Science of the Total Environment 472 (2014) 1145–1151

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Evidence of arsenic release promoted by disinfection by-products within drinking-water distribution systems Syam S. Andra a,b, Konstantinos C. Makris a,⁎, George Botsaris c, Pantelis Charisiadis a, Harris Kalyvas a, Costas N. Costa d a Water and Health Laboratory, Cyprus International Institute for Environmental and Public Health in association with Harvard School of Public Health, Cyprus University of Technology, Limassol, Cyprus b Harvard–Cyprus Program, Department of Environmental Health, Harvard School of Public Health, Boston, MA 02115, United States c Department of Agricultural Sciences, Biotechnology and Food Science, Cyprus University of Technology, Limassol, Cyprus d Department of Environmental Science and Technology, Cyprus University of Technology, Limassol, Cyprus

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Pipe scales act as sources of episodic release of contaminants to tap water. • For the first time, we report the effect of pre-formed disinfection by-products on the stability of arsenic-laden pipe scales. • Trihalomethanes and haloacetic acids act synergistically in promoting arsenic release.

a r t i c l e

i n f o

Article history: Received 14 August 2013 Received in revised form 8 November 2013 Accepted 8 November 2013 Available online 22 December 2013 Keywords: Arsenic Biofilm Disinfection by-products Pipe scales Drinking water distribution system Water and health

a b s t r a c t Changes in disinfectant type could trigger a cascade of reactions releasing pipe-anchored metals/metalloids into finished water. However, the effect of pre-formed disinfection by-products on the release of sorbed contaminants (arsenic-As in particular) from drinking water distribution system pipe scales remains unexplored. A bench-scale study using a factorial experimental design was performed to evaluate the independent and interaction effects of trihalomethanes (TTHM) and haloacetic acids (HAA) on arsenic (As) release from either scales-only or scalebiofilm conglomerates (SBC) both anchored on asbestos/cement pipe coupons. A model biofilm (Pseudomonas aeruginosa) was allowed to grow on select pipe coupons prior experimentation. Either TTHM or HAA individual dosing did not promote As release from either scales only or SBC, detecting b 6 μg As L−1 in finished water. In the case of scales-only coupons, the combination of the highest spike level of TTHM and HAA significantly (p b 0.001) increased dissolved and total As concentrations to levels up to 16 and 95 μg L−1, respectively. Similar treatments in the presence of biofilm (SBC) resulted in significant (p b 0.001) increase in dissolved and total recoverable As up to 20 and 47 μg L−1, respectively, exceeding the regulatory As limit. Whether or not, our

⁎ Corresponding author at: Water and Health Laboratory, Cyprus International Institute for Environmental and Public Health in Association with Harvard School of Public Health, Cyprus University of Technology, Irenes 95, Limassol 3041, Cyprus. Tel.: +357 25002398; fax: +357 25002676. E-mail address: [email protected] (K.C. Makris). 0048-9697/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.scitotenv.2013.11.045

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laboratory-based results truly represent mechanisms operating in disinfected finished water in pipe networks remains to be investigated in the field. © 2013 Elsevier B.V. All rights reserved.

1. Introduction Fifty percent of the waterborne disease outbreaks in the United States during 1995–2002 were ascribed to compromises in the integrity of finished tap water within urban drinking-water distribution systems (UDWDS) (WHO/NSF, 2003). Illustrative examples of finished water quality deficiencies within pipe networks of UDWDS were reviewed by Makris et al (2013). Released inorganic contaminants into finished water, for example lead (Pb), may pose a serious health risk to consumers via oral ingestion of lead-contaminated tap water, passing into the systemic circulation of sensitive subpopulations, i.e., children (Edwards et al., 2009; Schock et al., 2008). Other toxic inorganic contaminants, like arsenic (As) have been included in discussions within the European Commission to consider its monitoring within UDWDS (Jorgensen et al., 2008). The common sources of As in drinking water supplies are either natural deposits or anthropogenic activities (such as agriculture and industry) (U.S. EPA, 2010). Historical human exposures to low-level As contamination (b 50 µg As L−1) of drinking-water supplies were recently documented in a region of Cyprus. This resulted in switching of water source from the contaminated well to As-free surface water by the Cypriot authorities (Makris et al., 2012). Though treated water entering UDWDS typically may have As below 10 μg L−1, it tends to accumulate on the inner surfaces of UDWDS pipes (U.S. EPA, 2007). Indeed, the presence of As in pipe scales and corrosion products has been documented (Gerke et al., 2009; Schock et al., 2008), suggesting its widespread prevalence within UDWDS of U.S. cities (Lytle et al., 2004, 2010; Reiber and Dostal, 2000). Arsenic has been found in various types of pipe scales, including iron corrosion products (Lytle et al., 2004, 2010) and lead scales (Gerke et al., 2009; Schock et al., 2008), reaching levels as high as 14 mg As g−1 pipe scale (Lytle et al., 2010). Arsenic tenaciously held in pipe scales or corrosion products may be subject to sporadic release in As-free water, if and only if, drastic changes in water chemistry occur. Relevant examples of significance are: (i) changes in chloride and sulfate concentrations via blending of desalinated with conventionally-treated surface water, resulting in varying extent of iron release from pipe scales (Liu et al., 2013), (ii) changes in water pH along with chloride, sulfate, and orthophosphate concentrations affecting lead release from pipe scales (Cartier et al., 2012, 2013), and (iii) switch of disinfectant from chlorine to monochloramine resulting in elevated Pb concentrations in tap water (Switzer et al., 2006). For metals, like Pb and Cu, literature has already documented destabilization and release phenomena from pipe scales/corrosion products into finished water after a disinfectant change (Boyd et al., 2008; Huggins, 2008; Kim et al., 2011; Lin and Valentine, 2008). With the exception of As release after disinfection of well-water (Walker and Newman, 2011), literature is void of reports linking As release with changes in finished water chemistry, such as, the formation of disinfection by-products (DBP). Earlier work using the Plackett–Burman (PBD) and the central composite (CCD) experimental designs provided hints on candidate water chemistry variables promoting As release from pipe scales and solids (Andra and Makris, 2012). Using the PBD experimental setup, ten water quality chemical variables were screened for their significance on As release, measured as either total recoverable or dissolved As concentrations. The range of tested parameters was: pH, alkalinity, chloride, free chlorine, sulfate, orthophosphate, trihalomethanes (TTHM), haloacetic acids (HAA), hydroxylamine, and nitrosamines (Andra and Makris, 2012). Four significant variables as identified in the PBD experiment (TTHM, HAA, sulfate, orthophosphate) were further included in a separate experimental study design, i.e., that of a central composite design (CCD) (data not shown). Hints, but not conclusive evidence of the TTHM influence on promoting As release from pipe coupons using the CCD

experiments led us in performing a more detailed factorial experiment testing both TTHM and HAA effects on As release by pipe coupons. The objective of this study was to evaluate the independent and interaction effects of TTHM and HAA on As release from simulated pipe scales made of synthesized Fe–As precipitate laden onto asbestos–cement (A/C) pipe coupons in the presence and absence of a model microbe typically detected in a UDWDS biofilm community (Pseudomonas species) (Wang et al., 2012). Pseudomonas aeruginosa was selected for the purpose of this study because it is able to persist in a drinking-water for extended periods and it is often difficult to eliminate from water systems (Hardalo and Edberg, 1997). The organism can grow in very low nutrient aqueous environments and can survive for many months in water at ambient temperatures (Norton and LeChevallier, 2000). It is also an important opportunistic pathogen and is particularly significant as a cause of nosocomial infections (Hardalo and Edberg, 1997). Several studies report the isolation of this organism from drinking water distribution systems, while its persistence in drinking-water is attributed to its easy biofilm colonization (Hardalo and Edberg, 1997; Elhariry et al., 2012). 2. Materials and methods 2.1. Preparation of asbestos–cement pipe coupons with arsenic-laden pipe scales and biofilm conglomerates 2.1.1. A/C pipe coupons Previously-leaking A/C pipe segments that were disconnected from UDWDS in Nicosia, Cyprus were used in our experiments. Select A/C pipe coupons [~4(l) × 3(w) × 1.5(h) cm] were prepared adhering to safety precautions associated with handling asbestos materials. The A/C pipe coupons were used as working-substrates for anchoring synthesized Fe–As precipitates (scales only) and for culturing a model P. aeruginosa biofilm community (scales and biofilm conglomerate, SBC). 2.1.2. Fe–As pipe scales Amorphous ferric arsenate, which represents an environmentallyrelevant As form typically observed in pipe scales of UDWDS (Hill et al., 2010; Lytle et al., 2010), was synthesized following a modified co-precipitation method (Harvey et al., 2006) using sodium arsenate dibasic heptahydrate (Sigma) and iron(III) chloride hydrated (Fisher Chemical) at 1:5 mM ratio. Total As content in our environmentallyrelevant Fe–As precipitate (~5000 mg kg−1) was intentionally chosen to fall below the average total As content encountered in U.S. fieldcollected pipe scales and solids (Lytle et al., 2010). A ~0.5 g slurry made of synthesized Fe–As precipitate was uniformly applied to each A/C coupon. The smeared A/C coupons were allowed to dry for 24 h in an incubator at 50 °C to yield evenly distributed amorphous As-laden pipe scales on each coupon surface, simulating those observed in UDWDS (Picture SI-1), taking precautions such as low temperature and baking time to inhibit the formation of a more crystalline Fe–As scale, based on our earlier observations (unpublished data). Total As content in pipe scales was determined with an inductively coupled plasma mass spectrometer (ICP-MS) (Thermo X Series II, Thermo Scientific, Germany) after the U.S. EPA 3050 acid digestion method. 2.1.3. Scales-biofilm conglomerates (SBC) Pipe coupons were immersed in half nutrient broth (200 mL) and were inoculated with 1 mL of 108 overnight culture of P. aeruginosa. Coupons were then incubated at 30 °C for 48 h to allow biofilm formation. A/C coupons were then rinsed with deionized water and taken

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through the experimental procedure described later. Viability of grown P. aeruginosa biofilm on SBC coupons was assessed by taking a swab sample from the pipe surface (2 cm2) and diluted into 10 mL maximum recovery diluent (Merck KGaA, Germany). The presence of P. aeruginosa was molecularly confirmed by polymerase chain reaction (PCR) using primers PA SS-F and PA SS-R (Spilker et al., 2004). Amplification was carried out in a Bio-Rad DNA Engine TM thermal cycler. PCR products were viewed on a 2% agarose gel stained with SYBR red under a UV transilluminator (Text SI-1).

(90–105%) of trihalomethanes in spiked samples after 24-h reaction indicated absence of HAA to THM degradation reactions. Four treatments were tested by varying additions (0–100 μg L−1) of: (i) TTHM alone, (ii) HAA alone, and (iii–iv) TTHM at a single addition of HAA at 10 μg L−1 or 60 μg L−1. The spiked concentrations represent the sum of four THM species and nine HAA species. Current regulatory limits of TTHM at 100 μg L−1 (EU) and HAA5 at 60 μg L−1 (USA) were chosen as the upper level of spiked dosing. These spiked DBP concentrations were calculated on top of the existing background concentrations of DBP in tap water (Table SI-1).

2.2. Effect of trihalomethanes and haloacetic acids on arsenic release

2.2.2. Chemical and statistical analyses Beakers of 250 mL each, containing the A/C pipe coupon were filled up with 200 mL tap water containing proper amounts of TTHM and/or HAA, leaving minimal headspace for possible DBP losses. Each beaker was covered with parafilm to further minimize air–water exchange and right after spiking with proper amounts of TTHM and/or HAA. They were allowed to equilibrate for 24-h on an incubation-shaker at 100 revolutions per minute (rpm) at room temperature (23 °C). Water samples were safely preserved for less than a month prior analyses without losing their integrity, based on 3-mo sample stability tests. Following the 24-h period, water samples were collected and filtered using 0.45 μm filters and analyzed for dissolved As, while a portion of unfiltered water sample was acidified, left overnight and analyzed for total recoverable As content; both total and dissolved As concentrations were measured using an inductively coupled plasma mass spectrometer (Thermo X Series II, Thermo Scientific, Germany). The particulate As fraction was operationally defined by subtracting dissolved As from

2.2.1. Experimental design A factorial experiment was set-up using spiked additions of trihalomethanes (TTHM) and haloacetic acids (HAA) to As-laden scales (Table 1), while keeping the rest of tap water quality parameters constant as presented in Table SI-1. The stock TTHM solution was composed of chloroform, bromodichloromethane, dibromochloromethane and bromoform at a concentration of 2000 μg mL−1 (EPA 501 Trihalomethanes mix, product # RK30211, Restek Corporation, USA). The stock HAA solution was composed of bromoacetic acid, bromochloroacetic acid, bromodichloroacetic acid, chloroacetic acid, dibromochloroacetic acid, dibromoacetic acid, dichloroacetic acid, tribromoacetic acid, trichloroacetic acid at a concentration of 2000 μg mL−1 (EPA 552.2 Haloacetic Acids Mix, product # 49107-U, Sigma-Aldrich, USA). Working stock solutions of TTHM and HAA at 200 μg L−1 concentration were daily prepared in deionized water and used freshly upon preparation. Stable recovery

Table 1 Total recoverable As, dissolved As and soluble total Fe released into water from either scales-alone or scales-biofilm conglomerates (SBC) for varying spiked combinations of either TTHM, or HAA, or both. Means labeled with different letters were significantly different at p b 0.05 using Tukey's HSD test. Treatment (n = 2 replicates) Scales alone TTHM only

HAA only

TTHM with HAA @ 10 μg L−1

TTHM with HAA @ 60 μg L−1

SBC TTHM only

HAA only

TTHM with HAA @ 10 μg L−1

TTHM with HAA @ 60 μg L−1

DBP additions

Total recoverable As (μg L−1)

Dissolved As (μg L−1)

Soluble Fe (μg L−1)

Control (TTHM @ 0 μg L−1) TTHM @ 1 μg L−1 TTHM @ 10 μg L−1 TTHM @ 100 μg L−1 Control (HAA @ 0 μg L−1) HAA @ 1 μg L−1 HAA @ 10 μg L HAA @ 100 μg L Control (TTHM @ 0 μg L−1) TTHM @ 1 μg L−1 TTHM @ 10 μg L−1 TTHM @ 100 μg L−1 Control (TTHM @ 0 μg L−1) TTHM @ 1 μg L−1 TTHM @ 10 μg L−1 TTHM @ 100 μg L−1

0.79 4.05 4.75 6.46 0.79 2.48 3.02 5.75 4.99 8.89 8.20 23.7 9.47 19.9 32.7 94.7

± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±

0.10D 2.35D 0.80D 0.46D 0.10D 1.44D 1.86D 0.01D 0.66D 1.51D 1.17D 1.29BC 1.89D 5.59C 1.22B 7.63A

0.08 0.91 1.44 3.31 0.08 2.14 1.23 2.28 1.63 1.58 4.04 8.84 2.02 5.18 10.2 16.0

± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±

0.05E 0.16DE 0.04DE 1.16DE 0.05E 0.69DE 0.33DE 0.83DE 0.07DE 0.84DE 0.40CDE 3.00BC 0.19DE 0.33CD 2.90B 2.22A

130 115 135 220 130 145 190 430 195 240 385 450 315 390 565 780

± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±

14H 7H 21H 14FG 14H 7GH 28FGH 14C 21FGH 14EF 21CD 28C 7DE 28CD 49B 14A

Control (TTHM @ 0 μg L−1) TTHM @ 1 μg L−1 TTHM @ 10 μg L−1 TTHM @ 100 μg L−1 Control (HAA @ 0 μg L−1) HAA @ 1 μg L−1 HAA @ 10 μg L−1 HAA @ 100 μg L−1 Control (TTHM @ 0 μg L−1) TTHM @ 1 μg L−1 TTHM @ 10 μg L−1 TTHM @ 100 μg L−1 Control (TTHM @ 0 μg L−1) TTHM @ 1 μg L−1 TTHM @ 10 μg L−1 TTHM @ 100 μg L−1

1.45 0.91 2.40 4.00 1.45 1.39 1.15 1.15 4.59 6.59 8.41 14.6 6.28 8.60 17.7 46.5

± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±

0.64e 0.19e 0.61de 0.22de 0.64e 0.63e 0.47e 0.21e 0.72de 0.88de 0.45cd 5.04bc 1.98de 3.52cd 2.40bc 0.96a

0.09 0.14 1.54 0.70 0.09 0.42 0.42 0.93 1.57 0.83 5.46 15.3 1.67 5.28 11.8 20.1

± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±

0.11e 0.01e 0.01e 0.02e 0.11e 0.10e 0.08e 0.11e 0.10e 0.14e 0.01d 0.69b 0.22e 0.20d 2.39c 2.14a

170 175 215 210 170 185 230 255 235 235 345 400 280 380 415 610

± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±

14h 21gh 7efgh 28fgh 14h 7gh 14efg 21ef 21efg 21efg 7cd 14bc 14de 14bc 21b 14a

Mean comparisons were made between the treatments either within scales-only or SBC groups for each analyte of interest. Treatment means with different alphabets differ significantly in comparison with those in the four groups within each category viz., scales-only or SBC (p b 0.05). Graphical representation of effect of TTHM and HAA on total and dissolved As concentrations released into water from the two sets of A/C pipe coupons was presented as Figure SI-2 (Supplementary Information).

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the total recoverable As content. A standard reference material NIST 1643e (trace elements in water) was used as an external quality control. The limit of detection for As was 10 ng L−1. Linearity of the calibration curve ranged from 1 to 100 μg L−1, while the R2 was always N0.999. The recovery of the NIST standard reference material when spiked in the water was 98-100% (Table SI-2). Spectrophotometric determination of total iron (sum of Fe2+ and Fe3+) and ferrous component (Fe2+) was made on filtered water samples (no digestion) using spectrophotometric method #220 (PC Spectro II, Spectro Direct, Lovibond, Germany). A slightly modified EPA Method 551.1-1 was used to measure the THM in water using a gas chromatography coupled with a triple quadruple mass spectrometer (Agilent 7890A/7000B GC–MS/MS) (Agilent Technologies, Germany). In brief, a liquid–liquid extraction was adopted by mixing 15 mL water sample (spiked with surrogate solution at a final concentration of 10 μg L−1) with 2 mL of t-butyl methyl ether, adding 6.0 g of sodium sulfate to saturate the water sample and shaking gently for 5 min in a lab shaker at 100 rpm. Half of an mL of the organic phase was transferred in a GC autosampler vial, containing internal standard solution at final concentration of 10 μg L−1. Chromatograms from selected treatments (Figure SI-1) and details of the analytical protocol (Text SI1) were presented in the Supporting Information. Limits of detection for chloroform, bromodichloromethane, dibromochloromethane and bromoform were 0.13, 0.11, 0.13, and 0.11 μg L−1, respectively. Linearity of the calibration curve was extended from 0.77 to 99 μg L−1, while the R2 was always N 0.9999. Deionized water and bottled water were used as blanks, because they had no THM. Recoveries of THM in spiked blanks were in the acceptable range of 80–120%. The complexity of HAA measurements in water, coupled with the lack of in house analytical protocol did not allow us to measure them in this study. 3. Results and discussion 3.1. Effect of spiked disinfection by-products on arsenic release No adjustment of water pH in our experimental units was needed, since minor pH drifts were observed at the completion of the equilibration period (no more than 0.2 pH units) (final pH ranged between 7.8 and 8.1). Control treatments containing A/C coupons laden with As scales or SBC and in contact for 24-h with tap water media yielded an average total recoverable As concentration of b2 μg L−1 and dissolved As levels below limit of quantification (b0.05 μg L−1), suggesting that background tap water constituents like residual chlorine and/or nutrients exerted little influence on As release. The initial volume of organic solvents (methanol or t-butyl methyl ether) in the stock DBP solution (b10 μL) that was used to make working standard solutions was diluted to insignificant levels after preparation of working solution and further diluted upon preparation of the 200 mL final tap water volume. The nature of the factorial design allowed for the elimination of confounding effects by keeping most water chemistry variables constant except for TTHM and/or HAA. Additions of either TTHM or HAA only, up to 100 and 60 μg L−1 spike each, respectively, resulted in small, and non-significant (p N 0.05) increase in total and dissolved As concentrations (up to 6 μg L−1) for both scales-only and SBC coupon treatments (Figure SI-2). In the case of scales-only coupons, the combination of highest spike level of TTHM and HAA significantly (p b 0.001) increased dissolved As and total As concentrations to levels up to 16 and 95 μg L−1, respectively, in comparison with 0.1 and 0.8 μg As L−1 levels in control treatments (multiple comparisons of treatment means across the four groups using Tukey's HSD test (Tukey, 1953)) (Table 1). Similarly, in the presence of biofilm (SBC) the same treatments resulted in significant (p b 0.001) increase in dissolved and total recoverable As up to 20 and 47 μg L−1, respectively, in comparison with 0.1 and 1.5 μg As L−1 levels in control treatments (Table 1). Repeating ANOVAs using the particulate As expression (total minus dissolved As concentrations) did not change the significance or the direction of effects (data not shown). Indeed, suspended particulates were visually observed

(Picture SI-1F), and corroborated with the higher As release in the TTHM and HAA combination treatments. Addition of both TTHM and HAA yielded total recoverable As concentrations in finished water in the range of 1 to 100 μg L−1, and 1 to 50 μg L−1 for scales-only and SBC treatments, respectively, while nearly equal in magnitude release of As was documented for dissolved As for both scales-only and SBC treatments (up to 20 μg L−1) (Table 1). The presence of biofilm (SBC) on pipe scales did neither enhanced nor lowered dissolved As fraction in solutions (p N 0.05) in comparison to corresponding scales-only treatments (Fig. 1). However, in the presence of 60 μg L−1 HAA, an addition of 10 and 100 μg L−1 TTHM resulted in significantly higher levels of total recoverable As in scales-only treatments compared with corresponding SBC treatments (33 vs. 18 μg L−1 and 95 vs. 47 μg L−1, respectively) (Fig. 1). A particulate fraction of Asladen scale released into finished water contributed to the magnitude and variability of total recoverable As concentrations measured in water after spiking with the highest THM and HAA dose as visual fragments clouding the solution were visually observed (picture in supplementary information). 3.2. Possible role of disinfection by-products on arsenic release It was speculated that As release may be governed by a couple of environmental processes: the incongruent dissolution of scorodite (ferric arsenate scale) releasing ferric oxy(hydroxides) and HAsO2− at 4 a pH around 7.5 (Harvey et al., 2006), which was close to the average water pH of our experiments (pH 8.0). The relatively low ionic potential of iron permits the hydrolysis of ferric solids to form sparingly soluble Fe hydroxide species in water. In parallel, the acid dissociation constant (pKa) of acetic acid, major functional group of HAA in water is 4.76, suggesting full deprotonation of HAA at the measured pH (8.0) of our experiments. Hence, it was speculated that ferric-acetate soluble complexes may form at the expense of the ferric (oxy) hydroxide species being at equilibrium with the Fe solid (pipe) scale. Such formation of Fe–HAA complexes would require the replenishment of soluble Fe species via additional dissolution of both iron and As species from the pipe scale surface. Indeed, measured total recoverable As concentrations in finished water appeared in proportion with increasing soluble total Fe concentrations (Table 1). A significant (p b 0.001) correlation was observed between total recoverable As and total Fe concentrations. Similar trends were reported earlier: (i) significant correlation between water As and Fe concentrations (r = 0.96) and color (r = 0.89) (Lytle et al., 2010), and (ii) significant positive association between As and Fe release from drinking-water pipe solids (Copeland et al., 2007). These observations may as well suggest similarities between the observed As release from pipe scales here and that shown with pipe corrosion products (Lytle et al., 2010). Furthermore, the mixed composition of Fe surface charges (both +2 and +3) has been documented for several Fe minerals due to corrosion or reaction with natural organic matter (Elsner et al., 2004a,b; Sarin et al., 2003). The speculated presence of ferrous ions on ferric arsenate pipe scale could be involved with the sorption and reductive dehalogenation of TTHM, such as trichloronitromethane, trichloroacetonitrile, and trichloroacetaldehyde (Chun et al., 2005), and there by releasing sorbed As into water. Chlorinated and brominated DBP were susceptible to reduction by zerovalent Fe and a combination of iron oxide minerals, such as Fe(II)/goethite, and Fe(II)/magnetite (Chun et al., 2005, 2007; Lee et al., 2007; Zhang et al., 2004). Trichloronitromethane, an analog of halomethanes (TTHM) was degraded in the presence of ferrous iron via a reduction–oxidation step (Lee et al., 2008), suggesting that TTHM transformation could occur for the ferrous ion-containing ferricarsenate pipe scale. Measured TTHM concentrations in the end of the 24-h equilibration period were always a fraction of the initial spiked dose (Table 1). Measured residual TTHM concentrations in the runs containing both TTHM and HAA were negatively (p = 0.03 and p = 0.02) correlated with released As and Fe levels in water, respectively. With

1000

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175

900 800 700 600 500 400 300 200

25

150

100

Total THM at Time-Zero (µg L-1)

Soluble Total Fe (µg L-1)

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15

10

5

60

100 0

,H A A -6 00

,H A A -1 00

H M -1

H M -1

H M -1

TT

TT

80

TT

0

,H A A -0 00

,H A A -6

,H A A -0

H M -0

H M -0 TT

TT

40

0

0

Spiked DBP Treatment (µg L-1)

60 40

20

20 0

-6 A

A

A

A

,H

,H

00

00

-1

-1

M H TT

TT

H

M

M H TT

0

0 -1

-0 A A ,H 00 -1

M H TT

TT

H

M

-0

-0

,H

,H

A

A

A

A

-6

-0

0

0

Total THM at Time-Final (µg L-1)

80

Dissolved As (µ g L -1)

Total Recoverable As (µg L-1)

20

Spiked DBP Treatment (µg L-1) Scales Only-Final TTHM SBC-Final TTHM Initial TTHM Scales Only-Total Recoverable As SBC-Total Recoverable As Scales Only-Soluble Total Fe SBC-Soluble Total Fe

Fig. 1. Effect of spiked DBP (TTHM/HAA on total arsenic, residual TTHM, and soluble iron concentrations measured in water (μg L−1)) from two sets of A/C pipe coupons — with synthesized Fe–As pipe scales only (gray colored shapes represent data from A/C pipe coupons–with scales and absence of Pseudomonas aeruginosa) and with scales-biofilm conglomerates (SBC) (no-fill shapes represent data from A/C coupons with scales and Pseudomonas aeruginosa). Changes in dissolved arsenic concentrations were shown in the inset figure for the same treatments. Changes in trihalomethane levels were presented in Table SI-3 (Supplementary Information). TTHM and HAA denote trihalomethanes and haloacetic acids, respectively.

increasing Fe release in water (Table 1), a corresponding increase in % loss of added TTHM in water with As release was observed (Fig. 1 and Table SI-3), suggesting TTHM participation in reactions impacting particulate As release. Iron(II) concentrations (140 μg L−1) in time final of the experiments were detected only in one of the highest TTHM/HAA (100 μg L−1/10 μg L−1) treatments, but this may be due to the low sensitivity of the colorimetric test we used (method detection limit of 100 μg L−1). Only a small portion of TTHM losses could be explained by open atmosphere losses, because the absence of systematic volatilization trends across treatments suggested otherwise. Visual appearance of particulates in the beakers containing the highest spike of both TTHM and HAA was observed (Picture SI-1F), suggesting a role of particulate matter in the very high total recoverable As concentrations released from the pipe coupons. A few study limitations are noted here: Tap water was purposely used to simulate field conditions rather using a simpler matrix (deionized water), though the latter may have permitted for better control of experimental conditions. The question whether DBP promotes As release from pipe coupons still remains to be answered beyond our

speculated mechanistic reactions. The trend indicates occurrence of Fe(III) reduction and DBP oxidation, though it is unclear how conversion of Fe(III) to Fe(II) coupled with DBP reactions could result in increased As release. Ferric-ligand complexation reactions with HAA are plausible; however the preliminary data presents little supporting data. Acetate as a weak ligand may not have a stand-alone effect on iron dissolution and associated As release as observed by adding HAA alone in the range of 0 to 100 μg L−1. The need for the co-occurrence of TTHM an HAA at high concentrations for mobilizing As from pipe SBC is unclear. In addition, water pH remained noticeably unchanged (~pH 7.9 at time-zero) upon the addition of DBP solutions and at time-final that rules out effect of pH in the study treatments. Microbial activity could perhaps play a role in TTHM reduction and As release, but this was not the main focus of this study. 4. Implications for drinking-water distribution systems This work reported for the first time the concomitant influence of pre-formed TTHM and HAA on As release from pipe scales and SBC

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within UDWDS. As a result, our study's results are of preliminary nature and should be treated with caution. Whether or not, our laboratorybased results truly represent mechanisms operating in actual disinfected UDWDS remains to be investigated in more detail. Such episodic release events could be observed in district-metered areas within a single UDWDS that favor the formation of high TTHM and/or HAA concentrations, which in turn, could be implicated with the release of sorbed contaminants by pipe scales, like As or Pb. Co-occurrence of TTHM and HAA in the same water sample is frequently documented (Malliarou et al., 2005). Our recent work in UDWDS highlighted the wide spatial variability of DBP even within the same city, where certain neighborhoods encountered much higher THM concentrations than others, despite the same water source, treatment and chlorination (unpublished data from our lab). Residual chlorine levels within an aged UDWDS may not be homogeneously distributed across all network points, driving a similar distribution of formed DBP levels. The wide spatial variability of formed DBP even within the same city supplied with the same raw water source and disinfection treatment could prevent us from detecting episodic metal release events with current monitoring schemes that lack appropriate spatial resolution. Unless spatially-resolved mapping of areas with high risk for formation of TTHM and HAA is undertaken, such inorganic contaminant release events will go unnoticed. Our work has practical implications for all low-level As-contaminated areas using chlorine disinfection in their potable water; assuming historic accumulation of low level As on the internal surfaces of pipe scales and/or SBC, conformation of As-impacted authorities to the revised MCL of 10 μg As L−1 may have resulted in a distinct As concentration gradient at the pipe surfaceflowing water interface. If no changes in the UDWDS water quality are foreseen, historically dormant As will remain tenaciously sorbed by pipe scales. However, as our work suggested, chemisorbed As on pipes could react with formed TTHM and/or HAA upon excessive chlorination in summer, disrupting the pseudo-equilibrium As state, and, thus, initiating destabilization and release phenomena. Our study's TTHM and HAA concentrations were realistic, albeit in the upper range of their probability distribution curves, showcasing a worst-case scenario. Occurrence of high DBP levels in finished water has been reported (TTHM N 100 μg L− 1 by Lynberg et al., 2001; Miles et al., 2002; Nuckols et al., 2005; and HAA5 N 60 μg L− 1 and HAA9 N 100 μg L− 1 by Weinberg et al., 2002). The levels of HAA9 in UK drinking-water ranged between 35 and 95 μg L− 1 with a maximum concentration of 244 μg L− 1, while TTHM and HAA often coexisted in the same water sample (Malliarou et al., 2005). Even in a conservative scenario with lower HAA concentrations (e.g. 10 μg L −1 ), As release exceeded its current 10 μg L− 1 regulatory limit in the presence of TTHM at 100 μg L− 1. Hence, it is suggested that each UDWDS authority should identify those areas within their territory where TTHM/HAA concentrations fall above the 90th percentile of the seasonal quartile distribution plots. Those areas would require special attention, especially in the summer when the combination of high temperatures and higher dosing of disinfectant would exacerbate the magnitude of DBP formation. In addition to the elevated health risk posed by TTHM/HAA themselves, authorities would have to monitor possible leaching of sorbed metals by pipe scales into finished water, whenever increased DBP formation is observed in the respective district metered areas. Inevitably, because of the preliminary nature of our findings, more work at the pilot-, and field-scale is needed to validate and confirm our laboratory batch results. The role of biofilms and their interactions with pipe scales/corrosion products needs to be elucidated with respect to the stability of sorbed metals. The case of As was used here, but the observed trends could be also extrapolated to other sorbed trace elements within UDWDS. These findings intend to promote additional research to dispel the notion that conservative behavior of accumulated contaminants in UDWDS is an exception, rather a rule.

Conflict of interest The authors declare no conflict of interest. Acknowledgments We thank the funding agency supporting this work (EU FP7 Marie Curie # PIRG05-GA-2009-249271) to Konstantinos C. Makris. Special credit goes to Nicosia Water Board for permitting the use of A/C pipe segments and the kind contribution of Mr. Andreas Iliofotou in preparing A/C pipe coupons. Appendix A. Supplementary data Text: (i) Procedural details on biofilm molecular-confirmation on the A/C pipe coupons, and (ii) analytical method details for trihalomethane analysis in water samples using GC–MS/MS. Tables: (i) Water quality parameters of Limassol (Cyprus) tap water, and (ii) analytical procedural quality control for the ICP-MS analytical measurements. (iii) Concentrations and percent change in spiked trihalomethane concentrations in water samples at time-final across the study treatments. Figures: (i) Pictorial representation of GC–MS/MS chromatograms of trihalomethanes in water samples at time-final, from the selected treatments. (ii) Effect of controlled addition of TTHM and HAA on As release to water from pipe scales only and SBC A/C coupons in a factorial design experimental setup. Picture: Visualization of experimental pipe coupons and experimentation. Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.scitotenv. 2013.11.045. References Andra SS, Makris KC. Arsenic destabilization from synthesized ferric arsenate pipe solids found in asbestos–cement water distribution system. Oral Presentation at Proceedings at the Protection and Restoration of the Environment XI. Eleventh International Conference 2012; Thessaloniki, Greece; 2012. Boyd GR, Dewis KM, Korshin GV, Reiber SH, Schock MR, Sandvig AM, et al. Effects of changing disinfectants on lead and copper release: part 1 — literature review. J Am Water Works Assoc 2008;100:1–44. Cartier C, Nour S, Richer B, Deshommes E, Prévost M. Impact of water treatment on the contribution of faucets to dissolved and particulate lead release at the tap. Water Res 2012;1546(16):5205–16. Cartier C, Doré E, Laroche L, Nour S, Edwards M, Prévost M. Impact of treatment on Pb release from full and partially replaced harvested Lead Service Lines (LSLs). Water Res 2013;147(2):661–71. Chun CL, Hozalski RM, Arnold WA. Degradation of drinking water disinfection byproducts by synthetic goethite and magnetite. Environ Sci Technol 2005;39:8525–32. Chun CL, Hozalski RM, Arnold WA. Degradation of disinfection byproducts by carbonate green rust. Environ Sci Technol 2007;41:1615–21. Copeland RC, Lytle DA, Dionysiou DD. Desorption of arsenic from drinking water distribution system solids. Environ Monit Assess 2007;127:523–35. Edwards M, Triantafyllidou S, Best D. Elevated blood lead in young children due to lead-contaminated drinking water: Washington DC 2001–2004. Environ Sci Technol 2009;43:1618–23. Elhariry H, Gherbawy Y, El-Deeb B, Altalhi A. Molecular identification and biofilm-forming ability of culturable aquatic bacteria in microbial biofilms formed in drinking water distribution networks. Geomicrobiol J 2012;29(6):561–9. Elsner M, Schwarzenbach RP, Haderlein SB. Reactivity of Fe(II)-bearing minerals toward reductive transformation of organic contaminants. Environ Sci Technol 2004a;38:799–807. Elsner M, Haderlein SB, Kellerhals T, Luzi S, Zwank L, Angst W, et al. Mechanisms and products of surface-mediated reductive dehalogenation of carbon tetrachloride by Fe(II) on goethite. Environ Sci Technol 2004b;38:2058–66. Gerke TL, Scheckel KG, Schock MR. Identification and distribution of vanadinite (Pb5(V5+O4)3Cl) in lead pipe corrosion by-products. Environ Sci Technol 2009;43: 4412–8. Hardalo C, Edberg SC. Pseudomonas aeruginosa: assessment of risk from drinking water. Crit Rev Microbiol 1997;23:47–75. Harvey MC, Schreiber ME, Rimstidt JD, Griffith MM. Scorodite dissolution kinetics: implications for arsenic release. Environ Sci Technol 2006;40:6709–14. Hill AS, Friedman MJ, Reiber SH, Korshin GV, Valentine RL. Behavior of trace inorganic contaminants in drinking water distribution systems. J Am Water Works Assoc 2010;102:107–18. Huggins D. Remediation of lead levels in drinking water: the city of London's experience. Ontario Water Works Association Proceedings of the OWWA/OMWA Joint Annual Conference and Trade Show 2008; London, Ontario, Canada; 2008.

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