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H2O2 advanced oxidation process
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Simultaneous degradation of disinfection byproducts and earthy-musty odorants by the UV/H2O2 advanced oxidation process Chang Hyun Jo a,1, Andrea M. Dietrich a,*, James M. Tanko b a b
Department of Civil and Environmental Engineering, Virginia Polytechnic Institute and State University, VA 24061, USA Department of Chemistry, Virginia Polytechnic Institute and State University, VA 24061, USA
article info
abstract
Article history:
Advanced treatment technologies that control multiple contaminants are beneficial to
Received 27 September 2010
drinking water treatment. This research applied UV/H2O2 for the simultaneous degradation
Received in revised form
of geosmin, 2-methylisoborneol, four trihalomethanes and six haloacetic acids. Experi-
3 February 2011
ments were conducted in de-ionized water at 24 1.0 C with ng/L amounts of odorants
Accepted 4 February 2011
and mg/L amounts of disinfection byproducts. UV was applied with and without 6 mg/L
Available online 12 February 2011
H2O2. The results demonstrated that brominated trihalomethanes and brominated haloacetic acids were degraded to a greater extent than geosmin and 2-methylisoborneol.
Keywords:
Tribromomethane and dibromochloromethane were degraded by 99% and 80% respec-
Hydroxyl radical
tively at the UV dose of 1200 mJ/cm2 with 6 mg/L H2O2, whereas 90% of the geosmin and
Ultraviolet
60% of the 2-methylisoborneol were removed. Tribromoacetic acid and dibromoacetic acid
Geosmin
were degraded by 99% and 80% respectively under the same conditions. Concentrations of
2-Methylsioborneol
trichloromethane and chlorinated haloacetic acids were not substantially reduced under
Trihalomethanes
these conditions and were not effectively removed at doses designed to remove geosmin
Haloacetic acids
and 2-methylisoborneol. Brominated compounds were degraded primarily by direct photolysis and cleavage of the CeBr bond with pseudo first order rate constants ranging from 103 to 102 s1. Geosmin and 2-methylisoborneol were primarily degraded by reaction with hydroxyl radical with direct photolysis as a minor factor. Perchlorinated disinfection byproducts were degraded by reaction with hydroxyl radicals. These results indicate that the UV/H2O2 can be applied to effectively control both odorants and brominated disinfection byproducts. ª 2011 Elsevier Ltd. All rights reserved.
1.
Introduction
Advanced oxidation processes (AOP) in water treatment involve hydroxyl radicals ($OH) as intermediates. UV-based AOPs essentially mimic the photo-initiated oxidation processes in natural systems, such as sun light on surface
water or in the atmosphere (Oppenlander, 2003). Ultraviolet (UV) irradiation is well established for disinfection of water. The UV/H2O2 provides oxidation through generation of hydroxyl radicals by photolysis of H2O2 (Liao and Gurol, 1995; Rudra et al., 2005). This process degrades organic contaminants, including recalcitrant odorants such as geosmin and
* Corresponding author. E-mail address: [email protected] (A.M. Dietrich). 1 Present address: K-water, Daejeon, Korea. 0043-1354/$ e see front matter ª 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2011.02.006
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2-methylisoborneol (2-MIB), mainly by the reaction with hydroxyl radicals and partially by direct UV photolysis of the contaminant (Beltran et al., 1993; Stefan and Bolton, 1998; Rosenfeldt et al., 2005; Rudra et al., 2005; Paradis and Hoffman, 2006; Rosenfeldt and Linden, 2007). Solution-phase AOPs are advantageous for contaminant removal because residual solids are not produced such as would occur with spent activated carbon from an adsorption process. Recently, UV systems labeled as “dual purpose” were applied to fullscale water treatment plants. These dual systems combine low intensity UV for disinfection and high intensity UV for both disinfection and advanced oxidation of odorants. Their operation is limited by both UV transmittance and water quality constraints because low UV transmittance attenuates UV light available to produce hydroxyl radicals from H2O2, and carbonates scavenge hydroxyl radicals (Ho et al., 2004; Cotton and Collins, 2006). The optimum H2O2 dose in the UV/H2O2 process should be empirically determined because excess H2O2 can be a hydroxyl radical scavenger (Wang et al., 2000). Typical doses applied for taste-and-odor control in full-scale treatment plants vary from >0 to 11 mg/L H2O2 (Royce and Stefan, 2005). Geosmin and 2-MIB are earthy-musty odorants found in surface water and drinking water. They are produced by cyanobacteria or actinomycetes, commonly in the warm summer and early fall seasons (Ju¨ttner and Watson, 2007). Concentrations in drinking water can range from a few to >100 ng/L; to minimize consumer complaints and produce acceptable quality water, concentrations of these odorants should be brought to below their odor threshold levels of 1e10 ng/L (Rashash et al., 1997; Dietrich, 2006; Piriou et al., 2009). Although not regulated for health effects, esthetic guidelines of 10 ng/L geosmin and 10 ng/L 2-MIB have been established in Korea and as secondary standards in Japan. These compounds are difficult to remove by conventional water treatment processes; thus activated carbon or AOPs are frequently used for their control during full-scale treatment. 2-MIB has repeatedly proven the more difficult to remove and reacts slower with hydroxyl radicals compared to geosmin (Glaze et al., 1990; Westerhoff et al., 2006; Peter and Von Gunten, 2007). In addition to problematic taste and odor compounds, disinfection byproducts (DBPs) are another major issue in drinking water quality. Most organic DBPs form from the reaction with humic or fulvic acid and disinfectants. The major classes of regulated DBPs in drinking water are trihalomethanes (THMs), which are regulated at a total concentration of 80 mg/L, and haloacetic acids (HAAs) which are regulated at a total concentration of 60 mg/L (USEPA, 2011). Chlorinated DBPs including trichloromethane and mono, di-, and trichloroacetic acid are the most prevalent (Krasner et al., 2006), but the occurrence of brominated and iodinated DBPs is both pervasive and problematic because bromo-and iodocontaining DBPs are more toxic than their chlorinated analogs (Krasner et al., 2006; Richardson et al., 2007). Brominated DBPs can regionally constitute > 40% THMs and 10e25% HAAs (Krasner et al., 1989; Hyun et al., 2005). In addition to regulated DBPs, over 90 other brominated DBPs were identified by GCeMS (Richardson et al., 2007) and additional polar brominated DBPs were detected by LCeMS (Zhang et al., 2008).
Advanced treatment technologies that control multiple contaminants and provide disinfection in full-scale water treatment are immensely beneficial to the drinking water industry. Studies report that UV-based AOPs control DBP precursors and consequently reduce DBP level in finished water (Wang et al., 2000; Chin and Berube, 2005; Sarathy and Mohseni, 2010). However, many water utilities apply prechlorination to control taste/odor or iron/manganese or ammonia nitrogen, and to obtain required CT values. Prechlorination occurs prior to coagulation and produces a variety of DBPs while AOPs typically occur after filtration to increase UV transmission, which is also after formation of DBPs. There have been only a few studies on the removal of DBPs by UV/H2O2 and these reports are contradictory. Rudra et al. (2005) reported over 90% degradation of THMs at high UV (17 000 mJ/cm2) and 0.1% H2O2 dose. In pilot-scale and bench-scale studies with natural surface waters, THMs increased at lower UV/H2O2 doses and decreased with higher UV/H2O2 doses, and HAAs decreased for two samples and increased for one sample (Paradis and Hoffman, 2006). Second order rate constants for the reaction of DBPs and odorous contaminants with hydroxyl radical have been measured (Mezyk et al., 2006; Westerhoff et al., 2006; Peter and Von Gunten, 2007). Few studies have addressed the possibility and mechanisms of DBPs degradation by UV/H2O2 process in aqueous phase. Two mechanisms are associated with UV/H2O2 oxidation of contaminants (Scheme 1). Photolysis of H2O2 produces hydroxyl radicals that subsequently react with contaminants, generally by abstracting a hydrogen or by adding to an unsaturated site. Direct UV photolysis of the contaminant can occur and result in bond homolysis and radical generation; the resulting carbon-centered radicals subsequently are oxidized by reaction with H2O2, O2, and contaminants. Direct UV photolysis can contribute substantially to contaminant degradation. Geosmin and 2-MIB were decreased by 40% and 20%, respectively, with a UV dose of 1700 mJ/cm2 in raw water blends with 1.7e2.3 mg/L TOC, 105 mg/L alkalinity as CaCO3, and pH 8.0 to 8.3 (Rosenfeldt et al., 2005). Other source water algal-contaminants degraded by UV include microcystin (Qiao et al., 2005) and conjugated odorous aldehydes (Jo and Dietrich, 2009). In regard to UV photolysis of THMs, only the brominated THMs were photolyzed and quantum yield of the photolysis for all brominated THMs was 0.43 with a 253.7 nm lamp (Nicole et al.,
Direct Photolysis hv
R3C-X
R 3C
+
X
Hydroxyl radical generation (H2O2 photolysis; hydrogen abstraction) hv
H 2O 2
HO
+
R-H
2 HO
H2O
+
R
Scheme 1 e Mechanisms for oxidation by UV alone and with H2O2. (“X” refers to a halogen).
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1991). In the same research, polybrominated THMs were shown to be photolyzed faster than monobrominated compounds. Concentrations of tribromomethane and chlorodibromomethane in chlorinated swimming pool water decreased significantly with medium pressure UV irradiation of 145 mJ/cm2 (Cassan et al., 2006). Polyhalomethanes such as tribromomethane (CHBr3) and carbon tetrabromide (CBr4) were photolyzed by a proposed water-catalyzed OeH insertion/HBr elimination (Kwok et al., 2004; Lin et al., 2005; Zhao et al., 2005). For the direct UV photolysis and hydroxyl radical reactions of organic compounds such as geosmin/2-MIB and pharmaceutical compounds by UV/H2O2, pseudo-first order reaction models have been proposed because under constant irradiation, the steady-state concentration of the hydroxyl radical is approximately constant with respect to pollutant concentration because of the relatively high concentration of H2O2 (Rosenfeldt et al., 2005; Pereira et al., 2007). d½C ¼ k0 ½C dt Where, k0 ¼ k0d þ k0i k0 ¼ the observed pseudo-first order rate constant (s1) k0d ¼ the measured pseudo-first order rate constant of direct photolysis (s1) k0i ¼ the measured pseudo-first order rate constant of the reaction with $OH (s1) Second order reaction rate constants for reaction of hydroxyl radical with geosmin have been reported to range from 8.2 109 to 1.4 1010 M1s1, and for 2-MIB from 5.1 109 to 8.1 109 M1s1 (Glaze et al., 1990; Westerhoff et al., 2006; Peter and Von Gunten, 2007), which are nearly an order of magnitude more than for the reaction of hydroxyl radical with THMs (0.7 107e1.5 108 M1s1) or chlorinated HAAs (6 107e1.0 108 M1s1). The rate constant for reaction of tribromomethane with HO$ is greater than trichloromethane by a factor of 10 (Maruthamuthu et al., 1995; Mezyk et al., 2006). This paper summarizes our results from the investigation of the simultaneous degradation of odorants and DBPs at concentrations and conditions representative of those used for the removal of recalcitrant odorants in drinking water using UV/H2O2 based on typical H2O2 dose used in full-scale water treatment plant. The objectives of this research were to: 1) determine the degree of degradation of THMs and HAAs with respect to degradation of geosmin and 2-MIB, and 2) evaluate the role of UV photolysis and hydroxyl radical reaction in this degradation.
2.
Methods and materials
2.1.
Apparatus
Experiments were performed with a Rayonet RPR-100 photochemical reactor equipped with 253.7 nm wavelength UV lamps of 7.2 mW/cm2 total intensity, and quartz reaction vessels. UV dose was confirmed with the iodide/iodate
actinometer (Rahn, 2004; Rahn et al., 2006). Samples were completely mixed and headspace free while being irradiated with UV. A reaction temperature of 24 1.0 C was maintained by an electric fan set under the sample, and ice.
2.2.
Reagents and sample preparation
Samples for UV/H2O2 irradiation were prepared directly in deionized water (Nanopure) using individual neat compounds for DBPs; trichloromethane (99%, Fisher Scientific), tribromomethane (99%, Acros Organics), chloroacetic acid (MCAA) (99%, Aldrich), dichloroacetic acid (DCAA) (99%, SigmaeAldrich), trichloroacetic acid (TCAA) (99%, Alfa Aesar), bromoacetic acid (MBAA) (99%, SigmaeAldrich), dibromoacetic acid (DBAA) (99%, Fluka), tribromoacetic acid (TBAA) (99%, Acros Organics), tetrachloromethane (>99%, Acros Organics) and tetrabromomethane (>99%, Acros Organics). To prepare individual and mixed samples of HAAs to react with UV/H2O2, neat HAA compounds were dissolved in water because the commercial HAA standard mixture was dissolved in methyl tert-butyl ether (MTBE) and this solvent scavenges hydroxyl radicals (second order rate constant, k ¼ 3.9 109 M1s1) (Chang and Young, 2000). For quantitative chromatographic analyses, commercial standards solutions of THMs in methanol and HAAs in methyl-tert-butyl ether were purchased from Ultra Scientific. Geosmin (200 mg/L) and 2-MIB (100 mg/L) were purchased from Supelco as solutions in methanol and diluted in water to the appropriate concentration. Typical odorant solutions used in this research (Table 1) consistently contained about 2 mL or 1.5 mg/L (50 mM) methanol from dilution of the concentrated odorant solutions. Hydrogen peroxide (30%, Fisher) was
Table 1 e Typical initial concentrations of compounds in the research. Compounds
Typical concentrations mg/L
mM
a
Odorants Geosmin 2-MIB THMsb Trichloromethane Bromodichloromethane Dibromochloromethane Tribromomethane Tetrahalomethanes Carbon tetrachloride Carbon tetrabromide HAAsb Chloroacetic acid (MCAA) Dichloroacetic acid (DCAA) Trichloroacetic acid (TCAA) Bromoacetic acid (MBAA) Dibromoacetic acid (DBAA) Tribromoacetic acid (TBAA) Hydrogen Peroxide
0.04e0.2 0.1e0.3
0.0002e0.001 0.0006e0.002
60e500 90 80 80e550
0.5e4.2 0.5 0.4 0.3e2.2
350 1000
2.3 3.0
270 190 180 200 190 160 6000
2.9 1.5 1.1 1.4 0.9 0.5 176.5
a Added with z1.5 mg/L methanol. b It should be noted that these concentrations are slightly higher than those usually found in actual drinking waters.
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2.4.
Statistical analyses
Microsoft Excel (2007) was applied for data manipulation and calculation of apparent first order rate constant and standard errors.
3.
Results
3.1.
UV absorbance
To investigate the relative role of hydroxyl radical production versus direct UV photolysis of the organic contaminants in
3.2. Degradation of odorants and DBPs at typical concentrations found in drinking water Concentrations of ng/L geosmin/2-MIB and mg/L DBPs were prepared in mixed aqueous samples and contaminant removals were simultaneously followed under the same UV/ H2O2 and solution conditions. The solutions contained about 1.5 mg/L methanol, because it was the solvent used to dissolve geosmin and 2-MIB; methanol is known to act as a hydroxyl radical scavenger (Nickelsen et al., 1994). In practice, a utility would apply a UV/H2O2 dose for geosmin/2-MIB degradation that met their treatment need to control the odor problem. Thus, although the water being treated may contain other contaminants and radical scavengers that could absorb at UV 253.7 nm or react with hydroxyl radicals, the applied UV/H2O2 dose would be determined by treatment goals for removing geosmin and 2-MIB.
1.20 1.00 0.80 0.60 0.40 0.20 0.00
A
H 2O ge 2 os m in 2M IB C H Br 3 C H C l3 C Br 4 C C l4 M BA A D BA A TB AA M C AA D C AA TC AA
1200 1000 800 600 400 200 0
B
H 2O ge 2 os m in 2M IB C H Br 3 C H C l3 C Br 4 C Cl 4 M BA A D BA A TB AA M C AA D C A A TC AA
Geosmin and 2-MIB were measured by solid-phase microextraction (SPME, Supelco) as reported by other researchers (Watson et al., 1999, 2000). Compounds partitioned from water were sorbed onto an SPME fiber (65 mm, PDMS/DVB) for 10 min at 60 C. The SPME fiber was injected in the splitless mode into the GC/MS (Agilent 5973) at 220 C and desorbed for 2 min. An Rtx-5Sil column (30 m, 0.25 mm ID) with He carrier gas and a temperature program of 60 Ce180 C by 15 C/min was used. Approximate retention times were 5.4 min for 2-MIB and 7.9 min for geosmin; m/z values of 112, 125, 182 for geosmin and m/z values of 95, 108, 168 for 2-MIB were monitored for qualitative analysis and m/z values of 112 or 95 were used for quantitative analysis for geosmin and 2-MIB respectively in selective ion mode. Concentrations of disinfection byproducts were determined using standardized methods. THMs were measured based on Standard Method 6232.D by purge/trap (Tekmar 3000) and GC (Tremetrics 9001) with DB-624 column (J & W) and ECD detector. GC temperature was initially maintained at 45 C for 3 min, and then increased by 11 C/min up to 200 C. HAAs were determined by liquideliquid extraction method (EPA Method 552.2) and GC (HP 5890) with an ECD detector. The injector temperature was 210 C and the initial oven temperature was set to 35 C and increased up to 140 C. UV adsorption was measured on a UV/vis spectrophotometer (Beckman, DU640). H2O2 concentration was determined by triiodide (I 3 ) titration method (Klassen et al., 1994).
Molar extinction
Chemical analyses
coefficient (M -1 cm -1 )
2.3.
this UV/H2O2 process, molar extinction coefficients were measured (Fig. 1A). The brominated compounds, geosmin, and 2-MIB had at least two order of magnitude higher molar extinction coefficients than chlorinated compounds, and one order of magnitude lower than H2O2. Absolute molar extinction coefficients, however, can be misleading. At typical concentrations used in this research (Table 1), only H2O2 and the brominated DBPs would absorb an appreciable amount of UV as shown by their absorptions relative to H2O2 (Fig. 1B). The relative absorbance of geosmin, 2-MIB, and any of the chlorinated compounds are nil. The UV absorbances of the brominated DBPs are sufficiently high that a plausible mechanism for brominated DBPs elimination is by direct UV photolysis (CeBr bond cleavage). At 6 mg/L H2O2, hydrogen peroxide absorbs most of the UV at 253.7 nm, which indicates that hydroxyl radicals can be produced from UV photolysis of H2O2 under these conditions. In this research, the absorbance of 6 mg/L H2O2 at 253.7 nm was ca. 0.035 in the experimental apparatus.
Relative absorbance
diluted to desired concentrations of 6 mg/L which was selected based on typical concentration range in pilot-scale study (Paradis and Hoffman, 2006), and added into the samples immediately before UV irradiation. Typical initial concentrations of compounds used in the research are shown in Table 1; it should be noted that these concentrations are slightly higher than those usually found in actual drinking waters. Molar extinction coefficients at 253.7 nm were determined for each compound dissolved in distilled water; these concentrations were measured: 34 mg/L H2O2, 0.4 mg/L geosmin, 0.1 mg/L 2-MIB, 340 mg/L for CHBr3, CHCl3, or CBr4, 150 mg/L CCl4, and 1000 mg/L for MBAA, DBAA, TBAA, MCAA, DCAA, and TCAA. Absorptions relative to H2O2 were calculated by applying the concentrations in Table 1, and using 100 ng/L for geosmin and 2-MIB, and 300 mg/L for CHCl3 and CHBr3.
Fig. 1 e Molar extinction coefficients (A) and relative absorption (relative to H2O2 [ 1) (B) of odorant and DBP compounds and hydrogen peroxide measured at 253.7 nm and at the concentrations used in this research. Concentrations used for these measurements are provided in the methods and materials.
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3.2.1.
Geosmin and 2-MIB
The geosmin and 2-MIB results show 90 and 65% degradation, respectively, with a UV dose of 1200 mJ/cm2 and 6 mg/L H2O2. This dose was selected as it was sufficient to reduce the initial geosmin concentration of 40e43 ng/L to 4 ng/L (Figs. 2 and 3), which is below its guideline and upper odor threshold value of 10 ng/L. Under identical conditions, but in the absence of H2O2, only about 20% was degraded with UV photolysis (Fig. 2).Using the results in Figs. 2 and 3, apparent pseudo first order rate constants and half-lives for degradation of these odorants can be calculated (Table 2). The apparent rate constants for degradation of both geosmin and 2-MIB are approximately 3e7 greater when H2O2 is present. These results suggest that geosmin and 2-MIB concentrations are mainly reduced by reaction with hydroxyl radical (formed by photolysis of H2O2) though a small amount may also degraded by direct photolysis.
3.2.2.
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Fig. 3 e Photolysis of geosmin, 2-MIB and THMs in the presence of 6 mg/L H2O2. Initial concentration (C0): geosmin [ 43 ng/L, 2-MIB [ 100 ng/L, trichloromethane [ 63 mg/L, bromodichloromethane [ 94 mg/L, dibromochloromethane [ 76 mg/L, tribromomethane [ 82 mg/L, methanol z1.5 mg/L.
Geosmin/MIB and THMs
3.2.2.1. THMs degradation compared to geosmin/2-MIB by UV/H2O2. Brominated THMs were degraded better than trichloromethane at the UV/H2O2 dose effective for removing geosmin/2-MIB (Fig. 3). Tribromomethane and dibromochloromethane were degraded by 99% and 80%, respectively, at the UV dose of 1200 mJ/cm2 and 6 mg/L H2O2, while 90% of the geosmin and 65% of the 2-MIB were degraded at this dose. The brominated THMs with higher numbers of bromine atoms were degraded faster than ones with lower numbers of bromine or trichloromethane, which for all practical purposes, was not degraded by UV/H2O2 (Fig. 3). Tribromomethane was degraded faster than either geosmin/2-MIB. Another approach to determining an effective UVeH2O2 dose from the data in Fig. 3 is to use the guideline of 10 ng/L 2-MIB that has been established in Japan and Korea. Accordingly, a UV dose of 2700 mJ/cm2 and 6 mg/L H2O2 was required to remove 90% of the more recalcitrant odorant, 2-MIB, to a final concentration of 10 ng/L 2-MIB. At this UV/H2O2 dose for 90% removal of 2-MIB and the initial concentrations used for this experiment, the final concentrations of the other contaminants were: 0.8 ng/L geosmin (98% degradation);
Fig. 2 e Photolysis of geosmin and 2-MIB in the presence and absence of 6 mg/L H2O2. Initial concentration (C0): geosmin (no H2O2) [ 40 ng/L, geosmin (H2O2 6 mg/ L) [ 183 ng/L, 2-MIB (no H2O2) [ 108 ng/L, 2-MIB (6 mg/L H2O2) [ 306 ng/L, methanol z1.5 mg/L.
47 mg/L trichloromethane (26% degradation); 44 mg/L monobromodichloromethane (53% degradation); 0.6 mg/L dibromochloromethane (99% degradation); and <0.1 mg/L tribromomethane (>99.9% degradation). Although the THM concentrations used in this experiment were higher than concentrations typical for drinking water, the UVeH2O2 conditions applied in this research to degrade 2-MIB below its esthetic guideline would also be effective to eliminate dibromochloromethane and tribromomethane to below their regulatory limits.
3.2.2.2. Direct UV photolysis of brominated THMs. To investigate the contribution of direct UV photolysis, an aqueous mixture of THMs was reacted with UV in the absence of H2O2. Using the data in Figs. 3 and 4, apparent pseudo first order rate constants and half-lives were calculated (Table 3). The data clearly show that: a) the apparent rate constants for the brominated THMs were substantially faster than for trichloromethane, b) the rate constant for the brominated THMs increased with the number of bromine atoms in the molecule, c) the rate constants for the brominated THMs were identical in the absence and presence of H2O2. These observations make sense because the CeBr bond serves as the active chromophore which is effectively cleaved by UV (i.e., CeBr þ hn / C$ þ Br$) and chlorine-only containing compounds do not absorb appreciably at 253.7 nm, and thus, cannot be eliminated by direct photolysis.
3.2.2.3. Degradation mechanism for THMs. To further investigate the role of hydrogen abstraction and its effects on THM degradation, the reaction of carbon tetrachloride (CCl4) and carbon tetrabromide (CBr4) were compared to each other and to their corresponding trihalomethane analogs. All halomethane solutions were prepared by dissolving neat compounds in de-ionized water then individually treating with UV/H2O2. Because they do not possess abstractable hydrogens, carbon tetrachloride and carbon tetrabromide are effectively non-reactive toward hydroxyl radical and can only be degraded by direct photolysis. Both the tri- and tetrabrominated methanes were degraded faster than their
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Table 2 e Apparent first order rate constants for photo-induced removal of geosmin and 2-MIB in the presence and absence of 6 mg/L H2O2. hn/H2O2, UV 253.7 nm
Contaminant
kapp, s1 Geosmin
2-MIB
a b c d
1.6 1.9 1.6 7.2 6.8 6.0
(0.3) (0.3) (0.3) (0.2) (0.3) (0.5)
kapp, s1
t½, s 2 b
10 102 102 103 103 103
c d b c d
krela
hn only, UV 253.7 nm
1.2 (0.6) 10
43.3 36.5 43.3 96.3 101.9 115.5
t½, s 3 b
1.1 (0.1) 103
b
577.6
14.1
630.1
6.1
Relative rate constant: mean kapp(hn/H2O2)/kapp(hn only). Based upon data from Fig. 2; standard error of slopes provided. Based upon data from Fig. 3; standard error of slopes provided. Based upon data from Fig. 6; standard error of slopes provided.
chlorinated analogs as shown in Fig. 5. The pseudo first order rate constant for degradation of CCl4 was 5.6 (0.4) 103 s1 and for CBr4 it was 8.4 (0.8) 102 s1 (calculated from the data in Fig. 5). Rate constants for degradation of CX4 (X ¼ Cl or Br) were greater than CHX3, even though the latter possesses a hydrogen atom that can be abstracted by hydroxyl radical. These results further confirm that the different degradation rates in UV/H2O2 between chlorinated and brominated THMs results from the different UV photolysis rates and not from the hydrogen abstraction by hydroxyl radicals.
3.2.3.
Geosmin/2-MIB and HAAs
3.2.3.1. HAAs degradation compared to geosmin/2-MIB by UV/ H2O2. Treatment with UV/H2O2 degraded brominated HAAs faster than chlorinated HAAs (Fig. 6) in distilled water that also contained odorants. Applying a UV dose of 1200 mJ/cm2 and 6 mg/L H2O2 as was done for the geosmin/2-MIB/THM data in Fig. 3, the geosmin concentration was reduced to 7 ng/L (96% removal), while TBAA and DBAA were degraded by 99% and 80% respectively. Chlorinated HAAs with no bromine atoms were barely degraded by a UV doses of 0e4300 mJ/cm2 and 6 mg/L of H2O2. Brominated HAAs degradation rates increased in proportion to the number of bromine atoms in
Fig. 4 e Photolysis of brominated THMs in the presence and absence of H2O2. Initial concentration (C0): trichloromethane [ 63 mg/L, bromodichloromethane [ 94 mg/ L, dibromochloromethane [ 76 mg/L, tribromomethane [ 82 mg/L.
the molecule. Consequently, tribromoacetic acid had the highest degradation rate among all HAAs and geosmin/2-MIB.
3.2.3.2. UV photolysis of brominated HAAs. TBAA, DBAA, and MBAA were treated with UV in the absence of H2O2 to investigate the contribution of direct UV photolysis as the primary mechanism for degradation of brominated HAAs. As shown in Fig. 7, there is not a huge difference in the degradation rates for MBAA, DBAA and TBAA between UV photolysis and UV/ H2O2 process e in fact, the rates in the presence of H2O2 appear to be slightly lower. This becomes clearer when the apparent rate constants and half-lives are compared (Table 4). The fact that the rate constants did not increase in the presence of H2O2 suggests that brominated HAAs are degraded mainly by UV photolysis in UV/H2O2 process, rather than via reaction with $OH.
4.
Discussion
For ng/L odorant concentrations and mg/L THM and HAA concentrations typical of drinking water, UV/H2O2 provided substantial simultaneous degradation of geosmin, 2-MIB, brominated THMs and brominated HAAs; however, chlorinated DBPs experienced only minor degradation under the same conditions. This research confirmed the results of Rosenfeldt et al. (2005) that geosmin and 2-MIB were mainly eliminated by the reaction with hydroxyl radicals in the UV/ H2O2 process and that direct UV photolysis is minor. This research concurs with that of Rudra et al. (2005) who reported removal of THMs by UVeH2O2 and also Nicole et al. (1991) who reported direct UV as the removal mechanism for brominated THMs. Brominated HAAs were demonstrated to be degraded by direct UV photolysis and chlorinated HAAs by UV/H2O2 which is consistent with the results of Paradis and Hoffman (2006), who determined that UVeH2O2 decreased HAAs in two out of three samples. For halogenated THMs, a possible first reaction step in UV/ H2O2 advanced oxidation is either hydrogen abstraction by hydroxyl radical or carbonehalogen bond cleavage by direct UV photolysis. Bond dissociation energies of the CeH bond in trichloromethane (CHCl3) and tribromomethane (CHBr3) are very close, 100.0 and 99.9 kcal/mol, respectively (McGivern
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Table 3 e Apparent first order rate constants for photo-induced removal of brominated and chlorinated THMs in the presence and absence of 6 mg/L H2O2. Contaminant
hn/H2O2, UV 253.7 nm kapp, s1
CHCl3 CHBrCl2 CHBr2Cl CHBr3
1.5 (0.1) 2.2 (0.1) 1.8 (0.1) 1.0 (0.1) 1.3 (0.2) 2.9 (0.3) 2.7 (0.4)
3 b
10 103 b 103 c 102 b 102 c 102 b 102 c
krela
hn only, UV 253.7 nm t½, s
kapp, s1
t½, s
462.1 315.0 385.1 69.3 53.3 23.9 25.7
e 2.1 (0.2) 103 c
330.0
e 1.0
1.0 (0.1) 102 c
69.3
1.2
2.8 (0.3) 102 c
24.8
1.0
a Relative rate constant: mean kapp(hn/H2O2)/kapp(hn only). b Based upon data from Fig. 3; standard error of slopes provided. c Based upon data from Fig. 4; standard error of slopes provided.
et al., 2000) and thus, do not explain the faster degradation of brominated DBPs compared to perchlorinated DBPs. Carbonebromine cleavage due to UV photolysis is the likely mechanism of faster degradation of brominated DBPs. This is
Fig. 5 e Photolysis of halogenated methanes in the absence of H2O2. Initial concentration (C0): trichloromethane [ 514 mg/L, tribromomethane [ 523 mg/L, carbon tetrachloride [ 343 mg/L, carbon tetrabromide [ 926 mg/L.
Fig. 6 e Photolysis of geosmin, 2-MIB and HAAs in the presence of 6 mg/L H2O2. Initial concentration (C0): geosmin [ 183 ng/L, 2-MIB [ 306 ng/L, bromoacetic acid [ 202 mg/L, dibromoacetic acid [ 190 mg/L, tribromoacetic acid [ 161 mg/L, chloroacetic acid [ 271 mg/L, dichloroacetic acid [ 191 mg/L, trichloroacetic acid [ 176 mg/L, methanol z1.5 mg/L.
supported by the higher strengths of CeCl bonds than CeBr bonds in trichloromethane and tribromomethane, 80.1 and 70.4 kcal/mol, respectively (McGivern et al., 2000). The energy of a UV photon at 253.7 nm is about 113 kcal/mol, which is sufficient to break both the CeBr bond (Kwok et al., 2004) and CeCl (Kurac and Hlatka, 1992). The higher molar absorption coefficients of brominated DBPs (Fig. 1) and the equal rates of degradation of brominated THMs by UV photolysis with and without H2O2 (Fig. 4 and Table 3) indicate that brominated compounds absorb more UV and thus are more susceptible to photolysis than chlorinated compounds. Although the CeCl is weak, the chlorinated compounds are not degraded by direct UV photolysis because the CeCl chromophore has a low extinction coefficient, i.e., the molecule does not absorb sufficient light to be degraded by this mechanism. The degradation rates of the brominated DBPs increased with the number of bromine atoms present in the molecule. Tribromomethane and tribromoacetic acid were eliminated the fastest among the THMs and HAAs (Figs. 4 and 7, and Table 4). In this research, brominated THMs and HAAs were shown to be eliminated mainly by UV photolysis in the UV/ H2O2 process. Previously, Nicole et al. (1991) demonstrated that UV photolysis of CHBr3, CHBr2Cl, and CHCl2Br at 253.7 nm in aqueous solution led to complete conversion to the corresponding halide. Regarding UV (253.7) photolysis of tribromomethane, a water-catalyzed mechanism was proposed in which isotribromomethane isomerized from
Fig. 7 e Photolysis of brominated HAAs in the presence and absence of 6 mg/L H2O2. Initial concentration (C0): bromoacetic acid [ 202 mg/L, dibromoacetic acid [ 190 mg/L, tribromoacetic acid [ 161 mg/L.
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w a t e r r e s e a r c h 4 5 ( 2 0 1 1 ) 2 5 0 7 e2 5 1 6
Table 4 e Apparent first order rate constants for photo-induced removal of chlorinated and brominated acetic acids in the presence and absence of H2O2. (explanation of acronyms: AA [ acetic acid; B [ bromo; C [ chloro; M [ mono; D [ di; T [ tri). hn/H2O2, UV 253.7 nm
Contaminant
kapp, s1 MBAA DBAA TBAA MCAA DCAA TCAA
1.1 1.3 6.4 5.8 5.0 2.9
(0.1) (0.1) (0.4) (2.6) (1.2) (2.3)
103 102 102 104 104 104
kapp, s1
t½, s b b b b b b
krela
hn only, UV 253.7 nm
2.1 (0.1) 103 1.5 (0.1) 102 7.4 (1.1) 102 e e e
630.9 53.3 10.8 1195 1390 2390
t½, s c c c
330.0 46.2 9.4
0.52 0.87 0.86
a Relative rate constant: mean kapp(hn/H2O2)/kapp(hn only). b Based upon data from Fig. 6; standard error of slopes provided. c Based upon data from Fig. 7, standard error of slopes provided.
tribromomethane reacts with water molecule resulting in OH insertion and HBr elimination. The major products from tribromomethane were carbon monoxide and three equivalents of HBr for the major route and formic acid plus three equivalents of HBr for the minor route (Kwok et al., 2004). Brominated compounds are increasingly identified as toxic contaminants in drinking water. In addition to the regulated THMs and HAAs, other bromine-containing disinfection byproducts in drinking water include halonitromethanes, 3e4 carbon haloacids, haloaldehydes, haloketones, haloamides, and halogenated furanones (Krasner et al., 2006). Non-DBP brominated compounds of health concern also occur in source and treated water including the regulated volatile organic compound methyl bromide, brominated flame retardants, and odorous bromophenols (Piriou et al., 2007). In this research the UV/H2O2 process reduced the concentrations of geosmin and 2-MIB in distilled water. This is consistent with the results of Rosenfeldt et al. (2005), who reported degradation of geosmin and 2-MIB in distilled water, and also in source waters and treated drinking waters which possessed other reactants for UV and hydroxyl radicals including natural organic matter and bicarbonate/carbonate ions. The degradation of geosmin, 2-MIB, and perchlorinated DBPs depend on reaction with hydroxyl radicals. The degradation of brominated compounds is primarily through direct UV photolysis that would consume few hydroxyl radicals and would not be affected by the presence of hydroxyl radical scavengers. This is an advantage to the use of UV/H2O2 to simultaneously remove brominated compounds while removing geosmin and 2-MIB. Furthermore, water that contains greater than 0.10 mg/L bromide ion could result in bromate formation with ozonation (Song et al., 1997), and such source waters would be unsuitable for the application of ozone to control geosmin and/or MIB due to bromate formation (Ho et al., 2004; Westerhoff et al., 2006; Peter and Von Gunten, 2007); UV/H2O2 would be the preferred AOP in this case. This research has implications for application of UV at full-scale facilities to control pathogens and contaminants. Although application of UV at doses of approximately 40 mJ/cm2 is considered viable for full-scale drinking water disinfection (Linden et al., 2004), full-scale data for contaminant removal by UV are not readily available. Pilot- and bench-
scale research by Mofidi et al. (2002) demonstrated that in trials using 50 ng/L geosmin and 2-MIB added to source water, a UV dose of 1100 mJ/cm2 with 5.5 mg/L H2O2 reduced both odorants by >90% while a high UV dose of 10 100 mJ/cm2 was required for >90% removal in the absence of H2O2. The authors calculated that at a reasonable dose of 100 mJ/cm2 for disinfection, even with the addition of 5 mg/L H2O2, only 50e60% of the added geosmin and 2-MIB would be removed and the earthy-musty odors would persist. THM and HAA degradation were not investigated. Linden et al. (2004) concur that UV alone at doses compatible with disinfection will not be effective for destruction of organic contaminants. Their bench-scale studies with geosmin, 2-MIB, and pesticides indicated that degradation of organic contaminants would likely require 500e2000 mJ/cm2 and even then, some contaminants would not be removed without addition of H2O2. Our research confirms that degradation of geosmin and 2-MIB occurs at doses of 500e2000 mJ/cm2 with addition of H2O2, and that brominated DBPs are degraded under these conditions whereas perchlorinated DPBs are not well removed.
5.
Conclusions
The results demonstrate that mg/L concentrations of brominated THMs and HAAs can be simultaneously degraded by advanced oxidation when applying 6 mg/L H2O2 and a UV dose effective for removing ng/L concentrations of geosmin and 2-methylisoborneol. Trichloromethane and perchlorinated haloacetic acids could not be effectively degraded under these same advanced oxidation conditions. Geosmin, 2-MIB, and chlorinated DBPs were primarily degraded by the reaction with hydroxyl radicals, generated via the photolysis of H2O2. While direct UV photolysis played a minor role in removing geosmin and 2-methylisoborneol, the perchlorinated DBPs were not degraded by direct photolysis. Brominated DBPs were degraded primarily by direct photolysis, presumably via photo-induced CeBr bond cleavage. The apparent pseudo-first order rate constants for photolysis of the brominated THMs and HAAs were between
w a t e r r e s e a r c h 4 5 ( 2 0 1 1 ) 2 5 0 7 e2 5 1 6
103 to 102 s1 and increased with the number of bromine atoms in the molecule. Tribromomethane and dibromochloromethane were degraded by 99% and 80%, respectively, at the UV dose of 1200 mJ/cm2 with 6 mg/L H2O2, while 90% of the geosmin and 65% of the 2-methylisoborneol were degraded. Tribromoacetic acid and dibromoacetic acid were degraded by 99% and 80% respectively under the same conditions. Implications of this research are that the UV/H2O2 process, when implemented for odor control, can have the additional benefit of DBP elimination, especially in regions where source water bromide concentration leads to high concentrations of brominated DBPs. UV/H2O2 may also be more desirable than ozone in these regions due to possible formation of bromate from bromide ion during ozonation.
Acknowledgments The authors thank Kwater (Korea Water Resources Corporation) which provided a research fellowship for Dr. Chang Hyun Jo and the MILES (Macromolecular Interfaces with Life Science)eIGERT program at Virginia Tech (NSF agreement # DGE0333378) for experimental support. Neither funding agency contributed to the study design; collection, analysis or interpretation of data; decision to submit the paper for publication.
references
Beltran, F.J., Ovejero, G., Acedo, B., 1993. Oxidation of atrazine in water by ultraviolet-radiation combined with hydrogenperoxide. Water Research 27 (6), 1013e1021. Cassan, D., Mercier, B., Castex, F., Rambaud, A., 2006. Effects of medium-pressure UV lamps radiation on water quality in a chlorinated indoor swimming pool. Chemosphere 62 (9), 1507e1513. Chang, P.B.L., Young, T.M., 2000. Kinetics of methyl tert-butyl ether degradation and by-product formation during UV/hydrogen peroxide water treatment. Water Research 34 (8), 2233e2240. Chin, A., Berube, P.R., 2005. Removal of disinfection by-product precursors with ozone-UV advanced oxidation process. Water Research 39 (10), 2136e2144. Cotton, C.A., Collins, J.R., 2006. Dual purpose UV light: Using UV light for disinfection and for taste and odor oxidation. Proceedings of the 2006 Water Quality Technology Conference. American Water Works Association, Denver, CO, USA. Dietrich, A.M., 2006. Aesthetic issues for drinking water. Journal of Water and Health 4 (Suppl. 1), 11e16. Glaze, W.H., Schep, R., Chauncey, W., Ruth, C., Zarnoch, J.J., Aieta, E.M., Tate, C.H., McGuire, M.J., 1990. Evaluating oxidants for the removal of model taste and odor compound from a municipal water-supply. Journal of the American Water Works Association 82 (5), 79e84. Ho, L., Croue, J.P., Newcombe, G., 2004. The effect of water quality and NOM character on the ozonation of MIB and geosmin. Water Science and Technology 49 (9), 249e255. Hyun, K.S., Kim, Y.J., Choi, J.J., 2005. Profiles of THM and HAAs in the process of water treatment plants. Journal of Korean Society Water Science and Technology 13 (1), 55e66.
2515
Jo, C.-H., Dietrich, A.M., 2009. Removal and transformation of odorous aldehydes by UV/H2O2- efficiency and effect. Journal of Water Supply: Research and Technology-AQUA 58 (8), 580e587. Ju¨ttner, F., Watson, S.B., 2007. Biochemical and ecological control of geosmin and 2-methylisoborneol in source waters. Applied and Environmental Microbiology 73 (14), 4395e4406. Klassen, N.V., Marchington, D., McGowan, H.C.E., 1994. H2O2 determination by the I 3 method and by KMnO4 titration. Analytical Chemistry 66 (18), 2921e2925. Krasner, S.W., McGuire, M.J., Jacangelo, J.G., Patania, N.L., Reagan, K.M., Aieta, E.M., 1989. The occurrence of disinfection by-products in US drinking water. Journal of the American Water Works Association 81 (8), 41e53. Krasner, S.W., Weinberg, H.S., Richardson, S.D., Pastor, S.J., Chinn, R., Sclimenti, M.J., Onstad, G.D., Thruston, A.D., 2006. Occurrence of a new generation of disinfection byproducts. Environmental Science & Technology 40 (23), 7175e7185. Kurac, J., Hlatka, A., 1992. Photolysis of waterecarbon tetrachloride and waterechlorobenzene saturated solutions and their two phase system: quantum yields of chloride ions. Journal of Radioanalytical and Nuclear Chemistry 166 (3), 239e250. Kwok, W.M., Zhao, C.Y., Li, Y.L., Guan, X.G., Wang, D.G., Phillips, D.L., 2004. Water-catalyzed dehalogenation reactions of isobromoform and its reaction products. Journal of the American Chemical Society 126 (10), 3119e3131. Liao, C.-H., Gurol, M.D., 1995. Chemical oxidation by photolytic decomposition of hydrogen peroxide. Environmental Science and Technology 29 (12), 3007e3014. Lin, X.F., Guan, X.G., Kwok, W.M., Zhao, C.Y., Du, Y., Li, Y.L., Phillips, D.L., 2005. Water-catalyzed OeH insertion/HI elimination reactions of isodihalomethanes (CH2XeI, where X ¼ Cl, Br, I) with water and the dehalogenation of dihalomethanes in water-solvated environments. The Journal of Physical Chemistry A 109 (6), 981e998. Linden, C., Sharpless, C., Andrews, S., Atasi, K., Korategere, V., Stephan, M., Suffet, I.H., 2004. Innovative UV Technologies to Oxidize Organic and Organoleptic Chemicals. AwwaRF, Denver CO USA. Maruthamuthu, P., Padmaja, S., Huie, R.E., 1995. Rate constants for some reactions of free-radicals with haloacetates in aqueous-solution. International Journal of Chemical Kinetics 27 (6), 605e612. McGivern, W.S., Derecskei-Kovacs, A., North, S.W., Francisco, J.S., 2000. Computationally efficient methodology to calculate CeH and CeX (X ¼ F, Cl, and Br) bond dissociation energies in haloalkanes. The Journal of Physical Chemistry A 104 (2), 436e442. Mezyk, S.P., Helgeson, T., Cole, S.K., Cooper, W.J., Fox, R.V., Gardinali, P.R., Mincher, B.J., 2006. Free radical chemistry of disinfection-byproducts. 1. Kinetics of hydrated electron and hydroxyl radical reactions with halonitromethanes in water. The Journal of Physical Chemistry A 110 (6), 2176e2180. Mofidi, A., Min, J., Palencia, L., Coffey, B., Liang, S., Green, J., 2002. Advanced Oxidation Processes and UV Photolysis for Treatment of Drinking Water. California Energy Commission. http://www.energy.ca.gov/reports/2004-04-02_500-02-019/50002-019_A1.PDF (accessed January 2011). Nickelsen, M.G., Cooper, W.J., Lin, K., Kurucz, C.N., Waite, T.D., 1994. High energy electron beam generation of oxidants for the treatment of benzene and toluene in the presence of radical scavengers. Water Research 28 (5), 1227e1237. Nicole, I., De Laat, J., Dore, M., Duguet, J.P., Sutyh, E., 1991. Degradation of trihalomethanes in dilute aqueous solution by UV irradiation e determination of the quantum yield of photolysis at 253.7 nm. Environmental Technology 12, 21e31. Oppenlander, T., 2003. Photochemical Purification of Water and Air. Wiley-VHC.
2516
w a t e r r e s e a r c h 4 5 ( 2 0 1 1 ) 2 5 0 7 e2 5 1 6
Paradis, N., Hoffman, A., 2006. Mitigation of taste and odour compounds by UV/H2O2 advanced oxidation. In: Proceedings of the 2006 Water Quality Technology Conference. American Water Works Association, Denver, CO, USA November 6e9, 2006. Pereira, V.J., Weinberg, H.S., Linden, K.G., Singer, P.C., 2007. UV degradation kinetics and modeling of pharmaceutical compounds in laboratory grade and surface water via direct and indirect photolysis at 254 nm. Environmental Science and Technology 41 (5), 1682e1688. Peter, A., Von Gunten, U., 2007. Oxidation kinetics of selected taste and odor compounds during ozonation of drinking water. Environmental Science and Technology 41 (2), 626e631. Piriou, P., Devesa, R., De Lalande, M., Glucina, K., 2009. European reassessment of MIB and geosmin perception in drinking water. Journal of Water Supply: Research and TechnologyAQUA 58 (8), 532e538. Piriou, P., Soulet, C., Acero, J.L., Bruchet, A., Von Gunten, U., Suffet, I.H., 2007. Understanding medicinal taste and odour formation in drinking waters. Water Science and Technology 55 (5), 85e94. Qiao, R.P., Li, N., Qi, X.H., Wang, Q.S., Zhuang, Y.Y., 2005. Degradation of microcystin-RR by UV radiation in the presence of hydrogen peroxide. Toxicon 45 (6), 745e752. Rahn, R.O., 2004. Spatial distribution of upper-room germicidal UV radiation as measured with tubular actinometry as compared with spherical actinometry. Photochemistry and Photobiology 80 (2), 346e350. Rahn, R.O., Bolton, J., Stefan, M.I., 2006. The iodide/iodate actinometer in UV disinfection: determination of the fluence rate distribution in UV reactors. Photochemistry and Photobiology 82 (2), 611e615. Rashash, D.M.C., Dietrich, A.M., Hoehn, R.C., 1997. Flavor profile analysis of selected odorous compounds. Journal of the American Water Works Association 89 (4), 131e142. Richardson, S.D., Plewa, M.J., Wagner, E.D., Schoeny, R., DeMarini, D.M., 2007. Occurrence, genotoxicity, and carcinogenicity of regulated and emerging disinfection byproducts in drinking water: a review and roadmap for research. Mutation Research-Reviews in Mutation Research 636, 178e242. Rosenfeldt, E.J., Linden, K.G., 2007. The R-OH, R-UV concept to characterize and the model UV/H2O2 process in natural waters. Environmental Science and Technology 41 (7), 2548e2553. Rosenfeldt, E.J., Melcher, B., Linden, K.G., 2005. UV and UV/H2O2 treatment of methylisoborneol (MIB) and geosmin in water. Journal of Water Supply: Research and Technology-AQUA 54 (7), 423e434.
Royce, A., Stefan, M., 2005. Application of UV in drinking water treatment for simultaneous disinfection and removal of taste and odor compounds. In: Proceedings of the 2005 Water Quality Technology Conference. American Water Works Association, Quebec City, Canada November 6e10, 2005. Rudra, A., Thacker, N.P., Pande, S.P., 2005. Hydrogen peroxide and ultraviolet irradiations in water treatment. Environmental Monitoring and Assessment 109 (1e3), 189e197. Sarathy, S., Mohseni, M., 2010. Effects of UV/H2O2 advanced oxidation on chemical characteristics and chlorine reactivity of surface water natural organic matter. Water Research 44 (15), 4087e4096. Song, R., Westerhoff, P., Minear, R., Amy, G., 1997. Bromate minimization during ozonation. Journal of the American Water Works Association 89 (6), 69e78. Stefan, M.I., Bolton, J.R., 1998. Mechanism of the degradation of 1,4-dioxane in dilute aqueous solution using the UV hydrogen peroxide process. Environmental Science and Technology 32 (11), 1588e1595. United Stated Environmental Protection Agency (USEPA), 2011. National Primary Drinking Water Regulations. http://water.epa. gov/drink/contaminants/index.cfm#List (accessed 01.14.11). Wang, G.S., Hsieh, S.T., Hong, C.S., 2000. Destruction of humic acid in water by UV light e catalyzed oxidation with hydrogen peroxide. Water Research 34 (15), 3882e3887. Watson, S.B., Brownlee, B., Satchwill, T., Hargesheimer, E.E., 2000. Quantitative analysis of trace levels of geosmin and MIB in source and drinking water using headspace SPME. Water Research 34 (10), 2818e2828. Watson, S.B., Brownlee, B., Satchwill, T., McCauley, E., 1999. The use of solid phase microextraction (SPME) to monitor for major organoleptic compounds produced by chrysophytes in surface waters. Water Science and Technology 40 (6), 251e256. Westerhoff, P., Nalinakumari, B., Pei, P., 2006. Kinetics of MIB and geosmin oxidation during ozonation. Ozone-Science and Engineering 28 (5), 277e286. Zhang, X., Talley, J.W., Boggess, B., Ding, G.Y., Birdsell, D., 2008. Fast selective detection of polar brominated disinfection byproducts in drinking water using precursor ion scans. Environmental Science and Technology 42 (17), 6598e6603. Zhao, C.Y., Lin, X.F., Kwok, W.M., Guan, X.G., Du, Y., Wang, D.Q., Hung, K.F., Phillips, D.L., 2005. Water-catalyzed dehalogenation reactions of the isomer of CBr4 and its reaction products and a comparison to analogous reactions of the isomers of di- and trihalomethanes. Chemistrya European Journal 11 (4), 1093e1108.
Available at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
Simultaneous degradation of disinfection byproducts and earthy-musty odorants by the UV/H2O2 advanced oxidation process Chang Hyun Jo a,1, Andrea M. Dietrich a,*, James M. Tanko b a b
Department of Civil and Environmental Engineering, Virginia Polytechnic Institute and State University, VA 24061, USA Department of Chemistry, Virginia Polytechnic Institute and State University, VA 24061, USA
article info
abstract
Article history:
Advanced treatment technologies that control multiple contaminants are beneficial to
Received 27 September 2010
drinking water treatment. This research applied UV/H2O2 for the simultaneous degradation
Received in revised form
of geosmin, 2-methylisoborneol, four trihalomethanes and six haloacetic acids. Experi-
3 February 2011
ments were conducted in de-ionized water at 24 1.0 C with ng/L amounts of odorants
Accepted 4 February 2011
and mg/L amounts of disinfection byproducts. UV was applied with and without 6 mg/L
Available online 12 February 2011
H2O2. The results demonstrated that brominated trihalomethanes and brominated haloacetic acids were degraded to a greater extent than geosmin and 2-methylisoborneol.
Keywords:
Tribromomethane and dibromochloromethane were degraded by 99% and 80% respec-
Hydroxyl radical
tively at the UV dose of 1200 mJ/cm2 with 6 mg/L H2O2, whereas 90% of the geosmin and
Ultraviolet
60% of the 2-methylisoborneol were removed. Tribromoacetic acid and dibromoacetic acid
Geosmin
were degraded by 99% and 80% respectively under the same conditions. Concentrations of
2-Methylsioborneol
trichloromethane and chlorinated haloacetic acids were not substantially reduced under
Trihalomethanes
these conditions and were not effectively removed at doses designed to remove geosmin
Haloacetic acids
and 2-methylisoborneol. Brominated compounds were degraded primarily by direct photolysis and cleavage of the CeBr bond with pseudo first order rate constants ranging from 103 to 102 s1. Geosmin and 2-methylisoborneol were primarily degraded by reaction with hydroxyl radical with direct photolysis as a minor factor. Perchlorinated disinfection byproducts were degraded by reaction with hydroxyl radicals. These results indicate that the UV/H2O2 can be applied to effectively control both odorants and brominated disinfection byproducts. ª 2011 Elsevier Ltd. All rights reserved.
1.
Introduction
Advanced oxidation processes (AOP) in water treatment involve hydroxyl radicals ($OH) as intermediates. UV-based AOPs essentially mimic the photo-initiated oxidation processes in natural systems, such as sun light on surface
water or in the atmosphere (Oppenlander, 2003). Ultraviolet (UV) irradiation is well established for disinfection of water. The UV/H2O2 provides oxidation through generation of hydroxyl radicals by photolysis of H2O2 (Liao and Gurol, 1995; Rudra et al., 2005). This process degrades organic contaminants, including recalcitrant odorants such as geosmin and
* Corresponding author. E-mail address: [email protected] (A.M. Dietrich). 1 Present address: K-water, Daejeon, Korea. 0043-1354/$ e see front matter ª 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2011.02.006
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2-methylisoborneol (2-MIB), mainly by the reaction with hydroxyl radicals and partially by direct UV photolysis of the contaminant (Beltran et al., 1993; Stefan and Bolton, 1998; Rosenfeldt et al., 2005; Rudra et al., 2005; Paradis and Hoffman, 2006; Rosenfeldt and Linden, 2007). Solution-phase AOPs are advantageous for contaminant removal because residual solids are not produced such as would occur with spent activated carbon from an adsorption process. Recently, UV systems labeled as “dual purpose” were applied to fullscale water treatment plants. These dual systems combine low intensity UV for disinfection and high intensity UV for both disinfection and advanced oxidation of odorants. Their operation is limited by both UV transmittance and water quality constraints because low UV transmittance attenuates UV light available to produce hydroxyl radicals from H2O2, and carbonates scavenge hydroxyl radicals (Ho et al., 2004; Cotton and Collins, 2006). The optimum H2O2 dose in the UV/H2O2 process should be empirically determined because excess H2O2 can be a hydroxyl radical scavenger (Wang et al., 2000). Typical doses applied for taste-and-odor control in full-scale treatment plants vary from >0 to 11 mg/L H2O2 (Royce and Stefan, 2005). Geosmin and 2-MIB are earthy-musty odorants found in surface water and drinking water. They are produced by cyanobacteria or actinomycetes, commonly in the warm summer and early fall seasons (Ju¨ttner and Watson, 2007). Concentrations in drinking water can range from a few to >100 ng/L; to minimize consumer complaints and produce acceptable quality water, concentrations of these odorants should be brought to below their odor threshold levels of 1e10 ng/L (Rashash et al., 1997; Dietrich, 2006; Piriou et al., 2009). Although not regulated for health effects, esthetic guidelines of 10 ng/L geosmin and 10 ng/L 2-MIB have been established in Korea and as secondary standards in Japan. These compounds are difficult to remove by conventional water treatment processes; thus activated carbon or AOPs are frequently used for their control during full-scale treatment. 2-MIB has repeatedly proven the more difficult to remove and reacts slower with hydroxyl radicals compared to geosmin (Glaze et al., 1990; Westerhoff et al., 2006; Peter and Von Gunten, 2007). In addition to problematic taste and odor compounds, disinfection byproducts (DBPs) are another major issue in drinking water quality. Most organic DBPs form from the reaction with humic or fulvic acid and disinfectants. The major classes of regulated DBPs in drinking water are trihalomethanes (THMs), which are regulated at a total concentration of 80 mg/L, and haloacetic acids (HAAs) which are regulated at a total concentration of 60 mg/L (USEPA, 2011). Chlorinated DBPs including trichloromethane and mono, di-, and trichloroacetic acid are the most prevalent (Krasner et al., 2006), but the occurrence of brominated and iodinated DBPs is both pervasive and problematic because bromo-and iodocontaining DBPs are more toxic than their chlorinated analogs (Krasner et al., 2006; Richardson et al., 2007). Brominated DBPs can regionally constitute > 40% THMs and 10e25% HAAs (Krasner et al., 1989; Hyun et al., 2005). In addition to regulated DBPs, over 90 other brominated DBPs were identified by GCeMS (Richardson et al., 2007) and additional polar brominated DBPs were detected by LCeMS (Zhang et al., 2008).
Advanced treatment technologies that control multiple contaminants and provide disinfection in full-scale water treatment are immensely beneficial to the drinking water industry. Studies report that UV-based AOPs control DBP precursors and consequently reduce DBP level in finished water (Wang et al., 2000; Chin and Berube, 2005; Sarathy and Mohseni, 2010). However, many water utilities apply prechlorination to control taste/odor or iron/manganese or ammonia nitrogen, and to obtain required CT values. Prechlorination occurs prior to coagulation and produces a variety of DBPs while AOPs typically occur after filtration to increase UV transmission, which is also after formation of DBPs. There have been only a few studies on the removal of DBPs by UV/H2O2 and these reports are contradictory. Rudra et al. (2005) reported over 90% degradation of THMs at high UV (17 000 mJ/cm2) and 0.1% H2O2 dose. In pilot-scale and bench-scale studies with natural surface waters, THMs increased at lower UV/H2O2 doses and decreased with higher UV/H2O2 doses, and HAAs decreased for two samples and increased for one sample (Paradis and Hoffman, 2006). Second order rate constants for the reaction of DBPs and odorous contaminants with hydroxyl radical have been measured (Mezyk et al., 2006; Westerhoff et al., 2006; Peter and Von Gunten, 2007). Few studies have addressed the possibility and mechanisms of DBPs degradation by UV/H2O2 process in aqueous phase. Two mechanisms are associated with UV/H2O2 oxidation of contaminants (Scheme 1). Photolysis of H2O2 produces hydroxyl radicals that subsequently react with contaminants, generally by abstracting a hydrogen or by adding to an unsaturated site. Direct UV photolysis of the contaminant can occur and result in bond homolysis and radical generation; the resulting carbon-centered radicals subsequently are oxidized by reaction with H2O2, O2, and contaminants. Direct UV photolysis can contribute substantially to contaminant degradation. Geosmin and 2-MIB were decreased by 40% and 20%, respectively, with a UV dose of 1700 mJ/cm2 in raw water blends with 1.7e2.3 mg/L TOC, 105 mg/L alkalinity as CaCO3, and pH 8.0 to 8.3 (Rosenfeldt et al., 2005). Other source water algal-contaminants degraded by UV include microcystin (Qiao et al., 2005) and conjugated odorous aldehydes (Jo and Dietrich, 2009). In regard to UV photolysis of THMs, only the brominated THMs were photolyzed and quantum yield of the photolysis for all brominated THMs was 0.43 with a 253.7 nm lamp (Nicole et al.,
Direct Photolysis hv
R3C-X
R 3C
+
X
Hydroxyl radical generation (H2O2 photolysis; hydrogen abstraction) hv
H 2O 2
HO
+
R-H
2 HO
H2O
+
R
Scheme 1 e Mechanisms for oxidation by UV alone and with H2O2. (“X” refers to a halogen).
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1991). In the same research, polybrominated THMs were shown to be photolyzed faster than monobrominated compounds. Concentrations of tribromomethane and chlorodibromomethane in chlorinated swimming pool water decreased significantly with medium pressure UV irradiation of 145 mJ/cm2 (Cassan et al., 2006). Polyhalomethanes such as tribromomethane (CHBr3) and carbon tetrabromide (CBr4) were photolyzed by a proposed water-catalyzed OeH insertion/HBr elimination (Kwok et al., 2004; Lin et al., 2005; Zhao et al., 2005). For the direct UV photolysis and hydroxyl radical reactions of organic compounds such as geosmin/2-MIB and pharmaceutical compounds by UV/H2O2, pseudo-first order reaction models have been proposed because under constant irradiation, the steady-state concentration of the hydroxyl radical is approximately constant with respect to pollutant concentration because of the relatively high concentration of H2O2 (Rosenfeldt et al., 2005; Pereira et al., 2007). d½C ¼ k0 ½C dt Where, k0 ¼ k0d þ k0i k0 ¼ the observed pseudo-first order rate constant (s1) k0d ¼ the measured pseudo-first order rate constant of direct photolysis (s1) k0i ¼ the measured pseudo-first order rate constant of the reaction with $OH (s1) Second order reaction rate constants for reaction of hydroxyl radical with geosmin have been reported to range from 8.2 109 to 1.4 1010 M1s1, and for 2-MIB from 5.1 109 to 8.1 109 M1s1 (Glaze et al., 1990; Westerhoff et al., 2006; Peter and Von Gunten, 2007), which are nearly an order of magnitude more than for the reaction of hydroxyl radical with THMs (0.7 107e1.5 108 M1s1) or chlorinated HAAs (6 107e1.0 108 M1s1). The rate constant for reaction of tribromomethane with HO$ is greater than trichloromethane by a factor of 10 (Maruthamuthu et al., 1995; Mezyk et al., 2006). This paper summarizes our results from the investigation of the simultaneous degradation of odorants and DBPs at concentrations and conditions representative of those used for the removal of recalcitrant odorants in drinking water using UV/H2O2 based on typical H2O2 dose used in full-scale water treatment plant. The objectives of this research were to: 1) determine the degree of degradation of THMs and HAAs with respect to degradation of geosmin and 2-MIB, and 2) evaluate the role of UV photolysis and hydroxyl radical reaction in this degradation.
2.
Methods and materials
2.1.
Apparatus
Experiments were performed with a Rayonet RPR-100 photochemical reactor equipped with 253.7 nm wavelength UV lamps of 7.2 mW/cm2 total intensity, and quartz reaction vessels. UV dose was confirmed with the iodide/iodate
actinometer (Rahn, 2004; Rahn et al., 2006). Samples were completely mixed and headspace free while being irradiated with UV. A reaction temperature of 24 1.0 C was maintained by an electric fan set under the sample, and ice.
2.2.
Reagents and sample preparation
Samples for UV/H2O2 irradiation were prepared directly in deionized water (Nanopure) using individual neat compounds for DBPs; trichloromethane (99%, Fisher Scientific), tribromomethane (99%, Acros Organics), chloroacetic acid (MCAA) (99%, Aldrich), dichloroacetic acid (DCAA) (99%, SigmaeAldrich), trichloroacetic acid (TCAA) (99%, Alfa Aesar), bromoacetic acid (MBAA) (99%, SigmaeAldrich), dibromoacetic acid (DBAA) (99%, Fluka), tribromoacetic acid (TBAA) (99%, Acros Organics), tetrachloromethane (>99%, Acros Organics) and tetrabromomethane (>99%, Acros Organics). To prepare individual and mixed samples of HAAs to react with UV/H2O2, neat HAA compounds were dissolved in water because the commercial HAA standard mixture was dissolved in methyl tert-butyl ether (MTBE) and this solvent scavenges hydroxyl radicals (second order rate constant, k ¼ 3.9 109 M1s1) (Chang and Young, 2000). For quantitative chromatographic analyses, commercial standards solutions of THMs in methanol and HAAs in methyl-tert-butyl ether were purchased from Ultra Scientific. Geosmin (200 mg/L) and 2-MIB (100 mg/L) were purchased from Supelco as solutions in methanol and diluted in water to the appropriate concentration. Typical odorant solutions used in this research (Table 1) consistently contained about 2 mL or 1.5 mg/L (50 mM) methanol from dilution of the concentrated odorant solutions. Hydrogen peroxide (30%, Fisher) was
Table 1 e Typical initial concentrations of compounds in the research. Compounds
Typical concentrations mg/L
mM
a
Odorants Geosmin 2-MIB THMsb Trichloromethane Bromodichloromethane Dibromochloromethane Tribromomethane Tetrahalomethanes Carbon tetrachloride Carbon tetrabromide HAAsb Chloroacetic acid (MCAA) Dichloroacetic acid (DCAA) Trichloroacetic acid (TCAA) Bromoacetic acid (MBAA) Dibromoacetic acid (DBAA) Tribromoacetic acid (TBAA) Hydrogen Peroxide
0.04e0.2 0.1e0.3
0.0002e0.001 0.0006e0.002
60e500 90 80 80e550
0.5e4.2 0.5 0.4 0.3e2.2
350 1000
2.3 3.0
270 190 180 200 190 160 6000
2.9 1.5 1.1 1.4 0.9 0.5 176.5
a Added with z1.5 mg/L methanol. b It should be noted that these concentrations are slightly higher than those usually found in actual drinking waters.
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2.4.
Statistical analyses
Microsoft Excel (2007) was applied for data manipulation and calculation of apparent first order rate constant and standard errors.
3.
Results
3.1.
UV absorbance
To investigate the relative role of hydroxyl radical production versus direct UV photolysis of the organic contaminants in
3.2. Degradation of odorants and DBPs at typical concentrations found in drinking water Concentrations of ng/L geosmin/2-MIB and mg/L DBPs were prepared in mixed aqueous samples and contaminant removals were simultaneously followed under the same UV/ H2O2 and solution conditions. The solutions contained about 1.5 mg/L methanol, because it was the solvent used to dissolve geosmin and 2-MIB; methanol is known to act as a hydroxyl radical scavenger (Nickelsen et al., 1994). In practice, a utility would apply a UV/H2O2 dose for geosmin/2-MIB degradation that met their treatment need to control the odor problem. Thus, although the water being treated may contain other contaminants and radical scavengers that could absorb at UV 253.7 nm or react with hydroxyl radicals, the applied UV/H2O2 dose would be determined by treatment goals for removing geosmin and 2-MIB.
1.20 1.00 0.80 0.60 0.40 0.20 0.00
A
H 2O ge 2 os m in 2M IB C H Br 3 C H C l3 C Br 4 C C l4 M BA A D BA A TB AA M C AA D C AA TC AA
1200 1000 800 600 400 200 0
B
H 2O ge 2 os m in 2M IB C H Br 3 C H C l3 C Br 4 C Cl 4 M BA A D BA A TB AA M C AA D C A A TC AA
Geosmin and 2-MIB were measured by solid-phase microextraction (SPME, Supelco) as reported by other researchers (Watson et al., 1999, 2000). Compounds partitioned from water were sorbed onto an SPME fiber (65 mm, PDMS/DVB) for 10 min at 60 C. The SPME fiber was injected in the splitless mode into the GC/MS (Agilent 5973) at 220 C and desorbed for 2 min. An Rtx-5Sil column (30 m, 0.25 mm ID) with He carrier gas and a temperature program of 60 Ce180 C by 15 C/min was used. Approximate retention times were 5.4 min for 2-MIB and 7.9 min for geosmin; m/z values of 112, 125, 182 for geosmin and m/z values of 95, 108, 168 for 2-MIB were monitored for qualitative analysis and m/z values of 112 or 95 were used for quantitative analysis for geosmin and 2-MIB respectively in selective ion mode. Concentrations of disinfection byproducts were determined using standardized methods. THMs were measured based on Standard Method 6232.D by purge/trap (Tekmar 3000) and GC (Tremetrics 9001) with DB-624 column (J & W) and ECD detector. GC temperature was initially maintained at 45 C for 3 min, and then increased by 11 C/min up to 200 C. HAAs were determined by liquideliquid extraction method (EPA Method 552.2) and GC (HP 5890) with an ECD detector. The injector temperature was 210 C and the initial oven temperature was set to 35 C and increased up to 140 C. UV adsorption was measured on a UV/vis spectrophotometer (Beckman, DU640). H2O2 concentration was determined by triiodide (I 3 ) titration method (Klassen et al., 1994).
Molar extinction
Chemical analyses
coefficient (M -1 cm -1 )
2.3.
this UV/H2O2 process, molar extinction coefficients were measured (Fig. 1A). The brominated compounds, geosmin, and 2-MIB had at least two order of magnitude higher molar extinction coefficients than chlorinated compounds, and one order of magnitude lower than H2O2. Absolute molar extinction coefficients, however, can be misleading. At typical concentrations used in this research (Table 1), only H2O2 and the brominated DBPs would absorb an appreciable amount of UV as shown by their absorptions relative to H2O2 (Fig. 1B). The relative absorbance of geosmin, 2-MIB, and any of the chlorinated compounds are nil. The UV absorbances of the brominated DBPs are sufficiently high that a plausible mechanism for brominated DBPs elimination is by direct UV photolysis (CeBr bond cleavage). At 6 mg/L H2O2, hydrogen peroxide absorbs most of the UV at 253.7 nm, which indicates that hydroxyl radicals can be produced from UV photolysis of H2O2 under these conditions. In this research, the absorbance of 6 mg/L H2O2 at 253.7 nm was ca. 0.035 in the experimental apparatus.
Relative absorbance
diluted to desired concentrations of 6 mg/L which was selected based on typical concentration range in pilot-scale study (Paradis and Hoffman, 2006), and added into the samples immediately before UV irradiation. Typical initial concentrations of compounds used in the research are shown in Table 1; it should be noted that these concentrations are slightly higher than those usually found in actual drinking waters. Molar extinction coefficients at 253.7 nm were determined for each compound dissolved in distilled water; these concentrations were measured: 34 mg/L H2O2, 0.4 mg/L geosmin, 0.1 mg/L 2-MIB, 340 mg/L for CHBr3, CHCl3, or CBr4, 150 mg/L CCl4, and 1000 mg/L for MBAA, DBAA, TBAA, MCAA, DCAA, and TCAA. Absorptions relative to H2O2 were calculated by applying the concentrations in Table 1, and using 100 ng/L for geosmin and 2-MIB, and 300 mg/L for CHCl3 and CHBr3.
Fig. 1 e Molar extinction coefficients (A) and relative absorption (relative to H2O2 [ 1) (B) of odorant and DBP compounds and hydrogen peroxide measured at 253.7 nm and at the concentrations used in this research. Concentrations used for these measurements are provided in the methods and materials.
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3.2.1.
Geosmin and 2-MIB
The geosmin and 2-MIB results show 90 and 65% degradation, respectively, with a UV dose of 1200 mJ/cm2 and 6 mg/L H2O2. This dose was selected as it was sufficient to reduce the initial geosmin concentration of 40e43 ng/L to 4 ng/L (Figs. 2 and 3), which is below its guideline and upper odor threshold value of 10 ng/L. Under identical conditions, but in the absence of H2O2, only about 20% was degraded with UV photolysis (Fig. 2).Using the results in Figs. 2 and 3, apparent pseudo first order rate constants and half-lives for degradation of these odorants can be calculated (Table 2). The apparent rate constants for degradation of both geosmin and 2-MIB are approximately 3e7 greater when H2O2 is present. These results suggest that geosmin and 2-MIB concentrations are mainly reduced by reaction with hydroxyl radical (formed by photolysis of H2O2) though a small amount may also degraded by direct photolysis.
3.2.2.
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Fig. 3 e Photolysis of geosmin, 2-MIB and THMs in the presence of 6 mg/L H2O2. Initial concentration (C0): geosmin [ 43 ng/L, 2-MIB [ 100 ng/L, trichloromethane [ 63 mg/L, bromodichloromethane [ 94 mg/L, dibromochloromethane [ 76 mg/L, tribromomethane [ 82 mg/L, methanol z1.5 mg/L.
Geosmin/MIB and THMs
3.2.2.1. THMs degradation compared to geosmin/2-MIB by UV/H2O2. Brominated THMs were degraded better than trichloromethane at the UV/H2O2 dose effective for removing geosmin/2-MIB (Fig. 3). Tribromomethane and dibromochloromethane were degraded by 99% and 80%, respectively, at the UV dose of 1200 mJ/cm2 and 6 mg/L H2O2, while 90% of the geosmin and 65% of the 2-MIB were degraded at this dose. The brominated THMs with higher numbers of bromine atoms were degraded faster than ones with lower numbers of bromine or trichloromethane, which for all practical purposes, was not degraded by UV/H2O2 (Fig. 3). Tribromomethane was degraded faster than either geosmin/2-MIB. Another approach to determining an effective UVeH2O2 dose from the data in Fig. 3 is to use the guideline of 10 ng/L 2-MIB that has been established in Japan and Korea. Accordingly, a UV dose of 2700 mJ/cm2 and 6 mg/L H2O2 was required to remove 90% of the more recalcitrant odorant, 2-MIB, to a final concentration of 10 ng/L 2-MIB. At this UV/H2O2 dose for 90% removal of 2-MIB and the initial concentrations used for this experiment, the final concentrations of the other contaminants were: 0.8 ng/L geosmin (98% degradation);
Fig. 2 e Photolysis of geosmin and 2-MIB in the presence and absence of 6 mg/L H2O2. Initial concentration (C0): geosmin (no H2O2) [ 40 ng/L, geosmin (H2O2 6 mg/ L) [ 183 ng/L, 2-MIB (no H2O2) [ 108 ng/L, 2-MIB (6 mg/L H2O2) [ 306 ng/L, methanol z1.5 mg/L.
47 mg/L trichloromethane (26% degradation); 44 mg/L monobromodichloromethane (53% degradation); 0.6 mg/L dibromochloromethane (99% degradation); and <0.1 mg/L tribromomethane (>99.9% degradation). Although the THM concentrations used in this experiment were higher than concentrations typical for drinking water, the UVeH2O2 conditions applied in this research to degrade 2-MIB below its esthetic guideline would also be effective to eliminate dibromochloromethane and tribromomethane to below their regulatory limits.
3.2.2.2. Direct UV photolysis of brominated THMs. To investigate the contribution of direct UV photolysis, an aqueous mixture of THMs was reacted with UV in the absence of H2O2. Using the data in Figs. 3 and 4, apparent pseudo first order rate constants and half-lives were calculated (Table 3). The data clearly show that: a) the apparent rate constants for the brominated THMs were substantially faster than for trichloromethane, b) the rate constant for the brominated THMs increased with the number of bromine atoms in the molecule, c) the rate constants for the brominated THMs were identical in the absence and presence of H2O2. These observations make sense because the CeBr bond serves as the active chromophore which is effectively cleaved by UV (i.e., CeBr þ hn / C$ þ Br$) and chlorine-only containing compounds do not absorb appreciably at 253.7 nm, and thus, cannot be eliminated by direct photolysis.
3.2.2.3. Degradation mechanism for THMs. To further investigate the role of hydrogen abstraction and its effects on THM degradation, the reaction of carbon tetrachloride (CCl4) and carbon tetrabromide (CBr4) were compared to each other and to their corresponding trihalomethane analogs. All halomethane solutions were prepared by dissolving neat compounds in de-ionized water then individually treating with UV/H2O2. Because they do not possess abstractable hydrogens, carbon tetrachloride and carbon tetrabromide are effectively non-reactive toward hydroxyl radical and can only be degraded by direct photolysis. Both the tri- and tetrabrominated methanes were degraded faster than their
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Table 2 e Apparent first order rate constants for photo-induced removal of geosmin and 2-MIB in the presence and absence of 6 mg/L H2O2. hn/H2O2, UV 253.7 nm
Contaminant
kapp, s1 Geosmin
2-MIB
a b c d
1.6 1.9 1.6 7.2 6.8 6.0
(0.3) (0.3) (0.3) (0.2) (0.3) (0.5)
kapp, s1
t½, s 2 b
10 102 102 103 103 103
c d b c d
krela
hn only, UV 253.7 nm
1.2 (0.6) 10
43.3 36.5 43.3 96.3 101.9 115.5
t½, s 3 b
1.1 (0.1) 103
b
577.6
14.1
630.1
6.1
Relative rate constant: mean kapp(hn/H2O2)/kapp(hn only). Based upon data from Fig. 2; standard error of slopes provided. Based upon data from Fig. 3; standard error of slopes provided. Based upon data from Fig. 6; standard error of slopes provided.
chlorinated analogs as shown in Fig. 5. The pseudo first order rate constant for degradation of CCl4 was 5.6 (0.4) 103 s1 and for CBr4 it was 8.4 (0.8) 102 s1 (calculated from the data in Fig. 5). Rate constants for degradation of CX4 (X ¼ Cl or Br) were greater than CHX3, even though the latter possesses a hydrogen atom that can be abstracted by hydroxyl radical. These results further confirm that the different degradation rates in UV/H2O2 between chlorinated and brominated THMs results from the different UV photolysis rates and not from the hydrogen abstraction by hydroxyl radicals.
3.2.3.
Geosmin/2-MIB and HAAs
3.2.3.1. HAAs degradation compared to geosmin/2-MIB by UV/ H2O2. Treatment with UV/H2O2 degraded brominated HAAs faster than chlorinated HAAs (Fig. 6) in distilled water that also contained odorants. Applying a UV dose of 1200 mJ/cm2 and 6 mg/L H2O2 as was done for the geosmin/2-MIB/THM data in Fig. 3, the geosmin concentration was reduced to 7 ng/L (96% removal), while TBAA and DBAA were degraded by 99% and 80% respectively. Chlorinated HAAs with no bromine atoms were barely degraded by a UV doses of 0e4300 mJ/cm2 and 6 mg/L of H2O2. Brominated HAAs degradation rates increased in proportion to the number of bromine atoms in
Fig. 4 e Photolysis of brominated THMs in the presence and absence of H2O2. Initial concentration (C0): trichloromethane [ 63 mg/L, bromodichloromethane [ 94 mg/ L, dibromochloromethane [ 76 mg/L, tribromomethane [ 82 mg/L.
the molecule. Consequently, tribromoacetic acid had the highest degradation rate among all HAAs and geosmin/2-MIB.
3.2.3.2. UV photolysis of brominated HAAs. TBAA, DBAA, and MBAA were treated with UV in the absence of H2O2 to investigate the contribution of direct UV photolysis as the primary mechanism for degradation of brominated HAAs. As shown in Fig. 7, there is not a huge difference in the degradation rates for MBAA, DBAA and TBAA between UV photolysis and UV/ H2O2 process e in fact, the rates in the presence of H2O2 appear to be slightly lower. This becomes clearer when the apparent rate constants and half-lives are compared (Table 4). The fact that the rate constants did not increase in the presence of H2O2 suggests that brominated HAAs are degraded mainly by UV photolysis in UV/H2O2 process, rather than via reaction with $OH.
4.
Discussion
For ng/L odorant concentrations and mg/L THM and HAA concentrations typical of drinking water, UV/H2O2 provided substantial simultaneous degradation of geosmin, 2-MIB, brominated THMs and brominated HAAs; however, chlorinated DBPs experienced only minor degradation under the same conditions. This research confirmed the results of Rosenfeldt et al. (2005) that geosmin and 2-MIB were mainly eliminated by the reaction with hydroxyl radicals in the UV/ H2O2 process and that direct UV photolysis is minor. This research concurs with that of Rudra et al. (2005) who reported removal of THMs by UVeH2O2 and also Nicole et al. (1991) who reported direct UV as the removal mechanism for brominated THMs. Brominated HAAs were demonstrated to be degraded by direct UV photolysis and chlorinated HAAs by UV/H2O2 which is consistent with the results of Paradis and Hoffman (2006), who determined that UVeH2O2 decreased HAAs in two out of three samples. For halogenated THMs, a possible first reaction step in UV/ H2O2 advanced oxidation is either hydrogen abstraction by hydroxyl radical or carbonehalogen bond cleavage by direct UV photolysis. Bond dissociation energies of the CeH bond in trichloromethane (CHCl3) and tribromomethane (CHBr3) are very close, 100.0 and 99.9 kcal/mol, respectively (McGivern
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Table 3 e Apparent first order rate constants for photo-induced removal of brominated and chlorinated THMs in the presence and absence of 6 mg/L H2O2. Contaminant
hn/H2O2, UV 253.7 nm kapp, s1
CHCl3 CHBrCl2 CHBr2Cl CHBr3
1.5 (0.1) 2.2 (0.1) 1.8 (0.1) 1.0 (0.1) 1.3 (0.2) 2.9 (0.3) 2.7 (0.4)
3 b
10 103 b 103 c 102 b 102 c 102 b 102 c
krela
hn only, UV 253.7 nm t½, s
kapp, s1
t½, s
462.1 315.0 385.1 69.3 53.3 23.9 25.7
e 2.1 (0.2) 103 c
330.0
e 1.0
1.0 (0.1) 102 c
69.3
1.2
2.8 (0.3) 102 c
24.8
1.0
a Relative rate constant: mean kapp(hn/H2O2)/kapp(hn only). b Based upon data from Fig. 3; standard error of slopes provided. c Based upon data from Fig. 4; standard error of slopes provided.
et al., 2000) and thus, do not explain the faster degradation of brominated DBPs compared to perchlorinated DBPs. Carbonebromine cleavage due to UV photolysis is the likely mechanism of faster degradation of brominated DBPs. This is
Fig. 5 e Photolysis of halogenated methanes in the absence of H2O2. Initial concentration (C0): trichloromethane [ 514 mg/L, tribromomethane [ 523 mg/L, carbon tetrachloride [ 343 mg/L, carbon tetrabromide [ 926 mg/L.
Fig. 6 e Photolysis of geosmin, 2-MIB and HAAs in the presence of 6 mg/L H2O2. Initial concentration (C0): geosmin [ 183 ng/L, 2-MIB [ 306 ng/L, bromoacetic acid [ 202 mg/L, dibromoacetic acid [ 190 mg/L, tribromoacetic acid [ 161 mg/L, chloroacetic acid [ 271 mg/L, dichloroacetic acid [ 191 mg/L, trichloroacetic acid [ 176 mg/L, methanol z1.5 mg/L.
supported by the higher strengths of CeCl bonds than CeBr bonds in trichloromethane and tribromomethane, 80.1 and 70.4 kcal/mol, respectively (McGivern et al., 2000). The energy of a UV photon at 253.7 nm is about 113 kcal/mol, which is sufficient to break both the CeBr bond (Kwok et al., 2004) and CeCl (Kurac and Hlatka, 1992). The higher molar absorption coefficients of brominated DBPs (Fig. 1) and the equal rates of degradation of brominated THMs by UV photolysis with and without H2O2 (Fig. 4 and Table 3) indicate that brominated compounds absorb more UV and thus are more susceptible to photolysis than chlorinated compounds. Although the CeCl is weak, the chlorinated compounds are not degraded by direct UV photolysis because the CeCl chromophore has a low extinction coefficient, i.e., the molecule does not absorb sufficient light to be degraded by this mechanism. The degradation rates of the brominated DBPs increased with the number of bromine atoms present in the molecule. Tribromomethane and tribromoacetic acid were eliminated the fastest among the THMs and HAAs (Figs. 4 and 7, and Table 4). In this research, brominated THMs and HAAs were shown to be eliminated mainly by UV photolysis in the UV/ H2O2 process. Previously, Nicole et al. (1991) demonstrated that UV photolysis of CHBr3, CHBr2Cl, and CHCl2Br at 253.7 nm in aqueous solution led to complete conversion to the corresponding halide. Regarding UV (253.7) photolysis of tribromomethane, a water-catalyzed mechanism was proposed in which isotribromomethane isomerized from
Fig. 7 e Photolysis of brominated HAAs in the presence and absence of 6 mg/L H2O2. Initial concentration (C0): bromoacetic acid [ 202 mg/L, dibromoacetic acid [ 190 mg/L, tribromoacetic acid [ 161 mg/L.
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Table 4 e Apparent first order rate constants for photo-induced removal of chlorinated and brominated acetic acids in the presence and absence of H2O2. (explanation of acronyms: AA [ acetic acid; B [ bromo; C [ chloro; M [ mono; D [ di; T [ tri). hn/H2O2, UV 253.7 nm
Contaminant
kapp, s1 MBAA DBAA TBAA MCAA DCAA TCAA
1.1 1.3 6.4 5.8 5.0 2.9
(0.1) (0.1) (0.4) (2.6) (1.2) (2.3)
103 102 102 104 104 104
kapp, s1
t½, s b b b b b b
krela
hn only, UV 253.7 nm
2.1 (0.1) 103 1.5 (0.1) 102 7.4 (1.1) 102 e e e
630.9 53.3 10.8 1195 1390 2390
t½, s c c c
330.0 46.2 9.4
0.52 0.87 0.86
a Relative rate constant: mean kapp(hn/H2O2)/kapp(hn only). b Based upon data from Fig. 6; standard error of slopes provided. c Based upon data from Fig. 7, standard error of slopes provided.
tribromomethane reacts with water molecule resulting in OH insertion and HBr elimination. The major products from tribromomethane were carbon monoxide and three equivalents of HBr for the major route and formic acid plus three equivalents of HBr for the minor route (Kwok et al., 2004). Brominated compounds are increasingly identified as toxic contaminants in drinking water. In addition to the regulated THMs and HAAs, other bromine-containing disinfection byproducts in drinking water include halonitromethanes, 3e4 carbon haloacids, haloaldehydes, haloketones, haloamides, and halogenated furanones (Krasner et al., 2006). Non-DBP brominated compounds of health concern also occur in source and treated water including the regulated volatile organic compound methyl bromide, brominated flame retardants, and odorous bromophenols (Piriou et al., 2007). In this research the UV/H2O2 process reduced the concentrations of geosmin and 2-MIB in distilled water. This is consistent with the results of Rosenfeldt et al. (2005), who reported degradation of geosmin and 2-MIB in distilled water, and also in source waters and treated drinking waters which possessed other reactants for UV and hydroxyl radicals including natural organic matter and bicarbonate/carbonate ions. The degradation of geosmin, 2-MIB, and perchlorinated DBPs depend on reaction with hydroxyl radicals. The degradation of brominated compounds is primarily through direct UV photolysis that would consume few hydroxyl radicals and would not be affected by the presence of hydroxyl radical scavengers. This is an advantage to the use of UV/H2O2 to simultaneously remove brominated compounds while removing geosmin and 2-MIB. Furthermore, water that contains greater than 0.10 mg/L bromide ion could result in bromate formation with ozonation (Song et al., 1997), and such source waters would be unsuitable for the application of ozone to control geosmin and/or MIB due to bromate formation (Ho et al., 2004; Westerhoff et al., 2006; Peter and Von Gunten, 2007); UV/H2O2 would be the preferred AOP in this case. This research has implications for application of UV at full-scale facilities to control pathogens and contaminants. Although application of UV at doses of approximately 40 mJ/cm2 is considered viable for full-scale drinking water disinfection (Linden et al., 2004), full-scale data for contaminant removal by UV are not readily available. Pilot- and bench-
scale research by Mofidi et al. (2002) demonstrated that in trials using 50 ng/L geosmin and 2-MIB added to source water, a UV dose of 1100 mJ/cm2 with 5.5 mg/L H2O2 reduced both odorants by >90% while a high UV dose of 10 100 mJ/cm2 was required for >90% removal in the absence of H2O2. The authors calculated that at a reasonable dose of 100 mJ/cm2 for disinfection, even with the addition of 5 mg/L H2O2, only 50e60% of the added geosmin and 2-MIB would be removed and the earthy-musty odors would persist. THM and HAA degradation were not investigated. Linden et al. (2004) concur that UV alone at doses compatible with disinfection will not be effective for destruction of organic contaminants. Their bench-scale studies with geosmin, 2-MIB, and pesticides indicated that degradation of organic contaminants would likely require 500e2000 mJ/cm2 and even then, some contaminants would not be removed without addition of H2O2. Our research confirms that degradation of geosmin and 2-MIB occurs at doses of 500e2000 mJ/cm2 with addition of H2O2, and that brominated DBPs are degraded under these conditions whereas perchlorinated DPBs are not well removed.
5.
Conclusions
The results demonstrate that mg/L concentrations of brominated THMs and HAAs can be simultaneously degraded by advanced oxidation when applying 6 mg/L H2O2 and a UV dose effective for removing ng/L concentrations of geosmin and 2-methylisoborneol. Trichloromethane and perchlorinated haloacetic acids could not be effectively degraded under these same advanced oxidation conditions. Geosmin, 2-MIB, and chlorinated DBPs were primarily degraded by the reaction with hydroxyl radicals, generated via the photolysis of H2O2. While direct UV photolysis played a minor role in removing geosmin and 2-methylisoborneol, the perchlorinated DBPs were not degraded by direct photolysis. Brominated DBPs were degraded primarily by direct photolysis, presumably via photo-induced CeBr bond cleavage. The apparent pseudo-first order rate constants for photolysis of the brominated THMs and HAAs were between
w a t e r r e s e a r c h 4 5 ( 2 0 1 1 ) 2 5 0 7 e2 5 1 6
103 to 102 s1 and increased with the number of bromine atoms in the molecule. Tribromomethane and dibromochloromethane were degraded by 99% and 80%, respectively, at the UV dose of 1200 mJ/cm2 with 6 mg/L H2O2, while 90% of the geosmin and 65% of the 2-methylisoborneol were degraded. Tribromoacetic acid and dibromoacetic acid were degraded by 99% and 80% respectively under the same conditions. Implications of this research are that the UV/H2O2 process, when implemented for odor control, can have the additional benefit of DBP elimination, especially in regions where source water bromide concentration leads to high concentrations of brominated DBPs. UV/H2O2 may also be more desirable than ozone in these regions due to possible formation of bromate from bromide ion during ozonation.
Acknowledgments The authors thank Kwater (Korea Water Resources Corporation) which provided a research fellowship for Dr. Chang Hyun Jo and the MILES (Macromolecular Interfaces with Life Science)eIGERT program at Virginia Tech (NSF agreement # DGE0333378) for experimental support. Neither funding agency contributed to the study design; collection, analysis or interpretation of data; decision to submit the paper for publication.
references
Beltran, F.J., Ovejero, G., Acedo, B., 1993. Oxidation of atrazine in water by ultraviolet-radiation combined with hydrogenperoxide. Water Research 27 (6), 1013e1021. Cassan, D., Mercier, B., Castex, F., Rambaud, A., 2006. Effects of medium-pressure UV lamps radiation on water quality in a chlorinated indoor swimming pool. Chemosphere 62 (9), 1507e1513. Chang, P.B.L., Young, T.M., 2000. Kinetics of methyl tert-butyl ether degradation and by-product formation during UV/hydrogen peroxide water treatment. Water Research 34 (8), 2233e2240. Chin, A., Berube, P.R., 2005. Removal of disinfection by-product precursors with ozone-UV advanced oxidation process. Water Research 39 (10), 2136e2144. Cotton, C.A., Collins, J.R., 2006. Dual purpose UV light: Using UV light for disinfection and for taste and odor oxidation. Proceedings of the 2006 Water Quality Technology Conference. American Water Works Association, Denver, CO, USA. Dietrich, A.M., 2006. Aesthetic issues for drinking water. Journal of Water and Health 4 (Suppl. 1), 11e16. Glaze, W.H., Schep, R., Chauncey, W., Ruth, C., Zarnoch, J.J., Aieta, E.M., Tate, C.H., McGuire, M.J., 1990. Evaluating oxidants for the removal of model taste and odor compound from a municipal water-supply. Journal of the American Water Works Association 82 (5), 79e84. Ho, L., Croue, J.P., Newcombe, G., 2004. The effect of water quality and NOM character on the ozonation of MIB and geosmin. Water Science and Technology 49 (9), 249e255. Hyun, K.S., Kim, Y.J., Choi, J.J., 2005. Profiles of THM and HAAs in the process of water treatment plants. Journal of Korean Society Water Science and Technology 13 (1), 55e66.
2515
Jo, C.-H., Dietrich, A.M., 2009. Removal and transformation of odorous aldehydes by UV/H2O2- efficiency and effect. Journal of Water Supply: Research and Technology-AQUA 58 (8), 580e587. Ju¨ttner, F., Watson, S.B., 2007. Biochemical and ecological control of geosmin and 2-methylisoborneol in source waters. Applied and Environmental Microbiology 73 (14), 4395e4406. Klassen, N.V., Marchington, D., McGowan, H.C.E., 1994. H2O2 determination by the I 3 method and by KMnO4 titration. Analytical Chemistry 66 (18), 2921e2925. Krasner, S.W., McGuire, M.J., Jacangelo, J.G., Patania, N.L., Reagan, K.M., Aieta, E.M., 1989. The occurrence of disinfection by-products in US drinking water. Journal of the American Water Works Association 81 (8), 41e53. Krasner, S.W., Weinberg, H.S., Richardson, S.D., Pastor, S.J., Chinn, R., Sclimenti, M.J., Onstad, G.D., Thruston, A.D., 2006. Occurrence of a new generation of disinfection byproducts. Environmental Science & Technology 40 (23), 7175e7185. Kurac, J., Hlatka, A., 1992. Photolysis of waterecarbon tetrachloride and waterechlorobenzene saturated solutions and their two phase system: quantum yields of chloride ions. Journal of Radioanalytical and Nuclear Chemistry 166 (3), 239e250. Kwok, W.M., Zhao, C.Y., Li, Y.L., Guan, X.G., Wang, D.G., Phillips, D.L., 2004. Water-catalyzed dehalogenation reactions of isobromoform and its reaction products. Journal of the American Chemical Society 126 (10), 3119e3131. Liao, C.-H., Gurol, M.D., 1995. Chemical oxidation by photolytic decomposition of hydrogen peroxide. Environmental Science and Technology 29 (12), 3007e3014. Lin, X.F., Guan, X.G., Kwok, W.M., Zhao, C.Y., Du, Y., Li, Y.L., Phillips, D.L., 2005. Water-catalyzed OeH insertion/HI elimination reactions of isodihalomethanes (CH2XeI, where X ¼ Cl, Br, I) with water and the dehalogenation of dihalomethanes in water-solvated environments. The Journal of Physical Chemistry A 109 (6), 981e998. Linden, C., Sharpless, C., Andrews, S., Atasi, K., Korategere, V., Stephan, M., Suffet, I.H., 2004. Innovative UV Technologies to Oxidize Organic and Organoleptic Chemicals. AwwaRF, Denver CO USA. Maruthamuthu, P., Padmaja, S., Huie, R.E., 1995. Rate constants for some reactions of free-radicals with haloacetates in aqueous-solution. International Journal of Chemical Kinetics 27 (6), 605e612. McGivern, W.S., Derecskei-Kovacs, A., North, S.W., Francisco, J.S., 2000. Computationally efficient methodology to calculate CeH and CeX (X ¼ F, Cl, and Br) bond dissociation energies in haloalkanes. The Journal of Physical Chemistry A 104 (2), 436e442. Mezyk, S.P., Helgeson, T., Cole, S.K., Cooper, W.J., Fox, R.V., Gardinali, P.R., Mincher, B.J., 2006. Free radical chemistry of disinfection-byproducts. 1. Kinetics of hydrated electron and hydroxyl radical reactions with halonitromethanes in water. The Journal of Physical Chemistry A 110 (6), 2176e2180. Mofidi, A., Min, J., Palencia, L., Coffey, B., Liang, S., Green, J., 2002. Advanced Oxidation Processes and UV Photolysis for Treatment of Drinking Water. California Energy Commission. http://www.energy.ca.gov/reports/2004-04-02_500-02-019/50002-019_A1.PDF (accessed January 2011). Nickelsen, M.G., Cooper, W.J., Lin, K., Kurucz, C.N., Waite, T.D., 1994. High energy electron beam generation of oxidants for the treatment of benzene and toluene in the presence of radical scavengers. Water Research 28 (5), 1227e1237. Nicole, I., De Laat, J., Dore, M., Duguet, J.P., Sutyh, E., 1991. Degradation of trihalomethanes in dilute aqueous solution by UV irradiation e determination of the quantum yield of photolysis at 253.7 nm. Environmental Technology 12, 21e31. Oppenlander, T., 2003. Photochemical Purification of Water and Air. Wiley-VHC.
2516
w a t e r r e s e a r c h 4 5 ( 2 0 1 1 ) 2 5 0 7 e2 5 1 6
Paradis, N., Hoffman, A., 2006. Mitigation of taste and odour compounds by UV/H2O2 advanced oxidation. In: Proceedings of the 2006 Water Quality Technology Conference. American Water Works Association, Denver, CO, USA November 6e9, 2006. Pereira, V.J., Weinberg, H.S., Linden, K.G., Singer, P.C., 2007. UV degradation kinetics and modeling of pharmaceutical compounds in laboratory grade and surface water via direct and indirect photolysis at 254 nm. Environmental Science and Technology 41 (5), 1682e1688. Peter, A., Von Gunten, U., 2007. Oxidation kinetics of selected taste and odor compounds during ozonation of drinking water. Environmental Science and Technology 41 (2), 626e631. Piriou, P., Devesa, R., De Lalande, M., Glucina, K., 2009. European reassessment of MIB and geosmin perception in drinking water. Journal of Water Supply: Research and TechnologyAQUA 58 (8), 532e538. Piriou, P., Soulet, C., Acero, J.L., Bruchet, A., Von Gunten, U., Suffet, I.H., 2007. Understanding medicinal taste and odour formation in drinking waters. Water Science and Technology 55 (5), 85e94. Qiao, R.P., Li, N., Qi, X.H., Wang, Q.S., Zhuang, Y.Y., 2005. Degradation of microcystin-RR by UV radiation in the presence of hydrogen peroxide. Toxicon 45 (6), 745e752. Rahn, R.O., 2004. Spatial distribution of upper-room germicidal UV radiation as measured with tubular actinometry as compared with spherical actinometry. Photochemistry and Photobiology 80 (2), 346e350. Rahn, R.O., Bolton, J., Stefan, M.I., 2006. The iodide/iodate actinometer in UV disinfection: determination of the fluence rate distribution in UV reactors. Photochemistry and Photobiology 82 (2), 611e615. Rashash, D.M.C., Dietrich, A.M., Hoehn, R.C., 1997. Flavor profile analysis of selected odorous compounds. Journal of the American Water Works Association 89 (4), 131e142. Richardson, S.D., Plewa, M.J., Wagner, E.D., Schoeny, R., DeMarini, D.M., 2007. Occurrence, genotoxicity, and carcinogenicity of regulated and emerging disinfection byproducts in drinking water: a review and roadmap for research. Mutation Research-Reviews in Mutation Research 636, 178e242. Rosenfeldt, E.J., Linden, K.G., 2007. The R-OH, R-UV concept to characterize and the model UV/H2O2 process in natural waters. Environmental Science and Technology 41 (7), 2548e2553. Rosenfeldt, E.J., Melcher, B., Linden, K.G., 2005. UV and UV/H2O2 treatment of methylisoborneol (MIB) and geosmin in water. Journal of Water Supply: Research and Technology-AQUA 54 (7), 423e434.
Royce, A., Stefan, M., 2005. Application of UV in drinking water treatment for simultaneous disinfection and removal of taste and odor compounds. In: Proceedings of the 2005 Water Quality Technology Conference. American Water Works Association, Quebec City, Canada November 6e10, 2005. Rudra, A., Thacker, N.P., Pande, S.P., 2005. Hydrogen peroxide and ultraviolet irradiations in water treatment. Environmental Monitoring and Assessment 109 (1e3), 189e197. Sarathy, S., Mohseni, M., 2010. Effects of UV/H2O2 advanced oxidation on chemical characteristics and chlorine reactivity of surface water natural organic matter. Water Research 44 (15), 4087e4096. Song, R., Westerhoff, P., Minear, R., Amy, G., 1997. Bromate minimization during ozonation. Journal of the American Water Works Association 89 (6), 69e78. Stefan, M.I., Bolton, J.R., 1998. Mechanism of the degradation of 1,4-dioxane in dilute aqueous solution using the UV hydrogen peroxide process. Environmental Science and Technology 32 (11), 1588e1595. United Stated Environmental Protection Agency (USEPA), 2011. National Primary Drinking Water Regulations. http://water.epa. gov/drink/contaminants/index.cfm#List (accessed 01.14.11). Wang, G.S., Hsieh, S.T., Hong, C.S., 2000. Destruction of humic acid in water by UV light e catalyzed oxidation with hydrogen peroxide. Water Research 34 (15), 3882e3887. Watson, S.B., Brownlee, B., Satchwill, T., Hargesheimer, E.E., 2000. Quantitative analysis of trace levels of geosmin and MIB in source and drinking water using headspace SPME. Water Research 34 (10), 2818e2828. Watson, S.B., Brownlee, B., Satchwill, T., McCauley, E., 1999. The use of solid phase microextraction (SPME) to monitor for major organoleptic compounds produced by chrysophytes in surface waters. Water Science and Technology 40 (6), 251e256. Westerhoff, P., Nalinakumari, B., Pei, P., 2006. Kinetics of MIB and geosmin oxidation during ozonation. Ozone-Science and Engineering 28 (5), 277e286. Zhang, X., Talley, J.W., Boggess, B., Ding, G.Y., Birdsell, D., 2008. Fast selective detection of polar brominated disinfection byproducts in drinking water using precursor ion scans. Environmental Science and Technology 42 (17), 6598e6603. Zhao, C.Y., Lin, X.F., Kwok, W.M., Guan, X.G., Du, Y., Wang, D.Q., Hung, K.F., Phillips, D.L., 2005. Water-catalyzed dehalogenation reactions of the isomer of CBr4 and its reaction products and a comparison to analogous reactions of the isomers of di- and trihalomethanes. Chemistrya European Journal 11 (4), 1093e1108.