Effects of dissolved organic matter size fractions on trihalomethanes formation in MBR effluents during chlorine disinfection

Bioresource Technology 136 (2013) 535–541

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Effects of dissolved organic matter size fractions on trihalomethanes formation in MBR effluents during chlorine disinfection Defang Ma, Baoyu Gao ⇑, Shenglei Sun, Yan Wang, Qinyan Yue, Qian Li Shandong Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Shandong University, Ji’nan 250100, China

h i g h l i g h t s  MBR effluents were fractionated by UF according to molecular weight (MW).  Fractions of MW >30 kDa were the main source of THMs precursor.  THMs formation was mostly attributed to slow chlorine decay.  THMs yield coefficients of the MW >30 kDa fractions were low.  THMFP and specific THMFP increased linearly with DOC and SUVA, respectively.

a r t i c l e

i n f o

Article history: Received 29 December 2012 Received in revised form 27 February 2013 Accepted 2 March 2013 Available online 14 March 2013 Keywords: MBR Dissolved organic matter (DOM) Molecular weight (MW) Chlorination Trihalomethanes (THMs)

a b s t r a c t In this study, effects of dissolved organic matter (DOM) size fractions on trihalomethanes (THMs) formation in MBR effluents during chlorination were investigated by fractionating DOM into >100, 30–100, 10–30, 5–10 and <5 kDa fractions using ultrafiltration (UF) membranes based on molecular weight (MW). Fractions of MW > 30 kDa constituted 87% of DOM and were the main THMs precursors, which exhibited higher specific ultraviolet absorbance (SUVA) and THMs formation potential (THMFP) and should be reduced to control THMs formation. For these fractions, THMs formation was mostly attributed to slow chlorine decay, and THMs yield coefficients were low because halogenated intermediates derived from the macromolecular DOM were difficult to decompose to produce THMs. Moreover, there was a strong linear correlation between dissolved organic carbon (DOC) concentration and THMFP (R2 = 0.981), as well as between the SUVA and specific THMFP (R2 = 0.993) in all fractions. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction Membrane bioreactor (MBR) has been widely applied in wastewater treatment intended for water recycling (Zanetti et al., 2010), which has been considered as an option to relieve water scarcity issues (Chen et al., 2012). MBR technology is the combination of a biological treatment system with membrane filtration processes, by which organic matter, nutrient, suspended solid and microbial are simultaneously removed (Li et al., 2009). High quality reuse water suitable for various types of recycle applications can be produced by MBR systems (Xia et al., 2008). However, bacteria may regrow due to the presence of biodegradable dissolved organic matter (DOM) and nutrients at low concentrations in the reclaimed water entering the distribution system (Zhang and DiGiano, 2002). In order to prevent the potential transmission of pathogenic microorganisms, adequate disinfection for MBR effluents prior to reuse is required (USEPA, 2004). ⇑ Corresponding author. Tel.: +86 531 88366771; fax: +86 531 88364513. E-mail addresses: [email protected], [email protected] (B. Gao). 0960-8524/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biortech.2013.03.002

Chlorine is a widely used disinfectant for water and wastewater disinfection due to its strong sterilization ability, easiness to use and low cost. Moreover, the presence of residual chlorine can effectively prevent the potential regrowth of bacteria in water distribution systems. However, chlorine is a strong oxidant and will react with DOM in the water to produce genotoxic and cytotoxic disinfection by-products (DBPs) including trihalomethanes (THMs) and haloaceticacides (HAAs) (Chellam and Krasner, 2001) during the chlorine disinfection process. It has been indicated that THMs and HAAs in drinking water could induce DNA damage in humanderived hepatoma line (Zhang et al., 2012). HAAs were also demonstrated to be mutagenic in Salmonella and Chinese hamster ovary cells (Plewa et al., 2010; Zhang et al., 2010). Consequently, it is required to balance the pathogen control and DBPs formation to ensure the ecological and health safety of water supplies (USEPA, 2004). Most of the developed countries have published guidelines to control DBPs in drinking water, among which the USEPA issued the Stage 2 Disinfectants (D)/DBP Rule to regulate the permissible levels of total trihalomethanes (TTHM) (80 lg/L) and total haloaceticacides (THAA) (60 lg/L) (USEPA, 2006). In order

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to control DBPs formation during the chlorine disinfection of drinking or natural water, extensive researches on the complicated reactions of chlorine with DOM and the subsequent formation of DBPs have also been conducted (Uyak et al., 2005; Zhan et al., 2010). Since DOM in natural water is a heterogeneous mixture of various organic compounds, a common method to facilitate understanding the formation of DBPs during chlorination processes is fractionating DOM into more homogenous components according to their physical and chemical properties (Kitis et al., 2002). XAD resins isolation and ultrafiltration (UF) membranes separation are the most widely used methods for the fractionation of DOM in drinking or natural water (Gang et al., 2003; Kitis et al., 2002; Leenheer, 1981). Compared with drinking or natural water, compositions of DOM in biologically treated wastewater are more complex and distinctly different (Barker and Stuckey, 1999). It is the heterogeneous mixture of recalcitrant natural organic matter (NOM) from drinking water, synthetic organic chemicals added during anthropogenic use and soluble microbial products (SMP), etc. (Shon et al., 2006; Wert et al., 2007). Reactions between chlorine and DOM during wastewater chlorination are definitely more complex than that during drinking water chlorine disinfection. As a result, fractionation of DOM is necessary to better understand the formation of DBPs in biologically treated wastewater during the chlorine disinfection process. XAD resins isolation is the most common method at present for the fractionation of DOM in biologically treated wastewater (Imai et al., 2002; Wei et al., 2009). XAD resins can fractionate DOM into defined components (e.g., hydrophobic and hydrophilic fractions) that meet specific research needs based on the chemical properties of DOM (Leenheer, 1981). In a previous study, DOM in secondary effluents was isolated into humicsubstances, hydrophobic bases, hydrophobic neutrals, hydrophilic acids, hydrophilic bases, and hydrophilic neutral by XAD resins to investigate the DOM fractions distribution. And the results indicated that humicsubstances and hydrophilic acids were found to be the dominated DOM fractions (Imai et al., 2002). Zhang et al. (2008) also isolated DOM in reclaimed water into six classes using XAD resins to study the influences of DOM fractions on DBPs formation, confirming that hydrophobic acids and hydrophilic acids exhibited the highest specific HAAs and THMs formation potential, respectively. Nevertheless, unknown chemical alteration of DOM may occur during the XAD resins isolation processes due to the extremes in pH and its irreversible interactions with the resins (Leenheer, 1981; Song et al., 2009). Moreover, only 90% of the DOM in the water tested was recovered by using XAD resins adsorption method (Kitis et al., 2002). In contrast, UF separation has been used to fractionate DOM based on the molecular weight (MW) of DOM, not on the chemical properties. Harsh chemical conditions such as extreme pH or contact with chemical solvents are absent during UF fractionation processes. It was reported that more than 99% of DOC (dissolved organic carbon) and almost 100% of UV254 (ultraviolet absorbance at 254 nm) were recovered from the UF fractions during the size isolation of Mississippi River waters. In otherwords, UF separation is more efficient than XAD resins adsorption method in the fractionation of DOM (Gang et al., 2003). It has been confirmed that UF separation could primely preserve the DOM properties of the source waters including the specific ultraviolet absorbance at 254 nm (SUVA) and the reactivity with chlorine (Kitis et al., 2002). The relationship between THMs formation and chlorine decay kinetics to DOM molecular size during the chlorine disinfection of natural water has been widely investigated in previous researches (Gang et al., 2003; Zhao et al., 2006). However, studies on the role of DOM molecular size in DBPs formation during biologically treated wastewater chlorination are limited. In particular, information about the relationship between THMs formation and DOM size fractions during the

chlorine disinfection of municipal wastewater treated by MBR is limited. The primary objective of this research was, therefore, to investigate the role of DOM molecular size in THMs formation during the chlorine disinfection of MBR effluents. The municipal wastewater treated by a laboratory-scale submerged MBR was fractionated through UF separation method to study the DOM MW distribution and SUVA (specific ultraviolet absorbance at 254 nm) of each fraction. These fractions were further disinfected with chlorine, and the trihalomethanes formation potential (THMFP) of each fraction was evaluated to reveal the major THMs precursors among the different DOM fractions. TTHM data of those fractions having the highest THMFP were fitted to the DBPs formation model to investigate their THMs formation kinetics. 2. Methods 2.1. Experimental setup Municipal wastewater was treated by using a laboratory-scale submerged MBR which consisted of a rectangular tank having an operating volume of 18 L and a hollow-fiber vacuum-type membrane module submerged in the tank. The membrane module was made of poly-vinylidene fluoride (PVDF) membranes with a nominal pore size of 0.02 lm and a total effective membrane surface area of 0.175 m2. Aeration was done through oxygen nozzles beneath the membrane module to provide oxygen for the microorganism and to maintain a constant air flow shear force through the membrane module to reduce fouling and cake formation. A continuous mixer was used to ensure the homogeneous condition of sludge in the reactor. The municipal wastewater pumped from the detritus chamber of a local sewage treatment plant (Jinan, China) was pre-treated with a mesh sieve (0.2 mm) to reduce the braid formation of the membrane module. Then it was delivered to the MBR tank by a peristaltic pump, which was controlled by a water level sensor to maintain a constant water level in the bioreactor. The membrane-filtered effluent was extracted intermittently (8 min on and 2 min off) with a peristaltic pump controlled by the timer. The trans membrane pressure (TMP) was continuously monitored by a pressure gauge. A more detailed description of the MBR system including the flux diagram has been published earlier (Ma et al., 2013). 2.2. MBR operating conditions The MBR was seeded with activated sludge obtained from the aeration basin of a local sewage treatment plant (Jinan, China). A start-up period of two months was provided for the seeded sludge to acclimate the new environment, during which no sludge was wasted. Thereafter, the MBR was operated for the next two months to reach stabilization. During this period, the sludge retention time (SRT) was set at 180 days by regularly discharging a certain amount of sludge. When the mixed liquor suspended solids (MLSS) concentration in the reactor and five-day biological oxygen demand (BOD5) of the effluent kept constant for 20 days, UF fractionation and chlorination experiments for the MBR effluent were carried out. The membrane-filtered effluents used for UF fractionation and chlorination experiments were taken from the MBR system at 20, 23, 26, 29 and 35 days after steady state (i.e., 135, 138, 141, 144 and 150 days after the addition of seed sludge). The hydraulic retention time (HRT) of 6 h and dissolved oxygen (DO) concentration of 3.0 ± 0.5 mg/L were maintained during the entire start-up and experimental period. The MBR was operated under ambient temperature (27 ± 2 °C), and the pH was controlled within a range of 6.5–7.5.

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2.3. Fractionation of DOM samples A Minim™ II Tangential Flow Filtration (TFF) system (PALL, USA) was used to fractionate the MBR effluent into different DOM size fractions. The TFF system consisted of Minim™ capsules (PALL Corporation) with Omega™ membrane with molecular weight cut-off (MWCO) of 5, 10, 30 and 100 kDa. 5 L the water sample was filtrated through UF membranes in series with large pore-size membranes followed by smaller pore-size ones in sequence (Fig. 1) to generate five DOM size fractions. Each fraction was diluted to 5 L with ultrapure water. The preparation and MW range for each fraction are listed in Table 1. DOM recoveries (as % DOC and % UV254) were calculated after each fractionation experiment to evaluate any loss or contamination during the UF separation processes (Gang et al., 2003). DOC recovery was calculated as the ratio of the sum of fractions’ carbon mass to the carbon mass of the original unfractionated MBR effluents (Kitis et al., 2002). The carbon mass of each sample was calculated from its DOC concentration and corresponding volume (Zhao et al., 2006). UV254 recovery was also quantified by this calculation. 2.4. Chlorine disinfection process Chlorine disinfection experiments were conducted at 25 ± 1 °C. Chlorine dosing solutions with an active chlorine concentration (as Cl2) of 10 mg/L were prepared with sodium hypochlorite which was titrated according to the APHA standard method (APHA, 2005). The chlorine dosing solutions were immediately used for disinfection after preparation. In the THMFP analysis, the chlorine dosage is selected to yield at least 1–2 mg/L free chlorine residual in the water. A chlorine demand preliminary study using a series of four chlorine dosages of 10, 15, 20 and 40 mg/L was conducted to determine the 110 h chlorine demand of the MBR effluent. Since the average chlorine demand of the MBR effluent was 17.93 mg/L, the chlorine dosage in this study was selected as 20 mg/L. After 110 h chlorination, the average free chlorine residual in the MBR effluent was 2.07 mg/L. In the chlorination experiments, 1 mL chlorine dosing solution as mentioned above was added to a dark brown glass bottle with a PTFE-lined screw cap containing 500 mL water sample. After dosing, the bottle was covered, mixed, and kept in darkness. Then, THMs concentration and chlorine residual were measured at different times for each sample. The samples used for THMs analysis were immediately transferred to the bottles containing sodium thiosulphate solution (0.1 mol/L) to neutralize the chlorine residue. THMFP of each fraction was determined after a 110 h chlorine disinfection process. In order to investigate the THMs yield on a per carbon basis, the TTHM data were normalized relative to the DOC concentrations to obtain the specific THMFP. 2.5. THMs formation kinetics TTHM data of the fractions exhibiting the highest THMFP were fitted to the DBPs formation model derived from natural water

100 kDa > MW > 30 kDa

Retentate MBR effluent

100 kDa

Filtrate

where C0 (mg/L) is the initial chlorine dosage to provide a CT value (the product of total chlorine residual and contact time measured at the same point) of not less than 450 (mg min)/L at any time (USEPA, 2004); a is the TTHM yield coefficient (lg/L-TTHM/mg-Cl2 consumed), defined as the ratio of TTHM concentration (lg/L) to the chlorine consumed (mg/L); x is the fraction of chlorine added (C0) reduced by the rapid reacting substances; k1, k2 (h1) are the THMs formation kinetic constants for rapid and slow chlorine decay reactions, respectively, k1 > k2; t (h) is the reaction time. The mathematical treatment of curves started at 1 min from the dosing. Data were processed by the non-linear regression software (SigmaPlot 12.0, SPSS). 2.6. Analytical methods BOD5, COD, MLSS, NH4+ and TSS were measured according to the APHA standard methods (APHA, 2005). Turbidity was examined by a portable microprocessor turbidity meter (Hanna, Italy). DO and pH were measured by a portable DO meter (Precision & Scientific Instrument, China) and a pH meter (Luo Qi Te, China), respectively. DOC was quantified by a TOC-VCPH Total Organic Carbon Analyzer (SHIMADZU, Japan). UV254 was measured by using TU-1810 UV/VIS spectrophotometer (PGENERAL, China). SUVA was calculated as the ratio of UV254 to DOC. Free and total chlorine were measured by a Free & Total chlorine measuring meter (HANNA, Italy) according to the DPD powder pillow photometric method (APHA, 2005). THMs including bromodichloromethane (CHBrCl2), dibromochloromethane (CHBr2Cl), trichloromethane (CHCl3) and tribromomethane (CHBr3) were determined according to the headspace method by a gas chromatograph, GC-ECD (SHIMADZU, Japan). The GC-ECD operating conditions for THMs were as follows: detector temperature: 200 °C, injector temperature: 120 °C, injection volume: 1.00 lL.

3. Results and discussion 3.1. Performance of MBR MLSS concentrations in the MBR during the experiment were monitored (Fig. S1, Supplementary data). After seeding sludge, the biomass was found to decrease in the first 20 days. Hereafter the biomass increased and reached steady state after 115 days with the MLSS concentration of 6.51 ± 0.12 g/L.

DOM4

30 kDa > MW > 10 kDa 10 kDa > MW > 5 kDa

Retentate

30 kDa

TTHM ¼ aC 0 ½1  x expðk1 tÞ  ð1  xÞ expðk2 tÞ

DOM3

DOM2

DOM1 MW > 100 kDa

chlorine disinfection modeling to investigate the THMs formation kinetics (Gang et al., 2003). The model assumes that THMs formation is a function of chlorine consumption. Chlorine decay is presumed to proceed through two independent reactions: an initial rapid decay which is attributed to rapid reacting substances in the water, and a slow and continuous decay process due to slow reacting substances. The model was described as follows:

Filtrate

Retentate

10 kDa

Filtrate

Retentate

5 kDa

Fig. 1. Serial processing scheme in the Tangential Flow Filtration system.

Filtrate DOM5 MW < 5 kDa

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Table 1 Concentration of the individual THMs in DOM fractions after 110 h chlorination.a

CHCl3 (lg/L) CHBrCl2 (lg/L) CHBr2Cl (lg/L) CHBr3 (lg/L)

DOM1

DOM2

DOM3

DOM4

DOM5

145.59 ± 5.31 107.38 ± 5.01 94.83 ± 5.15 12.35 ± 0.20

109.40 ± 6.34 81.44 ± 1.47 73.86 ± 1.67 13.16 ± 0.52

15.33 ± 0.95 8.55 ± 0.41 5.12 ± 0.50 ND

1.70 ± 0.21 0.43 ± 0.08 0.16 ± 0.09 ND

1.05 ± 0.25 ND ND ND

a Sample mean ± standard deviation, number of measurements: 5 DOM1: >100 kDa fraction; DOM2: 30–100 kDa fraction; DOM3: 10–30 kDa fraction; DOM4: 5–10 kDa fraction; DOM5:<5 kDa fraction; ND: not detected.

3.2. DOM fractionation and MW distribution in MBR effluent DOM in the MBR effluent was isolated into the MW > 100 kDa, 30–100 kDa, 10–30 kDa, 5–10 kDa and <5 kDa fractions by using UF membranes. The average concentration of DOC in the five MBR effluents used for UF fractionation was 4.936 ± 0.127 mg/L. DOC and UV254 recovered in the UF fractions were 102.0% and 100.7%, respectively. The low-level positive errors in the DOM recoveries were attributed to the analytical errors (e.g., the very low-level DOC and UV254 measurements required for the fractions of MW less than 10 kDa). The results indicated that no mass loss or contamination occurred during the overall UF filtration process. Consequently, the fractionated waters could completely represent the composition of DOM in the MBR effluent. Fig. 2 shows the MW distribution of DOM in the MBR effluent. It was found that DOC proportions of DOM fractions reduced as MW decreased. In the MBR effluent, MW of more than 87% DOM was larger than 30 kDa. In particular, the >100 kDa fraction was dominant, contributing about 52% of the total DOC. The 30–100 kDa fraction was the second most abundant fraction, constituting about 35% of the total DOC. DOC percent of the 10–30 kDa fraction was about 10%. Only 1% of the DOC was in the <5 kDa fraction, which was low enough to be ignored. MW distribution of DOM varied substantially depending on the kind of waters and the type and operation conditions of the treatment technology. It was reported that for approximately half of DOM in natural water, MW was less than <1 kDa (Kitis et al., 2002). Liang et al. (2007) studied the MW distribution of synthetic wastewater treated by MBR. The results indicated that the <3 kDa fraction was dominant, whereas the >30 kDa fraction constituted 23–32% of DOM, which was due to the relative lower SRT (less than 40 days). In this study, DOM in the municipal wastewater treated by MBR was the mixture of initial soluble inert organic matters in the raw municipal wastewater and soluble microbial products (SMP) which were generated and released through microbial metabolic activities in the bioreactor. It has been indicated that SMP

60

50

DOC percentage (%)

In this study, the influent municipal wastewater was highly contaminated by organic matter with DOC concentration of 27.105 ± 3.300 mg/L. The water quality fluctuated widely (Ma et al., 2013). Despite this, high quality water was produced by the MBR due to the efficient removal of suspended solid and biodegradable organic matter (Ma et al., 2013). Detailed information about the performance of MBR with respect to the removal of BOD5, COD, DOC, NH4+, TSS, Turbidity and UV254 and its influence on TMHs formation during chlorination of MBR effluents can be found in other study (Ma et al., 2013). The TMP changes exerted on the membrane module over time were monitored. The TMP increased step by step from 0 to 10 kPa at the initial 20 days, and skyrocketed up to 50 kPa in the next 9 or 10 days, at which time obvious module sludging was observed. The sludge accumulation inside the membrane bundle could be easily removed by rinsing with fresh water. After cleaning, the membrane permeability could be completely recovered, which indicated that the membrane fouling was reversible.

40

30

20

10

0 >100

30– 100

10– 30 5 – 10 Molecular weight range (kDa)

<5

Fig. 2. Molecular weight distribution of DOM in the MBR effluent. (Error bars represent the standard deviations from 5 different MBR effluents).

constituted the majority of DOM in biological treatment effluents (Barker and Stuckey, 1999). SMP are generally classified into biomass associated products (BAP) and utilization associated products (UAP). UAP are associated with substrate metabolism and biomass growth (Barker and Stuckey, 1999), which are mostly composed of small molecules (86% and 76% <1 kDa for phenol and glucose, respectively) (Boero et al., 1996). BAP are associated with biomass decay and produced at a rate proportional to the concentration of biomass (Barker and Stuckey, 1999), which are mostly composed of large molecules (47% and 52% >100 kDa for phenol and glucose substrates, respectively) (Boero et al., 1996). It has been indicated that the high molecular fraction in SMP increases with sludge retention time (SRT) (Pribyl et al., 1997). Since MBR is generally operated at high SRT and biomass concentrations, the generation of SMP is mainly attributed to microbial decay through endogenous respiration (Hocaoglu and Orhon, 2010). In this study, the SRT of the MBR was up to 180 days, and the biomass concentration (as MLSS) was up to 6.51 ± 0.12 g/L, which led to a low sludge loading of 0.08 kg-BOD/(kg-MLSSd). Under this condition, microorganisms in the bioreactor were in starvation and excreted more macromolecular SMP (i.e., BAP) through metabolism of intracellular components. Consequently, DOM in the MBR effluent was mostly composed of high-molecular-weight organic compounds. 3.3. SUVA of DOM size fractions SUVA is known to be a surrogate parameter for aromaticity or hydrophobicity of waters (Drews, 2010). It has been routinely utilized to predict THMs formation during water chlorine disinfection due to the strong linear correlation between SUVA and specific THMFP (Zhao et al., 2006). High SUVA generally indicates high degree of specific THMFP in water or wastewater due to its high hydrophobicity or aromaticity.

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4

(a) 400 2

R = 0.981 300

THMFP ( g/L)

SUVA (L/mg m)

3

2

200

100

1 0

0

0.0

30– 100

10– 30

5 – 10

.5

1.0

<5

Molecular weight range (kDa)

2.0

2.5

Specific THMFP ( g/mg-C)

160

R2 = 0.993

140 120 100 80 60 40 20 0 1.6

1.8

2.0

2.2

2.4

2.6

2.8

3.0

3.4

Fig. 4. Correlations between THMs formation and DOM properties of each size fraction: (a) correlation between THMFP and DOC; (b) correlation between specific THMFP and SUVA. (Error bars represent the standard deviations from 5 different MBR effluents).

180

400 THMFP Specific THMFP

THMFP ( g/L)

The total yields of TTHM calculated by summing contributions of individual chlorinated fractions were 670.352 lg/L, which was statistically similar with that of the chlorinated unfractionated MBR effluent (665.070 lg/L). This result indicated that UF fractionation maintained the integrity and reactivity of DOM in respect of THMs formation. This finding also implied that the contribution of each fraction to total TTHM formation was linearly cumulative, and there was no synergistic effect among the DOM fractions with regard to react with chlorine to produce THMs. For each fraction, THMFP was directly related with the DOC content. As shown in Fig. 4(a), THMFP level increased linearly with increase of DOC content and the linear correlation coefficient (R2) was 0.981. It was consistent with the previous study which indicated that high DOC content led to high THMFP (Kitis et al., 2002). As demonstrated in Fig. 4(b), there was also a strong linear correlation between the SUVA and specific THMFP (R2 = 0.993) in all DOM fractions. This suggested that formation of THMs for the

3.2

SUVA (L/mg m)

160 140

300

3.4. THMs formation of DOM size fractions

3.0

(b) 180

Fig. 3. SUVA of each size fraction in the MBR effluent. (Error bars represent the standard deviations from 5 different MBR effluents).

As shown in Fig. 3, SUVA of the high-molecular-weight (MW larger than 30 kDa) fractions was higher than that of the lowmolecular-weight (MW less than 30 kDa) fractions. In particular, the 30–100 kDa fraction exhibited the highest SUVA (3.233 L/ mg m), which was possibly caused by hydrophobic humic acid-like or protein-like SMP (Wang and Zhang, 2010) generated through endogenous respiration due to the very high SRT (180 days) in the MBR (Hocaoglu and Orhon, 2010). The result indicated that the 30–100 kDa fraction had the highest hydrophobicity or aromaticity. As a result, it was predicted that the 30–100 k fraction produced the highest specific THMFP yields during chlorine disinfection processes, which would be confirmed in the following section. The >100 kDa fraction exhibited the second highest SUVA (3.143 L/mg m) in spite of its highest DOC content. Hydrophobicity of the >100 kDa fraction was slightly lower than that of the 30– 100 kDa fraction, which could be rationalized by that the >100 kDa fraction comprised more hydrophilic macromolecule polysaccharide. The 5–10 kDa and <5 kDa fractions had similar SUVA, and the value was approximately 1.8 L/mg m. It has been proved that the organic matter was mostly composed of nonhumic substances and was of low hydrophobicity when SUVA was less than 2 L/mg m (Chu et al., 2011).

1.5

DOC (mg/L)

120 100 200

80 60 100

40

Specific THMFP ( g/mg-C)

>100

20 0

0

>100

30– 100

10 – 30

5– 10

<5

Molecular weight range (kDa) Fig. 5. THMs formation of each size fraction in the MBR effluent. (Error bars represent the standard deviations from 5 different MBR effluents).

MBR effluent was fundamentally depended on the aromaticity of DOM. TTHM yields of these five DOM fractions after 110 h chlorine disinfection process were presented in Fig. 5. It could be seen that THMFP gradually reduced as MW decreased during chlorine

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disinfection processes for all fractions, confirming that the large size DOM was the primary source of THMs precursors in the MBR effluent. The results showed that more than 95% of the THMs formation was attributed to the chlorination of DOM with MW larger than 30 kDa. In particular, the >100 kDa fraction had the largest THMFP, contributing up to 54% of total TTHM formed in the unfractionated MBR effluent during the chlorine disinfection process, which was due to its highest DOC content. As predicted, the specific THMFP of the >100 kDa fraction was lower than that of the 30–100 kDa fraction due to the relative lower hydrophobicity of the >100 kDa fraction. Sum of the TTHM yields in chlorinated low-molecular-weight (<30 kDa) fractions constituted less than 5% of the total TTHM formation during chlorination of the unfractionated MBR effluent. In addition, these fractions produced much lower specific THMFP, which was consistent with their low SUVA values. The results indicated that the large size DOM originating from metabolism of intracellular components (e.g., hydrophobic humic acid-like and protein-like SMP) in the MBR effluent was more reactive to produce THMs during chlorine disinfection processes. Concentration of the individual THMs in DOM fractions after 110 h chlorination was listed in Table 1. CHCl3, CHBrCl2, CHBr2Cl and CHBr3 were all detected in the >100 kDa and 30–100 kDa fractions. For both fractions, the yield of THMs species in order based on mass amount was CHCl3 > CHBrCl2 > CHBr2Cl > CHBr3. For the 10–30 kDa and 5–10 kDa fractions, CHCl3 was also found to be the dominant species, whereas CHBr3 was not detected. CHCl3 was the only species detected in the <5 kDa fraction.

3.5. THMs formation kinetics The DBPs formation model was calibrated against TTHM data sets to obtain a single invariant set of parameters for the >100 kDa fraction and 30–100 kDa fraction which were the major THMs precursors. The fitted curves were shown in Fig. 6. TTHM formation kinetics parameters including rapid and slow reaction rate constants (k1 and k2), TTHM yield coefficient (a) and rapid reacting substances proportion (x) were listed in Table 2. For both fractions, the TTHM data fit the DBPs formation model well with correlation coefficients of 0.99 (Table 2). This result indicated that THMs formation was fundamentally depended on chlorine consumption during the chlorine disinfection of the fractionated MBR effluent (Gang et al., 2003). TTHM yield coefficients of the >100 kDa

Table 2 TTHM formation kinetics parameters. Fraction

DOM1 DOM2

Molecular weight range

k1 (h1)

k2(h1)

>100 kDa 30–100 kDa

1.816 2.050

0.038 0.028

x

a (lg/L-TTHM/

R2

mg-Cl2 consumed) 0.24 0.31

16.9 17.7

0.994 0.996

fraction and 30–100 kDa fraction were were lower than that of the fractionated natural waters, which were mostly composed of low-molecular-weight organics (Gang et al., 2003). This was due to the fact that halogenated intermediates formed from DOM in the MBR effluent were difficult to decompose to produce THMs. As can be seen from the x values listed in Table 2, about 30% of chlorine consumption was attributed to the rapid reaction for both fractions, which indicated that most of THMs yields were attributed to the chlorination of slow reacting components. Compared with the >100 kDa fraction, the 30–100 kDa fraction had a higher x value, suggesting that more chlorine followed the rapid decay in the 30–100 kDa fraction. The ratios of rapid reaction rate constant to slow reaction rate constant for the >100 kDa and 30–100 kDa fractions were approximately 47 and 73, respectively. These values were lower than that of a previous report, which indicated that the constants of rapid decay rate were about 90–150 times larger than those of the slower decay rate for fractionated surface water samples (Gang et al., 2003). The reaction rates were depended on the characteristics of the water samples and available chlorine residual concentrations in water samples at the end of the chlorine disinfection process. Compared with natural water, DOM in the MBR effluent was mostly composed of high-molecularweight SMP with complex chemical structure, resulting a slower rapid reaction rate. Moreover, the slow reactions were accelerated due to the high effective chlorine residual levels of both fractions (4.60 mg/L for the >100 kDa fraction and 9.16 mg/L for the 30– 100 kDa fraction) after 110 h chlorination process. Rapid reaction constant of the 30–100 kDa fraction was higher than that of the >100 kDa fraction. This indicated that THMs formation and chlorine decay in 30–100 kDa fraction was faster, confirming that the 30–100 kDa fraction comprised more organic matters that were apt to react with chlorine to generate THMs.

4. Conclusions

TTHM concentration ( g/L)

400

In this study, fractions of MW > 30 kDa constituted 87% of DOM and were the main THMs precursors, exhibiting higher SUVA and THMFP. Macromolecular DOM should be reduced to control THMs formation. For these fractions, THMs formation was mostly attributed to slow chlorine decay, and the THMs yield coefficients were low because halogenated intermediates derived from the macromolecular DOM were difficult to decompose to produce THMs. Moreover, there was a strong linear correlation between DOC concentration and THMFP (R2 = 0.981), as well as between SUVA and specific THMFP (R2 = 0.993) in all fractions.

300

200

>100 kDa fraction 30-100 kDa fraction Model

100

Acknowledgements 0 0

20

40

60

80

100

120

Time (h) Fig. 6. Modeling of THMs formation of the >100 kDa and 30–100 kDa fractions. (Error bars represent the standard deviation based on triplicate analyses of one MBR effluent sampled at 150 days after addition of seed sludge).

The research was financially supported by the Key Scientific Technology Program for Environmental Protection of Shandong, China (16) and the National Major Special Technological Programmes Concerning Water Pollution Control and Management in the Twelfth Five-year Plan Period (2012ZX07203004).

D. Ma et al. / Bioresource Technology 136 (2013) 535–541

Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.biortech.2013. 03.002.

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