) than in pipes transporting surface water (SW). There were strong positive correlations (R2 > 0.8) between the releases of Fe and Mn, Fe and Ni, Fe and Cu, and Fe and Pb. When switching to higher sulfate water, iron corrosion scales in all pipe loops tended to be more stable (especially in pipes transporting GW), with a larger proportion of stable constituents (mainly Fe3O4) and fewer unstable compounds (β-FeOOH, γ-FeOOH, FeCO3 and amorphous iron oxides). The main functional iron reducing bacteria (IRB) communities were favorable for the formation of Fe3O4. The transformation of corrosion scales and the growth of sulfate reducing bacteria (SRB) accounted for the gradually reduced heavy metal release with time. The higher metal release in pipes transporting GW could be due to increased Fe6(OH)12CO3 content under higher sulfate concentrations.


Introduction
The accumulation of heavy metals within drinking water distribution systems (DWDS) has caused considerable concern. Even if heavy metals are present at concentrations below their maximum contaminant levels (MCLs) or at non-detectable levels in treated water, they can accumulate within deposits and corrosion scales that exist in DWDS (Peng et al., 2012;Lytle et al. 2004Lytle et al. , 2010Fisher et al., 2000;Lytle et al., 2004;Schock et al., 2008). Some iron compounds commonly found in corrosion scales, such as goethite (a-FeOOH), lepidocrocite (g-FeOOH), magnetite (Fe 3 O 4 ) and ferrihydrite have strong affinity to adsorb and concentrate trace heavy metals, such as arsenic (As), chromium (Cr), lead (Pb), vanadium (V) and uranium (U) (Peng et al., 2010;Sarin et al., 2001). Relative to iron compounds, manganese (Mn) oxides/oxyhydroxides are commonly found in DWDS, and these are highly effective at adsorbing metal ions, such as Cr, copper (Cu), iron (Fe), nickel (Ni), Pb and strontium (Sr) ions (Gerke et al., 2016).
The primary concern with trace heavy metals in DWDS is the potential release of the metals back into the water (through particulate transport and/or chemical release, e.g. desorption, dissolution), which may result in elevated heavy metal levels in domestic tap water. For example, there was a major release of lead and other contaminants into a midsized Midwestern distribution system following the introduction of chlorination in the United States. Shortly after chlorination was instituted, numerous colored water complaints were received, and subsequent investigations found high levels of Fe, Cu and As in the colored water specimens (Reiber and Dostal, 2000). Lytle et al. (2010) reported As concentration in domestic water was as high as 299 mg L À1 , which was nearly six times the MCL of 50 mg L À1 . Therefore, drinking water providers should be aware of the potential release of trace heavy metals that are pre-accumulated in iron corrosion scales, since these might pose a health risk if these contaminants are released into drinking water.
Alternative water sources and long distance water transport can help solve water shortage problems. However, source water switches can cause undesirable water quality changes due to the disturbance of corrosion scales by changes in water chemistry. Sulfate was identified as one of the most significant water quality parameters in many water chemical stability models and indices such as the Larson index (Larson and Skold, 1958). Under source switch condition, sulfate has been recognized as a causative agent of red water in many cases, and elevated levels of Fe or Mn in bulk water are associated with discolored water (Li et al., 2010;Ginige et al., 2011). For example, in a red water case caused by source water switch in a northern China city, it was found that the new water source had significantly higher sulfate concentration compared with local water source (200 mg L À1 vs 30 mg L À1 ) (Yang et al., 2014), but the effect of sulfate on the deterioration of water quality within DWDS are currently not fully understood.
Moreover, iron release from corrosion scales/deposits in DWDS is usually not considered to be a health concern but rather an aesthetic issue associated with discolored water. Fe and Mn based compounds commonly existing in DWDS corrosion scales are capable of adsorbing and concentrating trace heavy metals. We wonder if heavy metal release accompanied by iron release could result from switching to alternative source water and long distance transported water. The relationship and the interaction mechanisms between Fe and Mn release and heavy metal release are not well understood. The main factors influencing the release behavior of heavy metals are also not known.
This work sought to identify and evaluate the release of trace heavy metals under switching water quality conditions. Iron pipe sections from actual DWDS with different source water supply histories were collected for pipe loop set-up. The relationships of total iron release, trace heavy metal release and corrosion scale constituents under changing sulfate water conditions were systematically analyzed and discussed.

Pipe loop set-up and operation
Test cast-iron pipes (approximately 20 years old) were obtained from two locations in a northern city of China, and labelled as Pipe-A and Pipe-B. Historically, Pipe-A was transporting finished surface water (SW1), and Pipe-B was transporting finished groundwater (GW1). Water sources, main finished water quality parameters and treatment processes of the corresponding water treatment plants are summarized in Table S1. Two pipe loop facilities were set up in parallel, and each facility consisted of five DN100 and 1-m-long test pipe sections and three 0.1-m-long coupon pipe sections (Fig. S1). The pipe loops with Pipe-A and Pipe-B were designated as Loop 1 (L1) and Loop 2 (L2), respectively.
A two-phase experiment was conducted for both pipe loops. Phase I: feed water was regular tap water collected from our lab without further disinfection (with low residual chlorine of 0.05e0.07 mg L À1 ). The sulfate concentration was about 75 mg L À1 , and the duration was 50 d (equilibration period for stabilizing the concentration of total iron and turbidity of effluents). Phase II: sulfate concentration of feed water was increased by adding sodium sulfate to the regular tap water. Three sulfate concentrations were applied for each loop: for L1, 200, 250, and 300 mg L À1 ; for L2, 200, 230, and 250 mg L À1 .
Experimental phases, operation conditions, and corresponding water types for each loop are listed in Table 1. We simulated the actual flow mode of water in the distribution system and the domestic use. A complete operation cycle was set at 48 h, and routine operation cycle procedures were as follows: first, a certain amount of test water was prepared in a water tank; then water circulation was started and continued for 36 h, corresponding to 16 h of hydraulic retention time (HRT) in cast-iron pipes; second, circulation was suspended and the water was kept motionless for 8 h, then samples were taken for measurement; at the end, used water was discharged and preparations for a new cycle were started. The 8 h motionless time represented the night status when there was little or no water use. Water was circulated in test pipe loops using a centrifugal pump at a flow rate of 3.2 m 3 h À1 , and the flow velocity in test pipe sections was approximately 0.1 m s À1 . The water temperature was maintained at 20 ± 1 C using a heat-exchanging unit.

Pipe section reactor set-up and operation
Several small pipe section reactors (Loop 3) were assembled with cast iron pipes, which were harvested from another GW (GW2) service area in a northern city of China (Fig. S2). These pipes (denoted as Pipe-C) were approximately 20 years old, with diameter of DN100. Each test pipe was cut into 0.1-m-long experimental sections, and the cutting faces were coated with epoxy resin to prevent water contact. The HRT in pipe section reactors was 8 h. A stirring device was installed in reactors to simulate water flow conditions in actual DWDS.
Raw tap water from the lab was fed into the pipe section reactors until the concentration of total iron and turbidity of effluents stabilized (phase I). Then the sulfate concentration was increased by adding sodium sulfate to the regular tap water at 200, 300, 400, 500, 600 mg L À1 (phase II), to evaluate sulfate effects on the release of heavy metals from iron corrosion scales. The test used six concentration gradients (from 75 to 600 mg L À1 ), and was replicated five times (ran 1~ran 5).

Analytical methods
Water quality parameters, such as pH, temperature, dissolved oxygen (DO), conductivity, turbidity, alkalinity, chloride (Cl À ) and SO 4 2À were measured for both feed and effluent water. The pH, temperature, DO and conductivity were measured using Portable Multiparameter meter (Sension156 HACH, USA), and the turbidity was measured using a 2100P Turbidimeter (HACH, USA). Free chlorine was measured using a HACH DR2800 spectrophotometer. SO 4 2À and Cl À were analyzed using the Ion Chromatography System-1000 (Dionex, USA). Fe, Mn, Cu, Zn, Pb, Ni, Cr and As were analyzed by Inductively Coupled Plasma Mass Spectrometer (ICP-MS, Perkin-Elmer Optima, 2000, USA). Unless otherwise specified, all chemicals used in this study were Analytical Reagent grade. Ultrapure nitric acid was used to digest samples for metal analysis.

Characterization of iron corrosion scales
Corrosion scale samples were collected from freshly obtained pipe sections and pipe loops L1 and L2 at the end of phase I and phase II, respectively. Typical corrosion scale samples with different appearances were classified and analyzed separately. Pretreatment procedures were as described previously (Yang et al., 2012). X-ray diffraction (XRD, D/max-rA, Rigaku Co., USA) and X-ray Fluorescence (XRF, Advant' XP, Thermo Electron, Switzerland) were performed to characterize the crystalline phase and elemental composition of corrosion scales. The instrumental information for XRD and XRF as well as the quantification method for XRD were the same as described previously (Yang et al., 2012).

Structural and elemental composition of initial corrosion scales
The morphology of corrosion scales on cast iron pipes differed depending on the type of historic source water (Fig. S3), which was described previously (Yang et al., 2012;Sun et al., 2014a). According to Yang et al. (2014), corrosion scales on pipes historically transporting treated SW (Pipe-A) with an appearance of a continuous thick layer (10e20 mm) could be divided into top surface and hard shell (THS), porous core layers (PCL), and entire tubercle (ET). THS and ET scales are at the interface between bulk water and the pipe wall, and were barriers that protected the pipe from further corrosion. Their stability changes have great impact on iron corrosion and metal release. Scales on pipes transporting GW (Pipe-B and Pipe-C) are relatively thin (only few millimeters or less than a millimeter) with non-tubercle-formed corrosion scale (TNCS). The hollow tubercle (HT), scale type which formed in the pipes transporting GW, was not detected in the present study.
A total of 22 inorganic elements (including Fe, Si, Ca, Al, Na, Mg, S, K, Mn, P, Ti, Zn, Cl, Mo, Ni, Cr, V, Sr, Cu, Pt, Pb and As) in DWDS corrosion scales were detected in the corrosion scale samples. Table 2 summarizes the contents (mg g À1 ) of some main elements in the initial corrosion scales. Fe was the largest component in all of the scales analyzed, with mean content of 519640 mg g À1 . The second most abundant element for the two scales of TNCS was Ca (mean ¼ 22100 mg g À1 ), and nearly 6.5 times of the Ca content in THS, PCL and ET (mean ¼ 3424 mg g À1 ). As, Cr, Cu, Mg, Mn, Pb and Zn also had higher levels in TNCS, compared with other scale samples. The contents of S in the THS, PCL and ET scales were greater than those in the two scales of TNCS, and the highest content was present in the PCL scale. S occurrence in scales is likely to be in the form of troilite, a common under-layer component of iron corrosion scales (Benjamin et al., 1996). We did not detect crystalline Fe x S y minerals by XRD, but the XRF results demonstrated that S was common.

Heavy metal release
Inorganic contaminants in DWDS solids came from several sources, but mainly from source water and pipe materials. According to Peng et al. (2012), As, V and U released from mineral deposits dissolution were more prevalent in GW supplies, since they could not be efficiently removed by traditional water treatment plants. Some other metal ions such as Pb, Cu, Cd, Cr and other impurities might originate from pipe materials and pipe joints in DWDS (Masters et al., 2016). We studied the effects of sulfate on seven heavy metals (including As, Cr, Cu, Mn, Ni, Pb and Zn) release using cast iron pipe-loops with different corrosion scales. The changes of heavy metal release under different sulfate concentrations are presented in Fig. 1. Releases of As, Cr, Cu, Mn, Ni and Pb increased immediately at higher sulfate concentration but the releases gradually decreased over time. Compared with L1, the   releases of As, Cu, Mn, Ni and Pb in L2 were greater. In pipe section reactors (L3), heavy metal release exhibited similarity to L1 and L2. The mean concentrations of these heavy metals under different sulfate concentrations are show in Table 3. There was no correlation between Zn release and sulfate concentration in all pipe loops (reactor).
Mn. Mn occurrence in corrosion scales was mainly due to the formation and deposition of Mn oxyhydroxide that originated from the source water (Peng et al., 2010;Swietlik et al., 2011). Under anoxic environments, Mn oxides/oxyhydroxides were prone to reduction and re-solubilization, with associated release of adsorbed substance (e.g. As, Pb, Ni, Cd, Cu, V) . As shown in Fig. 1, higher sulfate levels can cause higher release of Mn. Further correlation analyses (Fig. S4) indicated that Mn release was positively related to sulfate concentration in all pipe loops (R 2 : L1 ¼ 0.72; L2 ¼ 0.59). The mean effluent concentrations of Mn in L1 (154 mg L À1 ), L2 (115 mg L À1 ) and L3 (117 mg L À1 ) all exceeded the drinking water standards of China (100 mg L À1 , GB 5749-2006), when the sulfate concentration increased to 300 mg L À1 , 230 mg L À1 and 300 mg L À1 , respectively (Fig. 1, Table 3). XRF results demonstrated that the contents of Mn in THS, PCL, ET and TNCS (L2) decreased at higher sulfate water levels, and the decrements were 194, 233, 258 and 443 mg L À1 , respectively ( Table 2). The oxygen consumption for L1 and L2 also increased with sulfate, especially for L2 (Fig. S5). For example, the residual DO concentration decreased rapidly from 6.67 to 0.98 mg L À1 in L2, when the pipe loop operated for 8 h (an entire operation cycle was 48 h), which was much lower than that of L1 (3.65 mg L À1 after 26 h operation). Therefore, when switching to higher sulfate water, the anaerobic condition might contribute to the release of Mn in reduced and more soluble form, which needs to be confirmed.
Ni, Cu. Similar to Mn release, elevated levels of Ni and Cu were found in all pipe loops when initially increasing sulfate concentration ( Fig. 1, Table 3). The concentrations of Ni in L2 (22 mg L À1 ) and L3 (23 mg L À1 ) also exceeded the drinking water standards (20 mg L À1 ), when the sulfate concentrations increased to 250 mg L À1 (L2) and 400 mg L À1 (L3), respectively. XRF results showed that the contents of Ni and Cu both decreased, sometimes to undetectable levels, in most of the scales at phase II (Table 2). A strong positive relationship existed between the release of Ni and Mn (R 2 : L1 ¼ 0.95, L2 ¼ 0.90), while relatively weak correlation existed between the release of Ni and sulfate (R 2 : L1 ¼ 0.73, L2 ¼ 0.65) (Fig. S4). Shi and Taylor (2007) found that Cu release was controlled by the dissolution of the corrosion scales, not by the corrosion rate of the metal. Mn-containing deposits could absorb Cu in DWDS, and the destabilization/re-solubilization of Mn deposits might increase the Cu ions in consumer taps (Gerke et al., 2016). Cu release was strongly correlated with Mn release (R 2 : L1 ¼ 0.82, L2 ¼ 0.97), but weakly correlated with sulfate (R 2 : L1 ¼ 0.53, L2 ¼ 0.50) (Fig. S4). Mn compounds can also absorb Ni effectively (Trivedi et al., 2001). Hence, it might indicate that the release of Cu and Ni were mainly accompanied by the release of Mn under higher sulfate condition.
As, Cr, Pb. Under raw tap water condition, the effluent concentrations of As and Cr were very low, and often undetectable. At higher sulfate water levels, the release of As and Cr increased but remained lower than their influent concentrations (Fig. 1, Table 3). As release was positively correlated with Mn release (R 2 : L1 ¼ 0.78, L2 ¼ 0.75), while Cr release was more associated with sulfate levels (R 2 : L1 ¼ 0.86, L2 ¼ 0.84) (Fig. S4). Pb release clearly increased with sulfate level. But under raw tap water condition, Pb release was also lower than its influent concentration. A strong positive correlation existed between the release of Pb and Mn (R 2 : L1 ¼ 0.94, L2 ¼ 0.91), and relative weak relationship existed between Pb release and sulfate level (Fig. S4). In the presence of high concentrations of sulfate water, the contents of As, Cr, and Pb in all analyzed scales decreased ( Table 2). The ability for iron oxides to adsorb As is wellknown, and compounds commonly found in iron corrosion scales (such as a-FeOOH, g-FeOOH, Fe 3 O 4 et al.) have similar adsorptive properties, which could therefore concentrate As and potentially release it into the DWDS (Lytle et al., 2004). Mn oxides/oxhydroxides on pipe internal surfaces or iron corrosion scales also can effectively adsorb Cr, Cu, Pb et al. (Lytle et al., 2004;Gerke et al., 2016;Manceau et al., 1992). Therefore, we suspect that the adsorption of As, Cr and Pb by Fe/Mn based solids within corrosion scales accounts for their relatively lower effluent concentrations. Moreover, Mn oxides could oxidize some adsorbed metal ions, e.g., Cr (III) to Cr (VI), changing the toxicity of the heavy metals (Gerke et al., 2016). Cr (IV) bound to Mn oxides/oxhydroxides and ingested is a potential human health risk.

Iron release and relationships between iron and heavy metal release
Curves of total iron release for L1 and L2 exhibited similar variation during the entire experimental period (Fig. 2). The total iron concentrations increased at the initial stage with increasing sulfate concentration, and then decreased with time (in accord with the results observed in L3, see below). Compared to L1, total iron release in L2 was greatly affected by sulfate concentration. The concentrations of total iron in L2 nearly all exceeded 0.3 mg L À1 (upper limit of the drinking water standard for iron in China) when switching to higher sulfate water, and the effluents always exhibited a yellowish color. The total iron release fluctuation scenarios ("D total iron release" ¼ "total iron release sample measurement" minus "average total iron release in corresponding experimental phase") under different sulfate concentrations for L1 and L2 are illustrated in Fig. S6. The "D total iron release" for L2 changed over a wider range from À0.057e0.076 mg L À1 (phase I) to À0.078e0.069 mg L À1 , -0.120e0.182 mg L À1 and -0.170e0.290 mg L À1 , when the sulfate concentration increased to 200, 230 and 250 mg L À1 , respectively. However, the "D total iron release" for L1 had a narrower range of between À0.053e0.090 mg L À1 during the entire experimental phases. The total iron release fluctuation scenarios under various operation conditions can reflect pipe scale stability (Yang et al., 2014). These results demonstrated that the corrosion scales of GW pipes were more affected by sulfate and the corrosion scale stability of SW pipes was higher than that of GW pipes, since the same water matrix (lab tap water) was used in this study and the change in water quality was only the level of sulfate. Sulfate is often the causative agent of red water, and this is reflected in indices such as the Larson index (Li et al., 2010). Sulfate can increase ferrous iron diffusion from pipe surfaces into bulk water, resulting in increased iron concentrations, but the pathways involved are not well understood. In the present study, we found that increasing sulfate concentration leads to a relatively anaerobic environment in distribution systems (especially for GW pipes). For L2, the DO concentration decreased rapidly from 6.67 to 0.98 mg L À1 when the sulfate concentration increased to 250 mg L À1 . However, the concentration of DO in L1 decreased slowly from 6.67 to 3.65 mg L À1 , even when the sulfate concentration increased to 300 mg L À1 . According to the Kuch mechanism, it is possible for previously deposited ferric scales to act as an electron acceptor and produce ferrous iron to diffuse into water in the absence of oxygen (Eqn. (1)) (Kuch, 1988). Sarin et al. also reported that decreasing the oxidant concentrations, such as DO, in water can increase the amount of iron release from corroded iron pipes (Sarin et al., 2004). Therefore, the anaerobic condition caused by increased sulfate concentrations could be an important factor increasing total iron release.
Total iron release for L3 was similar to L1 and L2 when sulfate concentration increased (Fig. 3). Under a fixed sulfate concentration, total iron release in ran 1 was greater than that in ran 2, and ran 2 was greater than ran 3. The total iron concentrations in ran 3, ran 4, and ran 5 were closer. The reason for this might be the contents of iron oxides, which can be easily released into the water, decreased over time. Sulfate could cause a more protective scale on the pipe surface, or change the corrosion scales from less stable to more stable constituents (Yang et al., 2014;Imran et al., 2006;McNeill and Edwards, 2001). In the present study, the contents of stable iron components, such as a-FeOOH and Fe 3 O 4 all increased over time in the presence of higher sulfate water concentration, and this also could account for the gradual lowering of iron release. The following section provides corrosion scale comparison and discussion.
The total iron release was positively correlated with sulfate both in L1 (R 2 ¼ 0.70) and L2 (R 2 ¼ 0.58) (Fig. 4). Strong positive correlations also existed between the release of Fe & Mn, Fe & Ni, Fe & Cu, and Fe & Pb, with R 2 all being greater than 0.8. Compared to other heavy metals, Cr release was more closely related to sulfate level (R 2 : L1 ¼ 0.86, L2 ¼ 0.84) and As release was correlated to Mn release (R 2 : L1 ¼ 0.78, L2 ¼ 0.75). This suggests that total iron release triggered by sulfate could result in pre-accumulated/ adsorbed heavy metal release back into the drinking water. Therefore, the total iron release not only contributed to discolored water, but was also a health concern with concomitant heavy metal release into DWDS.

Effect of corrosion scale composition transformation on heavy metal release
Corrosion scales with different morphologies were collected from newly obtained pipe sections in L1 -L3 under different experimental phases, and the crystalline phases were characterized by XRD (Fig. 5). For the initial corrosion scale samples (from pipe sections in phase I), a-FeOOH and Fe 3 O 4 were the main constituents of THS and ET. Relatively higher content of unstable iron components, such as g-FeOOH, siderite (FeCO 3 ), green rust (Fe 6 (OH) 12 CO 3 , GR) and amorphous iron oxides existed in PCL. For the two scales of TNCS, amorphous iron oxides were in greatest abundance, but the contents of a-FeOOH were relatively lower compared with THS and ET.
After moving to a higher sulfate water concentration, some notable characteristics of the transformation of scale crystalline composition were observed. First, the relatively abundance of Fe 3 O 4 in all kinds of corrosion scales increased when at higher sulfate levels and the increased percentages were all >10% except PCL (6%).
The relative contents of a-FeOOH increased in all samples by 2%e 4%. Sulfate was deemed to increase iron corrosion, and accelerate the formation of ferrous iron from elemental iron (Eqn. (2)). Ferrous iron can be oxidized by DO to form a-FeOOH and Fe 3 O 4 under low oxidation rates (Eqn. (3)e(5)), according to the Siderite model proposed by Sontheimer et al. (1981).
Equations (3) Roh et al., 2003;Sun et al., 2014b). In the present study, the bacterial communities in scale samples were analyzed by 16S rRNA pyrosequencing (Supplementary materials A). Results showed that the relative abundances of IRB in all corrosion scales also increased when switching to higher sulfate water (Fig. S7, Table S3). The high abundance of IRB in scales could favor the formation and preservation of Fe 3 O 4 . Fe 3 O 4 is thermodynamically stable, which could produce a dense and compact corrosion scale. Corrosion scales with high Fe 3 O 4 content usually possesses protective and anti-corrosive properties, which serve as barriers hindering the diffusion of corrosive ions into corrosion scales (Sarin et al., 2001(Sarin et al., , 2004Lytle et al., 2005). Fe 3 O 4 is also an effective sorbent for some heavy metals such as Cu, Pb, Cr, As (Rajput et al., 2016;Yantasee et al., 2007;Hu et al., 2004). Therefore, Fe 3 O 4 could act as a barrier to hinder the release of heavy metals on the one hand, and also to act as a sink to adsorb released heavy metals. The formation of Fe 3 O 4 might be one of the most important factors for inhibiting the release of total iron and heavy metals. Second, the percentages of some unstable ferric or ferrous iron oxides (including b-FeOOH, g-FeOOH and FeCO 3 and amorphous iron oxides) in THS, ET, and two samples of TNCS all decreased when switching to higher sulfate water. The decreased percentages were all >10%. b-FeOOH, g-FeOOH, FeCO 3 and amorphous iron oxides were chemically unstable, and the relatively higher percentage of these in corrosion scales indicated the weak tolerance of distribution pipes to feed water chemistry changes (Yang et al., 2014). Therefore, the decrease of these unstable ferric or ferrous iron oxides means that the iron corrosion scales in pipe loops tend to be more stable, which could also account for the decreased release of heavy metals over time. Third, we found that when switching to higher sulfate water, the relative abundance of GR in THS decreased from 15% to 8%, while the contents of GR in the two samples of TNCS both increased (L2: from 16% to 26%, L3: from 16% to 28%). GR was chemically unstable, and could be oxidized by dissolved oxygen into Fe 3 O 4 , a-FeOOH, g-FeOOH and g-Fe 2 O 3 , depending on pH and oxidation rate (Pineau et al., 2008). Under anaerobic condition (as in L2 at phase II), GR can remain for a long time (Refait et al.,3003). The higher abundance of these unstable ferrous oxides could lead to more serious heavy metal release in DWDS. This might explain why the release of total iron and some heavy metals were more serious in L2 than those in L1. GR and sulfate reducing bacterial (SRB) always coexisted, due to the presence of GR, and this could favor the colonization of SRB (Refait et al., 2003). Bacterial community analyses showed that the relatively abundance of SRB in all corrosion scales increased under higher sulfate concentration, especially in the two samples of TNCS (Fig. S7, Table S3). SRB were usually associated with anaerobic iron corrosion, which could favor sulfide production and thus promote iron corrosion and ferrous iron formation (Lytle et al., 2005;Pineau et al., 2008). However, SRB could use sulfate as an electron acceptor and generate sulfide (S 2À ) in an anaerobic environment, the biogenically produced S 2À could react with dissolved metals to form metal sulfide precipitates, such as Cu, Zn, Ni, Fe, Al, As and Mn (Jong and Parry, 2003). Therefore, the growth of SRB could also decrease heavy metal release.

Conclusions
Changes in water chemistry can result in accumulated/adsorbed trace heavy metals in iron corrosion scales being released back into domestic water supplies thus leading to reduced water quality. The present work systematically investigated the effect of sulfate on the release of trace heavy metals from iron corrosion scales to bulk water.
The release of Mn, Ni, Cu, Pb, Cr and As from iron corrosion scales increased immediately after sulfate concentration increased and then gradually decreased over time. Pipes historically transporting GW had greater heavy metal release, compared with pipes historically transporting SW. Significant correlations (R 2 > 0. For the pipes historically transporting SW, corrosion scales tended to be more stable in the presence of higher sulfate levels, with a larger proportion of stable constituents (mainly Fe 3 O 4 ) and fewer unstable compounds (b-FeOOH, g-FeOOH, FeCO 3 and amorphous iron oxides). The main functional bacterial communities IRB were favorable for the formation of Fe 3 O 4 . For pipes historically transporting GW, constituents of corrosion scales also achieved more stability, while the anaerobic condition promoted the formation of GR, a main precursor of stable iron constituents. The higher metal release in pipes transporting GW can be partially attributed to the higher GR content. The relative abundance of SRB increased with sulfate concentration in all kinds of scales. The transformation of corrosion scales to a more stable state and the growth of SRB may explain the gradual decrease of heavy metal release over time.