Degradation of Drinking Water Disinfection Byproducts by Synthetic

Sep 21, 2005 - Impacts of bacteria and corrosion on removal of natural organic matter and disinfection byproducts in different drinking water distribu...
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Environ. Sci. Technol. 2005, 39, 8525-8532

Degradation of Drinking Water Disinfection Byproducts by Synthetic Goethite and Magnetite CHAN LAN CHUN, RAYMOND M. HOZALSKI, AND WILLIAM A. ARNOLD* Department of Civil Engineering, University of Minnesota, 500 Pillsbury Drive SE, Minneapolis, Minnesota 55455-0116

Corrosion of iron pipes leads to the release of ferrous iron, Fe(II), and the formation of iron oxides, such as goethite and magnetite, on the pipe surface. Fe(II), a potent reductant when associated with iron oxide surfaces, can mediate the reduction of halogenated organic compounds. Batch experiments were performed to investigate the kinetics and pathways of the degradation of selected chlorinated disinfection byproducts (DBPs) by Fe(II) in the presence of synthetic goethite and magnetite. Trichloronitromethane was degraded via reduction, while trichloroacetonitrile, 1,1,1-trichloropropanone, and trichloroacetaldyde hydrate were transformed via both hydrolysis and reduction. Chloroform and trichloroacetic acid were unreactive. Observed pseudo-first-order reductive dehalogenation rates were influenced by DBP chemical structure and identity of the reductant. Fe(II) bound to iron minerals had greater reactivity than either aqueous Fe(II) or structural Fe(II) present in magnetite. For DBPs of structure Cl3C-R, reductive dehalogenation rate constants normalized by the surface density of Fe(II) on both goethite and magnetite correlated with the electronegativity of the -R group and with one electron reduction potential. In addition to chemical transformation, sorption onto the iron oxide minerals was also an important loss process for 1,1,1-trichloropropanone.

Introduction Drinking water distribution system pipes are composed of cast iron, ductile iron, cement-lined ductile iron, steel, copper, poly(vinyl chloride), or fiberglass. Iron and steel pipes have been used commonly in drinking water distribution systems (1). According to a 1996 survey, 30.2% of the 370 000 miles of U.S. water mains consist of unlined cast/ductile iron or steel (2). Typically, the interior pipe surface is exposed to an oxidizing environment causing the buildup of corrosion products over time. The quantity and composition of corrosion deposits vary from system to system and within individual systems due to differences in hydraulic conditions, pipe age, and water quality (1, 3-5). Nevertheless, regarding corrosion product composition, some consistency is apparent. In iron and steel pipe from Champaign, IL (1), and Melbourne, Australia (4), goethite (R-FeOOH) was the predominant iron mineral in corrosion deposits, followed by magnetite (Fe3O4), and lepidocrocite (γ-FeOOH). Green * Corresponding author phone: (612)625-8582, fax: (612)626-7750; e-mail: [email protected]. 10.1021/es051044g CCC: $30.25 Published on Web 09/21/2005

 2005 American Chemical Society

rust was observed inside tubercles collected from pipes in Columbus, OH (3). The behavior of disinfection byproducts (DBPs) in distribution systems has drawn significant attention due to the suspected health effects of these compounds. Both DBP formation and degradation have been observed in full-scale distribution systems (6-8). DBPs are often small, halogenated organic compounds that may degrade via hydrolysis, biological transformation, and abiotic reduction to form a variety of products. DBPs may also sorb to surfaces within the pipe. Many studies have reported the hydrolysis of DBPs. Basecatalyzed hydrolysis of trihaloketones forms trihalomethanes (9-11), and haloacetonitriles are hydrolyzed to haloacetoamides and haloacetic acids (HAAs) (12, 13). Biodegradation of halogenated DBPs proceeds via two mechanisms: hydrolysis-oxidation (14-16) and reductive dechlorination (17). Compounds that are amenable to hydrolysis-oxidation, such as HAAs, are expected to be more labile because distribution systems tend to be aerobic. Losses of HAAs observed in distribution systems (18) and in aquifer storage recovery systems (19) are typically attributed to biodegradation. Although halogenated DBPs such as HAAs are susceptible to reduction by Fe(0) (20, 21), little is known about the reduction of DBPs by iron corrosion products. Given that most iron water mains have been in place for decades and are corroded, reduction by and adsorption to iron minerals are potential loss processes for DBPs. Degradation of disinfectants such as monochloramine by iron oxides (22) suggests that oxidized organic species may also be reduced by iron minerals. Also, ferrous iron produced via corrosion (5, 23) may sorb onto iron minerals enhancing the reactivity (24-30). This research focuses on the abiotic degradation of regulated DBPs (chloroform, TCM; trichloroacetic acid, TCAA) and emerging DBPs (1,1,1-trichloropropanone, 1,1,1TCP; trichloroacetaldehdye hydrate, TCAh; trichloroacetonitrile, TCAN; and trichloronitromethane, TCNM) in the presence of aqueous Fe(II), magnetite, and Fe(II) sorbed onto goethite or magnetite.

Experimental Section Additional details for the following sections are provided in the Supporting Information (SI). Chemicals. Chemicals, suppliers, and purities are listed in the SI. Iron Minerals. Goethite and magnetite were selected due to their common occurrence as pipe corrosion products. X-ray diffraction patterns (Cu KR radiation, Siemens Bruker AXS D5005) were used to confirm synthesized mineral identities. The BET surface areas (ASAP 2000) of freeze-dried goethite and magnetite were 7.49 and 87.00 m2/g, respectively. Batch Experiments. Batch experiments were carried out in 123 mL serum bottles containing a 0.8 g/L mineral suspension buffered at pH 7.5 with deoxygenated 25 mM morpholinopropanesulfonic acid. Ferrous sulfate was added to the reactors to achieve a total Fe(II) concentration of 1 mM. Analytical Methods. A 0.5 mL filtered sample was extracted with 1.0 mL of n-pentane or MTBE for the analysis of the parent DBPs and degradation products (except for TCAA and nitromethane, NM) using gas chromatography with electron capture detection (Trace GC, ThermoQuest). Capillary electrophoresis was used for analysis of TCAA (20). For NM analysis, the method described by ref 31 was used. Data Analysis. The overall, hydrolysis, and reductive dehalogenation pseudo-first-order rate constants of DBPs VOL. 39, NO. 21, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 1. Degradation of TCNM (b) and DCNM (4) by Fe(II)/goethite (a, c) and Fe(II)/magnetite (b, d). DCNM is an intermediate in TCNM degradation, and the final product is NM (0) for both compounds. The horizontal mark (-) represents the observed carbon mass balance, and solid lines are model fits of parent compound disappearance. All experiments were performed at pH 7.5 with 1 mM total Fe(II) and a mineral loading of 0.8 g/L. in each system were calculated using Scientists for Windows (v. 2.01, Micromath Research). Calculation of One-Electron Reduction Potentials. Oneelectron reduction potentials for the dehalogenation half reactions (E1H) were determined via free energies computed using computational chemistry software.

Results DBP Reaction Pathways. The reaction of TCNM with aqueous Fe(II), Fe(II) bound to goethite (Fe(II)/goethite, Figure 1a) or magnetite (Fe(II)/magnetite, Figure 1b), and magnetite alone resulted in the formation of dichloronitromethane (DCNM) and NM. DCNM was also reduced to NM in the Fe(II)/goethite (Figure 1c) and Fe(II)/magnetite systems (Figure 1d). Chloronitromethane (CNM) was detected as another intermediate product. CNM was not quantified because of interference from a contamination peak, which resulted in incomplete carbon mass recovery at intermediate times. Degradation of CNM was verified by observing the formation of NM and Cl- within 3 h in experiments where CNM was used as starting material (data not shown). NM was also degraded, likely to methylamine, which was not analyzed for, but at a slower rate than TCNM, DCNM, and CNM in the Fe(II)/goethite and Fe(II)/magnetite systems (t1/2 > 5 days, data not shown). No degradation of TCNM was observed over 150 h in the presence of goethite without Fe(II) (data not shown). The degradation of TCAN resulted in the simultaneous formation of two major products, dichloroacetonitrile (DCAN) and trichloroacetamide (TCAM), in the presence of all 8526

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reductants as shown in Figure 2. TCAN was degraded to DCAN via hydrogenolysis. No further reduction of DCAN was observed over 150 h. TCAM is the hydrolysis product and was the only product in the presence of goethite alone (data not shown). The hydrolysis rate of TCAN increased with increasing pH (see SI), indicating that the hydrolysis of TCAN was base-catalyzed. TCAM was the major hydrolysis product at circum-neutral pH, but TCAM was further hydrolyzed to TCAA at high pH. This result is consistent with the findings of Glezer et al. (12). The overall carbon mass balance ranged from 84% to 106%. Rapid loss of 1,1,1-TCP from the aqueous phase was observed within 30 min without the formation of products (data not shown), suggesting sorption of 1,1,1-TCP to the iron oxide minerals. Sorption isotherms for 1,1,1-TCP onto goethite and magnetite are presented in the SI. Linear isotherms gave apparent iron mineral-water distribution coefficients (Kd) for goethite (0.22 ( 0.028 L/g, r2 ) 0.95) and magnetite (0.12 ( 0.004 L/g, r2 ) 0.99). The transformation of 1,1,1-TCP as a function of time in the Fe(II)/goethite and Fe(II)/magnetite systems is shown in Figure 3. The concentrations of 1,1,1-TCP shown are total concentrations including aqueous and sorbed 1,1,1-TCP, which was determined by extracting unfiltered samples (mineral + water). 1,1,1-TCP predominantly hydrolyzed to TCM in both systems. The hydrogenolysis product, 1,1dichloropropanone (1,1-DCP), was only observed with Fe(II)/magnetite. It is possible that Fe(II)/goethite may reduce 1,1,1-TCP, but because no reduction products were observed (due to rapid hydrolysis compared to dechlorination), a

FIGURE 2. Degradation of TCAN (b) in the presence of: (a) Fe(II)/goethite, (b) Fe(II)/magnetite, (c) Fe(II), and (d) magnetite. Major products are DCAN (4) and TCAM (0). The horizontal mark (-) represents the observed carbon mass balance, and the solid lines are model fits of parent compound disappearance and product formation. All experiments were conducted at pH 7.5 with 1 mM total Fe(II). The mineral loading was 0.8 g/L where applicable.

FIGURE 3. Degradation of 1,1,1-TCP (b) by (a) Fe(II)/goethite and (b) Fe(II)/magnetite. Major products are 1,1-DCP (4) and TCM (0). The horizontal mark (-) represents the observed carbon mass balance, and the solid lines are model fits of parent compound disappearance and product formation. All experiments were at conducted at pH 7.5 with 1 mM total Fe(II) and a mineral loading of 0.8 g/L. dechlorination rate constant could not be determined. With Fe(II)/goethite, 15% of the 1,1,1-TCP was adsorbed and the remaining 85% underwent hydrolysis. With Fe(II)/magnetite, sorption of 1,1,1-TCP accounted for 9% of the loss, the contribution of reduction was 10%, and hydrolysis accounted for 81%. No reduction was observed with magnetite alone. The degradation of TCAh and the simultaneous formation of two major products, TCM (via hydrolysis) and dichloroacetaldehyde (DCAh; via hydrogenolysis), in the presence of Fe(II)/goethite and Fe(II)/magnetite are shown in Figure 4. Over 350 h, only 35% and 10% of the TCAh was degraded by Fe(II)/goethite and Fe(II)/magnetite, respectively. For Fe(II)/goethite, the total carbon mass balance averaged 90%, and for Fe(II)/magnetite it averaged 100%. In the Fe(II) and buffer-only systems, only the TCM was observed. No loss of TCM or TCAA was observed over 150 h in the presence of iron minerals, aqueous Fe(II), or both.

Reaction Kinetics. Calculated reaction rate constants are summarized in Table 1. The overall rate constants for DBPs exhibited the following trend with respect to the reductant used: Fe(II) bound to goethite or magnetite > aqueous Fe(II) > magnetite.

Discussion Sorption of 1,1,1-TCP. To the best of our knowledge, this work presents the first report of 1,1,1-TCP sorption to iron oxide surfaces. 1,1,1-TCP sorbed more strongly onto goethite than magnetite, even though magnetite had a larger surface area. One possible explanation is that the point of zero charge of goethite is greater than that of magnetite. At pH 7.5, goethite has a more positively charged surface, potentially leading to stronger interaction with partial negative charges in the 1,1,1TCP molecules. TCAA, 1,1-DCP, and TCAh, which have similar chemical structures to 1,1,1-TCP, did not sorb to either VOL. 39, NO. 21, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 4. Degradation of TCAh (b) by (a) Fe(II)/goethite and (b) Fe(II)/magnetite. Major products are DCAh (4) and TCM (0). The horizontal mark (-) represents the observed carbon mass balance, and the solid lines are model fits of parent compound disappearance and product formation. All experiments were conducted at pH 7.5 with 1 mM total Fe(II) and a mineral loading of 0.8 g/L.

TABLE 1. Reaction Rate Constants for DBPs at pH 7.5 DBP TCNM

DCNM

TCAN

1,1,1-TCP

TCAh

e

system

C0a (µM)

kob (h-1)

khc (h-1)

krd (h-1)

red.e (%)

Fe(II)/goethite Fe(II)/magnetite Fe(II) magnetite Fe(II)/goethite Fe(II)/magnetite Fe(II) magnetite Fe(II)/goethite Fe(II)/magnetite Fe(II) magnetite goethite Fe(II)/goethite Fe(II)/magnetite Fe(II) magnetite goethite Fe(II)/goethite Fe(II)/magnetite Fe(II) magnetite goethite

71.1 ( 1.2f 27.9 ( 1.2 85.7 ( 1.1 88.9 ( 1.5 51.2 ( 1.1 43.3 ( 1.1 61.8 ( 1.0 56.4 ( 1.0 82.2 ( 2.1 86.6 ( 3.0 88.1 ( 3.1 79.3 ( 4.3 87.8 ( 3.2 17.3 ( 0.9 18.2 ( 0.5 18.6 ( 0.8 18.3 ( 0.6 18.9 ( 0.7 103.7 ( 2.8 100.5 ( 2.0 79.3 ( 2.8 82.9 ( 1.3 127 ( 1.2

7.67 ( 0.44 8.05 ( 0.67 3.63 ( 0.23 0.05 ( 0.01 1.13 ( 0.06 1.22 ( 0.09 0.17 ( 0.01 0.02 ( 0.002 3.7((0.3) × 10-2 2.5((0.3) × 10-2 2.4((0.2) × 10-2 1.5((0.3) × 10-2 1.4((0.1) × 10-2 1.0((0.2) × 10-2 9.5((0.9) × 10-3 8.3((0.8) × 10-3 8.0((0.9) × 10-3 6.5((0.8) × 10-3 1.8((0.4) × 10-3 4.5((1.8) × 10-4 2.4((0.7) × 10-4 1.9((0.4) × 10-4 2.1((0.2) × 10-4

n/hg n/h n/h n/h n/h n/h n/h n/h 5.1((0.9) × 10-3 1.4((0.1) × 10-2 1.5((0.1) × 10-2 1.3((0.2) × 10-2 1.4((0.1) × 10-2 1.0((0.2) × 10-2 8.6((0.5) × 10-3 8.3((0.8) × 10-3 8.0((0.9) × 10-3 6.5((0.8) × 10-3 1.9((1.4) × 10-4 3.2((0.9) × 10-4 2.4((0.7) × 10-4 1.9((0.4) × 10-4 2.1((0.2) × 10-4

7.67 ( 0.44 8.05 ( 0.67 3.63 ( 0.23 0.05 ( 0.01 1.13 ( 0.06 1.22 ( 0.09 0.17 ( 0.01 0.02 ( 0.002 3.2((0.2) × 10-2 1.1((0.1) × 10-2 8.2((1.0) × 10-3 2.1 ( (1.3) × 10-3 n/rh n/r 0.9((0.4) × 10-3 n/r n/r n/r 5.7((1.4) × 10-4 1.3((0.9) × 10-4 n/r n/r n/r

100 100 100 100 100 100 100 100 86.2 42.4 34.7 13.6 n/r n/r 9.5 n/r n/r n/r 31.5 28.9 n/r n/r n/r

a Initial concentration. b Overall pseudo-first-order reaction rate constant. c Hydrolysis rate constant. d Reductive dechlorination rate constant. Contribution of reductive dechlorination to overall reaction. f Errors represent 95% confidence limits. g No hydrolysis. h No reduction.

goethite or magnetite. This suggests there is a unique aspect in the (electronic) structure or physical-chemical properties (e.g., hydrophobicity) of 1,1,1-TCP that facilitates the sorption to goethite and magnetite. Reaction Pathways. The identity of the fourth substituent on the compounds with CCl3-R structure plays a role in the reductive pathways. Sequential hydrogenolysis appears to be dominant in the reduction of TCAN and 1,1,1-TCP. The appearance of DCNM as TCNM is degraded indicates that hydrogenolysis occurs. For TCNM, however, another pathway besides sequential hydrogenolysis may be important. In the TCNM experiments, a carbon deficit appears soon after initiating the experiment and increases in magnitude up to intermediate reaction times. Sorptive losses of TCNM, DCNM, and NM were negligible. Thus, the incomplete carbon mass balances were attributed to our inability to quantify CNM. If TCNM degradation occurred primarily via sequential hydrogenolysis, then the deficit should have appeared later, after significant concentrations of DCNM were formed and subsequently degraded to CNM. It appears that DCNM and CNM were formed simultaneously in these systems. Toward the end of the experiments, carbon mass balances improve 8528

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as NM accumulates. Experiments using CNM as a starting material resulted in the formation of NM and chloride (data not shown). Past studies have shown that reduction of TCNM results in simultaneous formation of DCNM and CNM in Fe(II) bearing systems (32, 33). The process may be the result of R-elimination or another pathway leading to rapid removal of two chlorine atoms. With iron metal, parallel pathways of hydrogenolysis (61%) and reductive R-elimination (39%) have been observed for TCNM (with DCNM and methylamine (MA) as major products (31)). Reductive R-elimination has also been observed for other structurally related compounds such as 1,1,1-trichloroethane (34) and carbon tetrachloride (30, 35, 36) with a variety of reductants. For the reaction of TCAh with Fe(II)/magnetite, it appears that the major products have been detected (TCM and DCAh). With Fe(II)/goethite, TCM and DCAh do not account for all of the TCAh loss. Further degradation of TCM and DCAh is unlikely. Both were unreactive when used as starting materials. Sorptive losses of TCAh, TCM, and DCAh were also negligible. The incomplete carbon balance may be due to the formation of an additional product. Trichloroethanol has been observed as a reduction product in biological systems

TABLE 2. Relationship between Normalized Reductive Dechlorination Rate Constants and Group Electronegativity (Xg) of the -R Group or One-Electron Standard Reduction Potential (E1H) of DBPs of Structure Cl3C-R Fe(II)/goethite surface area: 7.49 m2/g surface density of [Fe(II)]ads: 0.010 mmol/m2 R group -NO2 -CN -COCH3 -CHO -COOH -H CHCl2-NO2g

kra (h-1)

kr,sb (h-1 m-2 L)

7.67 3.2 × 10-2 -d 5.7 × 10-4

0.819 3.4 × 10-3

1.13

0.121

6.1 × 10-5

Fe(II)/magnetite surface area: 87.00 m2/g surface density of [Fe(II)]ads: 0.002 mmol/m2

kr,sdc (h-1 mmol-1 m2) 744.15 3.1 5.5 × 10-2 110.08

kr (h-1)

kr,s (h-1 m-2 L)

kr,sd (h-1 mmol-1 m2)

8.05 1.1 × 10-2 9.0 × 10-4 1.3 × 10-4

0.074 9.7 × 10-5 1.0 × 10-5 1.2 × 10-6

3890.82 5.1 0.4 6.3 × 10-2

1.22

0.011

588.99

Xg (44)

E1H (V)

3.63 3.10 2.84 2.85 2.73 2.20 n/ah

-0.274,e 0.647f 0.454 0.300 -0.116 0.128 -0.159 n/a

Pseudo-first-order reductive dechlorination rate constant as determined by Scientist for Windows (h-1). b Reductive dechlorination rate constant normalized by surface area (h-1 m-2 L). c Reductive dechlorination rate constant normalized by surface density of adsorbed Fe(II) (h-1 mmol-1 m2). d No reductive dechlorination observed. e Value determined using computational chemistry techniques. f Value determined from the linear regression relating Xg and E1H of the other five DBPs. g Data for this compound are relevant to the discussion of the normalization of the rate constants, but not the correlations with E1H or Xg. h Not available. a

(37). Further study is needed to identify any other reaction products. Kinetics of DBP Reductive Dechlorination. The reductive dechlorination rate constant (kr) of the DBPs depends on the chemical structure of the DBP as well as the reductant identity (Table 1). Comparing the reductants, the reactivity of Fe(II) associated with iron oxide minerals was greater than either structural or aqueous Fe(II). Higher reactivity of surfacebound Fe(II) for DBP reduction was expected based on numerous studies, which have noted that surface-bound Fe(II) accelerated reduction rates of various organic compounds (24-30) and heavy metals (38-40). For reactions mediated by surface-associated Fe(II), the concentration of reactive sites and hence the pseudo-firstorder rate constant is affected by the mineral surface area, the extent of Fe(II) sorption, the presence of structural Fe(II), and Fe(II) surface speciation (29, 41, 42). To compare dechlorination rates for Fe(II) bound to different minerals, a suitable normalization scheme is needed. Elsner et al. (29) normalized rate constants with respect to surface area and amount of sorbed Fe(II), and then suggested normalization of rate constants by surface density of sorbed Fe(II) as a practical approach to estimate the effect of reactive site concentration. Klupinski et al. (41) also found better correlation between rate constants and surface density of sorbed Fe(II) as compared to a relationship that accounted for Fe(II) surface speciation. To compare the reactivity of goethite and magnetite in the degradation of DBPs, reductive dehalogenation rate constants were normalized by mineral surface area and the surface density of sorbed Fe(II) (Table 2). Although it may be useful to compare reaction rate constants with surfacebound Fe(II) at different pH values, experiments in this work were performed at a single pH. Thus, normalization by surface complex concentrations is similar to normalization by surface density of sorbed Fe(II) because only a single distribution of surface complex concentrations exists at a single pH. It could be argued that the contribution of aqueous Fe(II) to overall reductive dechlorination should be subtracted. Given the recently elucidated and as of yet incompletely understood role of aqueous Fe(II) in reductions mediated by iron oxides (as shown in ref 42), we have chosen to normalize total reductive dechlorination rate constants. The surface area-normalized reductive dehalogenation rate constants (kr,s) for degradation of TCNM, DCNM, TCAN, and TCAh by Fe(II)/goethite were greater than the respective constants obtained for degradation by Fe(II)/magnetite (Table 2). Nevertheless, the trend was not consistent with the observed reaction behavior of 1,1,1-TCP. The reduction

of 1,1,1-TCP in the presence of Fe(II)/goethite was negligible in comparison to the hydrolysis rate, but reduction was observed in the presence of Fe(II)/magnetite. This suggests that normalizing by surface area alone does not properly account for the number of reactive sites. The rate constants normalized by the surface density of sorbed Fe(II) (kr,sd) for the Fe(II)/magnetite system, however, were consistently greater than those for the Fe(II)/goethite system (Table 2). Normalization of reduction rate constants by the surface density of sorbed Fe(II) appears to be the most useful approach for comparing the reduction kinetics for DBPs at the same pH. After normalization by the surface density of sorbed Fe(II), the following reactivity trend results: TCNM > TCAN > 1,1,1,-TCP ≈ TCAh . TCAA ≈ TCM. These compounds have a common structure Cl3C-R, where -R is -NO2, -CN, -C(d O)CH3, -CHO, -COO-, or -H. Group electronegativity (Xg) has been used as a parameter to develop quantitative structure activity relationships (43, 44). The potential for correlating normalized dechlorination rate constants with the group electronegativity of the -R groups or the E1H of the DBPs was investigated. The Xg values (44) are listed in Table 2. The relationship between log kr,sd and Xg is plotted in Figure 5a. Other Xg values are available (45, 46) and provide similarly good correlation. The more reactive compounds have higher (more electronegative) Xg values. This suggests that this group may pull electron density away from the -CCl3 moiety, making the dechlorination reaction more facile. Figure 5a shows that for a given Fe(II)/mineral system, predictions about whether a compound of structure Cl3C-R will be reductively dechlorinated may be made based on a minimum Xg value (∼2.8). Reduction potentials are often used to develop linear free energy relationships. Values are not readily available for all of the DBPs of interest, necessitating their calculation via computational means. Values determined in this work are listed in Table 2, and the relationship between log kr,sd and E1H is plotted in Figure 5b. In testing such a relationship, it is assumed that the first electron transfer is the rate-limiting step. Note that the E1H value for TCNM determined computationally (-0.274 V) does not fit with the trends in Xg or log kr,sd. Additional calculations with nitro-substituted species suggest that the method used does not accurately predict energies for small nitro compounds. Thus, an E1H was extrapolated on the basis of the relationship between E1H and Xg for the other DBPs (see SI). In Figure 5a and b, a single regression captures both the Fe(II)/goethite and the Fe(II)/magnetite data. For Xg, the relationship is log kr,sd ) 5.31((1.23)Xg - 16.0((3.9), and for VOL. 39, NO. 21, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 5. (a) Correlation of dechlorination rate constants normalized by surface density of sorbed Fe(II) plotted with group electronegativity of the -R groups (Xg; 44). Open symbols are rate constants with Fe(II)/goethite, and solid symbols are rate constants with Fe(II)/magnetite. (b) Correlation of dechlorination rate constants normalized by surface density of sorbed Fe(II) plotted with one electron reduction potential (E1H). Open symbols are rate constants with Fe(II)/goethite, and solid symbols are rate constants with Fe(II)/magnetite. E1H, the relationship is log kr,sd ) 5.29((2.94)E1H - 1.03((1.30). Reported errors are 95% confidence limits. Regression of the data for the individual reductants does not give slope or intercept values that are statistically different from those above, but the associated errors are larger given that the individual regressions contain fewer data points. The fact that a single relationship predicts the reactivity of two minerals using rate constants normalized by the surface density of sorbed Fe(II) suggests that the form and/or reactivity of the reactive sites on goethite and magnetite in the presence of aqueous Fe(II) are similar for the DBPs studied. Testing with additional compounds is necessary to fully develop the predictive relationships (especially given that the shape of the log kr,sd vs E1H plot could be curvilinear) and to gain additional insight into the nature of the reactive sites. Although the correlation with Xg is informative, it is limited to compounds with the structure Cl3C-R. The correlation with E1H is universal and could prove more useful because both the halogenated end of the molecule and the -R group could be varied. Such a correlation could be extended to include monohalogenated and dihalogenated compounds including those containing other halogens other than chlorine. 8530

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The occurrence of reduction enhances the degradation rate as compared to hydrolysis alone and may affect overall toxicity of DBP mixtures in drinking water. As transformation proceeds from TCNM (reduction-dominant) to DCNM and CNM, genotoxicity may be reduced but chronic cytotoxicities are similar (47). Unfortunately, we are unaware of any reports in which the toxicities of emerging DBPs and all of the potential degradation products were tested in a single study. In the case of hydrolysis-dominant DBPs such as 1,1,1-TCP and TCAN, hydrolysis leads to formation of TCM and TCAA, respectively, which are more toxic than the parent compounds (48). The experiments reported herein were performed in the presence of synthetic iron minerals at pH 7.5 for up to 15 days in the absence of competing oxidants (i.e., dissolved oxygen (DO) or chlorine). In typical iron water distribution pipes, multiple amorphous and crystalline forms of iron are present (4, 5, 49). Experiments with single minerals were performed to simplify data analysis and interpretation and to ascertain the importance of different minerals in the degradation and sorption of DBPs. The minerals were selected on the basis of their widespread occurrence in water distribution systems (1, 3-5, 49). The results obtained provide the basis for our ongoing DBP degradation experiments using iron corrosion products collected from water distribution systems. In distribution systems, typical conditions include: pH of 6.5-9; hydraulic retention time of 1-15 days; and oxidizing conditions due to the presence of DO and chlorine (1, 5, 23, 50). As pH increases, (base-catalyzed) hydrolysis should increase in importance relative to reduction. The effect of pH on DBP reduction rates is complicated because changes in pH can alter both the extent of Fe(II) adsorption as well as the speciation of the surface-associated Fe(II) (22, 24-28, 39, 41). To extrapolate the findings to the full range of conditions expected in water distribution systems, experiments are needed to evaluate the effects of pH and the competition for reductive capacity from oxygen and chlorine. Fe(II) concentrations in the bulk water in drinking water distribution systems are expected to be negligible due to the presence of numerous potential oxidants. Close to the pipe wall or in stagnant regions of the distribution system, significant Fe(II) concentrations are possible given that Fe(II) is released during the corrosion of iron metal and Fe(II) is the primary form of iron released from corroded iron pipe surfaces (4, 5, 23). Aqueous Fe(II) concentrations up to 0.02 mM were observed in a pipe loop reactor constructed from corroded 70-year-old galvanized iron pipe and operated under anoxic conditions (5). Although the 1 mM Fe(II) used in this research is likely greater than that in drinking water distribution systems, the amount of surface-associated Fe(II) in the reactors is similar to the Fe(II) content in surficial pipe deposits (3-30%; 1, 4). Such Fe(II) concentrations were also necessary to provide sufficient reducing capacity to monitor the DBP reaction kinetics and pathways.

Acknowledgments This work was supported by a grant from the National Science Foundation (BES-0332085). We thank the Minnesota Supercomputing Institute for computational resources, Dr. Christopher J. Cramer for computational chemistry guidance, and Dr. Raul Caretta for performing the BET analyses.

Supporting Information Available Chemical purity and supplier information, iron mineral synthesis, XRD patterns of the synthesized goethite and magnetite, reactor setup and sampling, analytical methods, kinetic modeling, details of the computation of one-electron reduction potentials, TCAN hydrolysis as a function of pH, sorption isotherms of 1,1,1-TCP, and correlation between

E1H and Xg. This material is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) Sarin, P.; Snoeyink, V. L.; Bebee, J.; Kriven, W. M.; Clement, J. A. Physicochemical characteristics of corrosion scales in old iron pipes. Water Res. 2001, 35, 2961-2969. (2) American Water Works Association; Water:/Stats, the Water Utility Database: 1996 Survey (CD); American Water Works Association: Denver, CO, 1998. (3) Tuovinen, O. H.; Button, K. S.; Vuorinen, A.; Carlson, L.; Mair, D. M.; Yut, L. A. Bacterial, chemical, and mineralogical characteristics of tubercles in distribution pipelines. J.-Am. Water Works Assoc. 1980, 72 (11, Pt 1), 626-635. (4) Lin, J.; Ellaway, M.; Adrien, R. Study of corrosion material accumulated on the inner wall of steel water pipe. Corros. Sci. 2001, 43, 2065-2081. (5) Sarin, P.; Snoeyink, V. L.; Bebee, J.; Jim, K. K.; Beckett, M. A.; Kriven, W. M.; Clement, J. A. Iron release from corroded iron pipes in drinking water distribution systems: Effect of dissolved oxygen. Water Res. 2004, 38, 1259-1269. (6) Singer, P. C.; Obolensky, A.; Greiner, A. DBPs in chlorinated North Carolina drinking waters. J.-Am. Water Works Assoc. 1995, 87 (10), 83-92. (7) Arora, H.; LeChevallier, M. W.; Dixon, K. L. DBP occurrence survey. J.-Am. Water Works Assoc. 1997, 89 (6), 60-68. (8) Williams, D. T.; LeBel, G. L.; M.Benoit, F. Disinfection byproducts in Canadian drinking water. Chemosphere 1997, 34, 299-316. (9) Gurol, M. D.; Wowk, A.; Myers, S.; Suffet, I. H. In Kinetics and mechanism of haloform formation: Chloroform formation from trichloroacetone. Water Chlorination: Environmental Impact and Health Effects; Robert, L. J., et al., Eds.; Ann Arbor Science Publishers: Ann Arbor, MI, 1983; Vol. 4. (10) Krasner, S. W.; McGuire, M. J.; Jacangelo, J. G.; Patania, N. L.; Reagan, K. M.; Aieta, E. M. The occurrence of disinfection byproducts in US drinking water. J.-Am. Water Works Assoc. 1989, 81 (8), 41-53. (11) Nikolaou, A. D.; Lekkas, T. D.; Kostopoulou, M. N.; Golfinopoulos, S. K. Investigation of the behaviour of haloketones in water samples. Chemosphere 2001, 44, 907-912. (12) Glezer, V.; Harris, B.; Tal, N.; Iosefzon, B.; Lev, O. Hydrolysis of haloacetonitriles: Linear free energy relationship, kinetics and products. Water Res. 1999, 33, 1938-1948. (13) Reckhow, D. A.; Platt, T. L.; MacNeill, A. L.; McClellan, J. N. Formation and degradation of dichloroacetonitrile in drinking waters. J. Water Supp. Res. Technol. - AQUA 2001, 50, 1-13. (14) Weightman, A. L.; Weightman, A. J.; Slater, J. H. Microbial dehalogenation of trichloroacetic acid. World J. Microbiol. Biotechnol. 1992, 8, 512-518. (15) Yu, P.; Welander, T. Growth of an aerobic bacterium with trichloroacetic acid as the sole source of energy and carbon. Appl. Microbiol. Biotechnol. 1995, 42, 769-774. (16) McRae, B. M.; LaPara, T. M.; Hozalski, R. M. Biodegradation of haloacetic acids by bacterial enrichment cultures. Chemosphere 2004, 55, 915-925. (17) De Wever, H.; Cole, J. R.; Fettig, M. R.; Hogan, D. A.; Tiedje, J. M. Reductive dehalogenation of trichloroacetic acid by Trichlorobacter thiogenes gen. nov., sp. nov. Appl. Microbiol. Biotechnol. 2000, 66, 2297-2301. (18) Rodriguez, M. J.; Serodes, J.-B.; Levallois, P. Behavior of trihalomethanes and haloacetic acids in a drinking water distribution system. Water Res. 2004, 38, 4367-4382. (19) Landmeyer, J. E.; Bradley, P. M.; Thomas, J. M. Biodegradation of disinfection byproducts as a potential removal process during aquifer storage recovery. J. Am. Water Resour. Assoc. 2000, 36, 861-867. (20) Hozalski, R. M.; Zhang, L.; Arnold, W. A. Reduction of haloacetic Acids by Fe0: Implications for treatment and fate. Environ. Sci. Technol. 2001, 35, 2258-2263. (21) Zhang, L.; Arnold, W. A.; Hozalski, R. M. Kinetics of haloacetic acid reactions with Fe(0). Environ. Sci. Technol. 2004, 38, 6881-6889. (22) Vikesland, P. J.; Valentine, R. L. Iron oxide surface-catalyzed oxidation of ferrous iron by monochloramine: Implications of oxide type and carbonate on reactivity. Environ. Sci. Technol. 2002, 36, 512-519. (23) Sarin, P.; Clement, J. A.; Snoeyink, V. L.; Kriven, W. M. Iron release from corroded, unlined cast-iron pipe. J.-Am. Water Works Assoc. 2003, 95 (11), 85-96.

(24) Klausen, J.; Troeber, S. P.; Haderlein, S. B.; Schwarzenbach, R. P. Reduction of substituted nitrobenzenes by Fe(II) in aqueous mineral suspensions. Environ. Sci. Technol. 1995, 29, 23962404. (25) Amonette, J. E.; Workman, D. J.; Kennedy, D. W.; Fruchter, J. S.; Gorby, Y. A. Dechlorination of carbon tetrachloride by Fe(II) associated with goethite. Environ. Sci. Technol. 2000, 34, 46064613. (26) Pecher, K.; Haderlein, S. B.; Schwarzenbach, R. P. Reduction of polyhalogenated methanes by surface-bound Fe(II) in aqueous suspensions of iron oxides. Environ. Sci. Technol. 2002, 36, 17341741. (27) Charlet, L.; Silvester, E.; Liger, E. N-compound reduction and actinide immobilization in surficial fluids by Fe(II): The surface tFeIIIOFeIIOH° species, as major reductant. Chem. Geol. 1998, 151, 85-93. (28) Wang, S.; Arnold, W. A. Abiotic reduction of dinitroaniline herbicides. Water Res. 2003, 37, 4191-4201. (29) Elsner, M.; Schwarzenbach, R. P.; Haderlein, S. B. Reactivity of Fe(II)-bearing minerals toward reductive transformation of organic contaminants. Environ. Sci. Technol. 2004, 38, 799807. (30) Elsner, M.; Haderlein, S. B.; Kellerhals, T.; Luzi, S.; Zwank, L.; Angst, W.; Schwarzenbach, R. P. Mechanisms and products of surface-mediated reductive dehalogenation of carbon tetrachloride by Fe(II) on goethite. Environ. Sci. Technol. 2004, 38, 2058-2066. (31) Pearson, C. R.; Hozalski, R. M.; Arnold, W. A. Degradation of chloropicrin in the presence of Fe0. Environ. Toxicol. Chem., in press. (32) Cervini-Silva, J.; Wu, J.; Larson, R. A.; Stucki, J. W. Transformation of chloropicrin in the presence of iron-bearing clay minerals. Environ. Sci. Technol. 2000, 34, 915-917. (33) Larson, R. A.; Cervini-Silva, J. Dechlorination of substituted trichloromethanes by an iron (II) porphyrin. Environ. Toxicol. Chem. 2000, 19, 543-548. (34) Fennelly, J. P.; Roberts, A. L. Reaction of 1,1,1-trichloroethane with zerovalent metals and bimetallic reductants. Environ. Sci. Technol. 1998, 32, 1980-1988. (35) Choi, W.; Hoffman, M. R. Kinetics and mechanism of CCl4 photoreductive degradation on TiO2: The role of trichloromethyl radical and dichlorocarbene. J. Phys. Chem. 1996, 100, 21612169. (36) McCormick, M. L.; Bouwer, E. J.; Adriaens, P. Carbon tetrachloride transformation in a model iron-reducing culture: Relative kinetics of biotic and abiotic reactions. Environ. Sci. Technol. 2002, 36, 403-410. (37) Tabakoff, B.; Vugrincic, C.; Anderson, R.; Alivisatos, S. G. A. Reduction chloral hydrate to trichloroethanol in brain extracts. Biochem. Pharmacol. 1974, 23, 455-460. (38) Cui, D.; Eriksen, T. E. Reduction of pertechnetate by ferrous iron in solution: Influence of sorbed and precipitated Fe(II). Environ. Sci. Technol. 1996, 30, 2259-2262. (39) Liger, E.; Charlet, L.; Van Cappellen, P. Surface catalysis of uranium (VI) reduction by iron (II). Geochim. Cosmochim. Acta 1999, 63, 2939-2955. (40) Buerge, I. J.; Hug, S. J. Influence of mineral surfaces on chromium (VI) reduction by iron (II). Environ. Sci. Technol. 1999, 33, 42854291. (41) Klupinski, T. P.; Chin, Y.-P.; Traina, S. J. Abiotic degradation of pentachloronitrobenzene by Fe(II): Reactions on goethite and iron oxide nanoparticles. Environ. Sci. Technol. 2004, 38, 43534360. (42) Williams, A. G. B.; Scherer, M. M. Spectroscopic evidence for Fe(II)-Fe(III) electron transfer at the iron oxide-water interface. Environ. Sci. Technol. 2004, 38, 4782-4790. (43) Wu, H. Chemical property calculation through JavaScript and applications in QSAR. Molecules (Electron. Pub.) 1999, 4, 1627. (44) Xie, Q.; Sun, H.; Xie, G.; Zhou, J. An iterative method for calculation of group electronegativities. J. Chem. Inf. Comput. Sci. 1995, 35, 106-109. (45) Boyd, R. J.; Edgecombe, K. E. Atomic and group electronegativities from the electron-density distributions of molecules. J. Am. Chem. Soc. 1988, 110, 4182-4186. (46) Boyd, R. J.; Boyd, S. L. Group electronegativities from the bond critical point model. J. Am. Chem. Soc. 1992, 114, 1652-1655. (47) Plewa, M. J.; Wagner, E. D.; Jazwierska, P.; Richardson, S. D.; Chen, P. H.; McKague, A. B. Halonitromethane drinking water VOL. 39, NO. 21, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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disinfection byproducts: Chemical characterization and mammalian cell cytotoxicity and genotoxicity. Environ. Sci. Technol. 2004, 38, 62-68. (48) Moudgal, C. J.; Lipscomb, J. C.; Bruce, R. M. Potential health effects of drinking water disinfection byproducts using quantitative structure toxicity relationship. Toxicology 2000, 147, 109131. (49) Valentine, R. L.; Angerman, B.; Hackett, S.; Vikesland, P.; Slattenow, S. Characterization of disinfectant decay and DBP formation in the presence of water distribution system deposits; AWWA Water Quality Technology Conference: Tampa, FL, 1999.

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(50) Sander, A.; Berghult, B.; Broo, A. E.; Johansson, E. L.; Hedberg, T. Iron corrosion in drinking water distribution systems-the effect of pH, calcium and hydrogen carbonate. Corros. Sci. 1996, 38, 443-455.

Received for review June 3, 2005. Revised manuscript received August 10, 2005. Accepted August 15, 2005. ES051044G