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Naturally occurring processes such as biological reaction might modify the properties of dissolved organic matter (DOM) for binding with heavy metals. Complexation of heavy metals with DOM determines their environmental and ecological impacts. In this study, biodegradation experiments were carried out separately for river water and river water spiked with DOM in wastewater treatment plant (WWTP) effluent to evaluate the biological alteration of zinc complexation characteristics of DOM under river water environment. Zinc complexation parameters, conditional stability constant and binding site concentrations were determined using anodic stripping voltammetry with Scatchard linearization. Total ambient zinc binding site concentration of river water DOM was reduced from 410 nM to 74 nM (82% decline) during two weeks of incubation. Compared to the river water DOM, 1-month incubation of DOM in WWTP effluent under the river water environment, showed only 22% decline in total ambient Zn binding sites. On the other hand, conditional stability constants, for Zn binding sites of DOM from WWTP effluent, did not vary during 1 month of incubation. The result suggests that metal (Zn) binding sites of DOM from WWTP effluents are biologically persistent in the urban river water environment.
Journal of Water and Environment Technology, Vol. 8, No.4, 2010 Address correspondence to Tushara CHAMINDA G. G, Environmental Science Center, The University of Tokyo, Email: chaminda@env.t.u-tokyo.ac.jp Received May 10, 2010; Accepted September 13, 2010. - 403 - Biological Alteration of Zinc Complexation Characteris- tics of Dissolved Organic Matter in Domestic Wastewa- ter Treatment Plant Effluent under River Water Environment Tushara CHAMINDA G. G*, Fumiyuki NAKAJIMA*, Ikuro KASUGA** *Environmental Science Center, The University of Tokyo, 7-3-1 Hongo, Bunkyo, Tokyo, 113- 0033, Japan **Department of Urban Engineering, School of Engineering, The University of Tokyo, 7-3-1, Hongo, Bunkyo, Tokyo, 113-8656, Japan ABSTRACT Naturally occurring processes such as biological reaction might modify the properties of dissolved organic matter (DOM) for binding with heavy metals. Complexation of heavy metals with DOM determines their environmental and ecological impacts. In this study, biodegradation experiments were carried out separately for river water and river water spiked with DOM in wastewater treatment plant (WWTP) effluent to evaluate the biological alteration of zinc complexation characteristics of DOM under river water environment. Zinc complexation parameters, conditional stability constant and binding site concentrations were determined using anodic stripping voltammetry with Scatchard linearization. Total ambient zinc binding site concentration of river water DOM was reduced from 410 nM to 74 nM (82% decline) during two weeks of incubation. Compared to the river water DOM, 1-month incubation of DOM in WWTP effluent under the river water environment, showed only 22% decline in total ambient Zn binding sites. On the other hand, conditional stability constants, for Zn binding sites of DOM from WWTP effluent, did not vary during 1 month of incubation. The result suggests that metal (Zn) binding sites of DOM from WWTP effluents are biologically persistent in the urban river water environment. Keywords: biodegradation, dissolved organic matter, river water, wastewater treatment plant effluent, zinc complexation. INTRODUCTION AND BACKGROUND Dissolved organic matter (DOM) is important in the transport of metals in aquatic systems. It also provides protection to the aquatic organisms from heavy metal stress by forming strong complexes, which results in the decline of the free metal ion concentration. Many water bodies in the urban environment, such as rivers and bays, receive effluent from wastewater treatment plant (WWTP) which may contain considerable concentrations of toxic metals as well as DOM that control the metal speciation. The chemical structure of DOM, which originated from biological wastewater treatment plant, is much complicated and includes organic compounds in different groups (Ma et al., 2001). Recent studies have revealed that DOM in treated effluent tends to bind with many metals (eg; Zn and Cu) to a greater degree than natural organic matter (NOM) does (Sarathy and Allen., 2005; Cheng and Allen., 2006; Chaminda et al., 2008). No matter how strong the complexations of metals are with DOM, naturally occurring processes such as biological degradation and photochemical reaction might modify the Journal of Water and Environment Technology, Vol. 8, No.4, 2010 - 404 - properties of DOM in binding with heavy metals. Although the photodegradation effects of DOM on heavy metal speciation have been investigated (Shiller et al., 2006; Brooks et al., 2007) to some extent, biodegradation effects of DOM have not been well understood. The microbial degradation of organic ligands in natural water bodies may release free metal ions and subsequently increase the metal toxicity. In this study, biodegradation experiments were carried out for river water and river water spiked with DOM from WWTP effluent to evaluate the biological alteration of zinc complexation characteristics of DOM under river water environment. Zinc complexation parameters (conditional stability constant and binding site concentration), which characterize the nature of DOM responsible for complexing with Zn, were determined using the square wave anodic stripping voltammetry (SWASV). MATERIALS AND METHODS Sampling The effluent used in this study was collected from a WWTP, where domestic wastewater was treated using conventional activated sludge process and disinfected by chlorination, in the late morning of a dry day in April, 2009 (pH = 7.0; dissolved organic carbon (DOC) = 8.1 mg/L; average ambient temp. = 19°C). River water samples for preliminary study and biodegradation study were collected from the Edogawa River (+35° 45' 56.3", +139° 52' 46.1") in January, 2009 (pH = 7.2; DOC = 2.1 mg/L; average ambient temp. = 7°C) and April, 2009 (pH = 7.1; DOC = 0.9 mg/L; average ambient temp. = 19°C), respectively. River water samples were immediately filtered by 10 µm PTFE filter (Millipore) to remove large particles and organisms and were stored in a refrigerator for biodegradation experiments. A portion of river water samples was also filtered through 0.5 µm PTFE filter (Millipore) to separate the suspended particles and was kept in a refrigerator at 4 o C in the laboratory for dissolved and labile Zn analysis. Collected WWTP effluent was filtered through 0.5 µm PTFE filter (Millipore) and stored in a freezer for further analysis and biodegradation experiments. Biodegradation experiment Preliminary experiments were conducted using the river water samples collected in January, 2009 to find whether the metal (Zn) binding sites in river water DOM are biodegradable. The river water sample (filtered by 10 µm PTFE filter) was transferred to 1 L bottles which were covered to avoid exposure to light. Then, two separate bottles (bottle 1 and bottle 2) were subjected to incubation at controlled temperature under 20°C for 6 weeks. The bottles were stopped with air permeable silicon caps to allow the passage of air and to prevent the system from contamination and from becoming anaerobic. Headspaces were also left in the bottles in such a way that air exchange could take place during the incubation. Throughout the incubation, the bottles were shaken manually at 1 - 2 day intervals. Water samples were withdrawn after two (from bottle 1) and six (from bottle 2) weeks and analyzed for Zn complexation parameters (conditional stability constant, K' and binding site concentration, [L]) using voltammetry. Water samples were also withdrawn in 3-7 day intervals for DOC, UV absorbance at 254 nm (UV 254 ), SUVA (specific UV absorbance) and dissolved and labile Zn analysis. Journal of Water and Environment Technology, Vol. 8, No.4, 2010 - 405 - Although the pH was not maintained at a constant value, the results showed that the variation of pH (7.0 - 7.2) was not significant during the incubation. Cations in the filtered WWTP effluent samples (< 0.5 µm) were removed by cation exchange resin AGMP-50 (H + form, 100 - 200 mesh, Bio Rad) in a batch mode and then DOC was concentrated by freeze drying (FDU 540, EYELA). Subsequently, the concentrated DOM of WWTP effluent was spiked into river water sample which was collected in April and pre-incubated for two weeks. Two weeks were chosen, because the preliminary results, which will be discussed in the succeeding section of the paper, suggested that Zn binding sites in river water DOM seemed to be persistent after two- week incubation. The mixture of DOM of WWTP effluent and the pre-incubated river water (DOC in the mixture = 4.9 mg/L) was incubated, in three separate bottles (bottle 1, bottle 2 and bottle 3), for one month under the same condition as explained in the preliminary experiment. Samples were withdrawn after 4 days (bottle 1), 7 days (bottle 2) and 1 month (bottle 3) to determine the Zn complexation parameters. Determination of dissolved and labile Zn, DOC, UV absorbance and SUVA Dissolved organic carbon concentration was measured as non-purgeable organic carbon with a TOC analyzer (TOC-V CSH, Shimadzu, Japan). Absorbance at 254 nm (UV 254 ) was obtained by a Hitachi U-2000 UV spectrophotometer using a 1.0 cm quartz cuvette. Specific UV absorbance is defined as the UV 254 divided by the DOC concentration. Dissolved Zn and labile Zn concentrations were analyzed by an ICP-MS (HP 4500, Yokogawa, Japan), in accordance with the methods described by Chaminda et al. (2008). The labile Zn was operationally defined by fractionation using a chelating disk cartridge (Empore) (Chaminda et al., 2008). Zinc complexation analysis using voltammetry The filtered (< 0.5 µm) samples of the WWTP effluent and river water were subjected to SWASV to determine the conditional stability constant (K') and binding site concentration ([L]) of Zn. All the samples were either concentrated by freeze drying or diluted by Milli-Q water to maintain an equal DOC concentration (3.5 mg/L) and then subjected to zinc titration with Nano-Band TM Explorer portable system (TraceDetect), equipped with a thin film mercury electrode. Zinc titration against DOM in water samples were conducted by the addition of known increments of Zn followed by the measurement of electrochemical labile Zn after a certain equilibrium period. The Zn titration data were then linearly transformed using the Scatchard equation, as explained in Chaminda et al. (2010) and in the Appendix of this paper. RESULTS AND DISCUSSION Biodegradability of zinc binding sites in river water DOM Fig. 1 represents the variation of DOC, UV 254 and SUVA during the river water incubation for 6 weeks in the preliminary study. During the first 4 days of incubation, DOC concentration declined by 54%. No further significant DOC reduction was observed up to 6 weeks of incubation as shown in Fig. 1. Similar to the trend in DOC, there was a momentous change in UV 254 and SUVA, which is a good indicator of aromaticity, during the first 4 days and no significant changes were observed afterwards. The original river water and the withdrawn samples after 2 weeks and 6 weeks of Journal of Water and Environment Technology, Vol. 8, No.4, 2010 - 406 - 0.000 0.005 0.010 0.015 0.020 0.025 0.030 0.035 0.040 0.0 1.0 2.0 3.0 4.0 5.0 0 1020304050 UV 254, cm -1 DOC (m g L⁻¹), SUVA (Lmg -1 m -1 ) Incubation period, days Note: Error bars represent the standard deviation of triplicate measurements DOC, mg L -1 SUVA₂₅₄, L mg -1 m -1 UV₂₅₄, cm -1 Fig. 1 - Variation of DOC, UV and SUVA during the incubation of river water 0.0 0.5 1.0 1.5 2.0 0.00.51.01.52.02.53.03.5 ZnL / Zn' ZnL, mol/L x 10 -7 Original After 2 weeks After 6 weeks Strong type binding sites, L₁ Weak type binding sites, L₂ Note: [Zn'] is the SWASV la bile Zn concentration, [ZnL] is the complexed Zn concentration and K' is the co nditiona l sta bility constant (with respect to SWASV la bile Zn) in the Sca tcha rd linearization equation; (Chaminda et al, 2010). L][-][L ][ L][ T Zn'K'K Zn Zn = ' Fig. 2 - Scatchard linearization results on Zn titration for DOM in river water in the original sample, after 2 weeks, and after 6 weeks of incubation incubation were titrated separately with Zn to determine the conditional stability constant K' and total binding site concentration [L T ]. As shown in Fig. 2 and Table 1, Zn titration data obtained for DOM in the original river water sample seems to fit into two linear portions indicating two classes of binding sites. In contrast, DOMs in the samples after 2 and 6 weeks fitted to a single line, signifying only one type of binding site. It seems that the weak type of binding site in the river water DOM might have been degraded by biological activities. On the other hand, total ambient binding site concentration was reduced from 410 nM to 74 nM (82% decline) during 2 weeks of incubation, while there was no significant changes, either in K' or [L T ], observed afterwards. These results agreed with the SUVA results between 2 and 6 weeks of incubation. This further implied that the weak Zn binding sites in the river water DOM were significantly altered and more likely removed by biodegradation during the first 4 days of incubation period and became persistent. This degradation of binding sites was also in line with the increment of labile Zn concentration during the first 4 days of incubation. Consequently, labile Zn concentration increased from 3.7 ± 0.4 µg/L to 5.4 ± 0.3 µg/L during the first 4 days of incubation. From this preliminary experiment, it was concluded that Zn binding sites in DOM in river water became persistent within or less than two weeks. Biodegradability of zinc binding sites in DOM of WWTP effluent in river water Fig. 3 shows the variation of DOC, UV 254 and SUVA during the incubation of DOM in WWTP effluent with river water, where river water was pre-incubated for two DOC, mg/L SUVA 254 , L/mg/m UV 254 , 1/cm (Chaminda et al, 2010) Journal of Water and Environment Technology, Vol. 8, No.4, 2010 - 407 - Table 1 - Conditional stability constants, K', and the binding site concentrations, [L T ] for Zn during the incubation of river water DOM and DOM of WWTP effluents in river water Normalized binding site concentration Ambient binding site concentration log K 1 ' log K 2 ' [L 1,T ] [L 2,T ][L T ] = [L 1,T ] + [L 2,T ][L 1,T ] [L 2,T ] [L T ] = [L 1,T ] + [L 2,T ] Sample (nM/µM C) (nM) Original sample 6.9 6.1 0.79 1.6 2.4 140 ± 10 270 ± 20 410 ± 22 After 2 weeks 6.8 ND 1.0 ND 1.0 74 ± 17 ND 74 ± 17 River water DOM After 6 weeks 6.7 ND 1.1 ND 1.1 72 ± 8 ND 72 ± 8 Original sample 7.2 6.3 0.85 1.5 2.4 350 ± 7 610 ± 12 960 ± 14 After 4 days 7.2 6.3 0.84 1.4 2.2 340 ± 5 550 ± 8 890 ± 9 After 7 days 7.2 6.3 0.86 1.4 2.3 350 ± 7 550 ± 12 900 ± 14 DOM of WWTP effluent in river water After 1 month 7.2 6.3 0.83 1.3 2.1 290 ± 14 460 ± 22 750 ± 26 Note: K' = conditional stability constant with respect to electrochemical labile Zn (L/mol), [L 1,T ] and [L 2,T ] represent two types of binding sites: strong and weak, respectively; [L T ] = total binding site concentration; ND = not detected; Ambient binding site concentration = Normalized binding site concentration × DOC concentration (Titration condition: DOC = 3.5 mg/L; pH≈7; ionic strength = 0.02 M; temperature = 25ºC) weeks in advance. It was observed that the initial DOC concentration in river water decreased from 0.9 mg/L to 0.4 mg/L during the two weeks of pre-incubation period, indicating the availability and activities of microorganism in river water. However, as shown in Fig. 3, there was no considerable change of DOC in the mixture of DOM of WWTP effluents and pre-incubated river water within the first week of incubation. Slight DOC reduction (14%) was observed after one month of incubation. In contrast to the results of the preliminary experiment for river water, there was also no significant variation of SUVA in the incubation of DOM in WWTP effluent. The results for the conditional stability constants, K', and the concentrations of the binding sites, [L T ] for Zn are listed in Table 1. Scatchard linearization for the Zn titration data of four samples (original sample, after 4-days, 7-days and 1-month incubation) indicated the existence of two classes of binding sites: strong type, K 1 ' and weak type, K 2 ' (Fig. 4). As shown in Fig. 4, titration plots of four samples are close to each other, indicating no considerable difference in the conditional stability constants, which were deduced from the slopes of the lines in a particular binding site class. Conditional stability constant for the strong type of binding site remained at log K 1 ' = 7.2 while that in weak type remained at log K 2 ' = 6.3 for all the four samples throughout the incubation period. The conditional stability constant for Zn in this study was consistent with the values reported (log K 1 ' = 7.1 - 7.4 and log K 2 ' = 6.2 - 6.4) for the effluents from three different WWTPs (Chaminda et al. 2009). Moreover, the normalized binding site concentrations ([L T ] divided by DOC concentration), as listed in Table 1, show no significant changes (2.1 - 2.4 nM/ µM Journal of Water and Environment Technology, Vol. 8, No.4, 2010 - 408 - 0.00 0.01 0.02 0.03 0.04 0.05 0.06 0.07 0.08 0.09 0.10 0.0 1.0 2.0 3.0 4.0 5.0 6.0 0 5 10 15 20 25 30 35 UV 254, cm -1 DOC (mg L -1 ), SUVA (Lmg -1 m -1 ) Incubation period, days DOC, m g L -1 SUVA₂₅₄, L mg -1 m -1 UV₂₅₄, cm -1 Note: Error bars represent the standard deviation of triplicate measurements Fig. 3 - Variation of DOC, UV and SUVA during the incubation of DOM of WWTP effluent with river water 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5 ZnL / Zn' ZnL, mol/L x 10 -7 Original After 4 days After 7 days After 1month Strong type binding sites, L₁ Weak type binding sites, L₂ Note: [Zn'] is the SWASV labile Zn concentration, [ZnL] is the complexed Zn concentration and K' is the conditional stability constant (with respect to SWASV la b ile Zn) in the Sca tcha rd linearization equation; (Chaminda et al, 2010). L][-][L ][ L][ T Zn'K'K Zn Zn = ' Fig. 4 - Scatchard linearization results on Zn titration for DOM in WWTP effluent in the original sample, after 4 days, after 7 days and after 1 month of incubation C). As the bottles were covered to avoid exposure to light and as the incubation was done in a controlled temperature, it was assumed that the changes in metal complexation parameters were only due to the activities of indigenous microbes in the river water. Therefore, it can be hypothesized that, unlike binding sites in DOM of the river water, binding sites in DOM of WWTP effluents were less biodegradable. In other words, treatment of wastewater by activated sludge process has shifted the DOM composition to more biologically recalcitrant compounds in terms of zinc complexation. On the other hand, a further examination of the variation of labile Zn during the incubation revealed a slight increment of the labile Zn concentration (from 2.1 µg/L to 2.5 µg/L) in the sample after 1 month of incubation. Decline of total ambient binding site concentration (from 960 nM to 750 nM) by 22% after one month of incubation, might have caused this slight increase of labile Zn concentration. However, the residence time of urban rivers is usually shorter than one month and hence, it can be suggested that the DOM from WWTP effluent seems persistent in receiving waters of urban environment. It may be important to study the long-term biological alteration of organic matter for the prediction of heavy metal speciation in bay or seawater. It is also similarly important to study in details the microbial community involved in the biodegradation of metal binding sites to further understand the fate of metal speciation in the water environment. DOC, mg/L SUVA 254 , L/mg/m UV 254 , 1/cm (Chaminda et al, 2010) Journal of Water and Environment Technology, Vol. 8, No.4, 2010 - 409 - CONCLUSIONS Biodegradation experiments revealed that 82% of Zn binding sites of river water DOM was degraded during 2 weeks of incubation. In contrast, Zn binding sites of DOM in WWTP effluent as well as their conditional stability constants showed no significant changes during 1 month of incubation and hence, no significant effects on Zn complexation. This implies that metal (Zn) binding sites in the WWTP effluent were biologically resistant under the river water environment. ACKNOWLEDGEMENT This study was funded by Grants-in-Aid for Scientific Research (B) (20360239) of Japan Society for the Promotion of Science (JSPS). REFERENCES Brooks M. L., Meyer J. S. and Boese C. J. (2007). Toxicity of copper to larval Pimephales promelas in the presence of photodegraded natural dissolved organic matter, Can. J. Fish. Aquat. Sci. 64, 391-401. Chaminda G. G. T., Nakajima F. and Furumai H. (2008). Heavy metal (Zn and Cu) complexation and molecular size distribution in wastewater treatment plant effluent, Water Sci. and Technol.,58 (6), 1207-1213. Chaminda G. G. T., Nakajima F., Furumai H., Kasuga I. and Kurisu F. (2009). Zn and Cu complexation with DOM in wastewater treatment plant effluent, Proceeding of SETAC Europe 19th Annual Meeting, Göteborg, 40-41. Chaminda G. G. T., Nakajima F., Furumai H., Kasuga I. and Kurisu F. (2010). Comparison of Metal (Zn and Cu) Complexation Characteristics of DOM in Different Urban Wastewaters, Water Sci. and Technol., (in press). Cheng T. and Allen H. E. (2006). Comparison of zinc complexation properties of dissolved natural organic matter from different surface waters, Journal of Environmental Management, 80, 222-229. Ma H., Allen H. E. and Yin Y. (2001). Characterization of isolated fractions of dissolved organic matter from natural waters and a wastewater effluent, Water Research, 35 (4), 985-996. Sarathy V. and Allen H. E. (2005). Copper complexation by dissolved organic matter from surface water and wastewater effluent, Ecotoxicology and Environmental Safety, 61, 337-344. Shiller A. M., Duan S., van Erp P. and Bianchi T. S. (2006). Photo-oxidation of dissolved organic matter in river water and its effect on trace element speciation, Limnol. Oceanogr., 51 (4), 1716-1728. APPENDIX Model and linearization technique The metal titration data in this study were linearly transformed using Scatchard equation which was derived from the appropriate mass balance and conditional stability constant, K', relationship. Assuming 1:1 stoichiometry between electrochemical labile metal M' and binding sites L, the conditional stability constant can be represented by: Journal of Water and Environment Technology, Vol. 8, No.4, 2010 - 410 - ][][]'[ ii MLLM ↔+ (1) ]]['[ ][ ' i i i LM ML K = (2) ][][][ ,Tiii LMLL =+ i =1, 2 (3) where [M'] is the electrochemical labile metal concentration, [ML i ] is the complexed metal concentration, [L i ] is the free binding site concentration, [L i,T ] is the total available binding site concentration and K i ' is the conditional stability constant (with respect to ASV labile metal). The Scatchard linearization equation which is derived from the above equations (2) and (3) can be discribed as in equation (4). ]['][' ]'[ ][ , iiTii i MLKLK M ML −= (4) The Scatchard linearization consists of plotting the [ML i ]/[M'] against [ML i ]. The labile metal concentration, [M'], was electrochemically determined by ASV and [ML] was calculated from the mass balance equation ([M T ] – [M']), where [M T ] was the total metal concentration calculated from the initial concentration and the cumulative titrated volume. If the transformation results are linear (Fig. A1(a)), it indicates that the sample contains a class of ligands having one representative conditional stability constant K'. The slope of the line corresponds to the conditional stability constants K'. The y-intercept of the line gives the product of the conditional stability constant K' and total ligand concentration [L i,T ] and the x-intercept directly gives [L i,T ]. Distinctively separate linear regions within a transformation curve indicate that more than one class of ligands are available in the tested samples. Those lines can be separately extrapolated to determine the parameters related to each ligand class. Fig. A1(b) shows the idealized Scatchard plots of titration data for the two-ligand case. In this study, conditional stability constants were determined with respect to ASV labile, or electrochemically labile, metal. The ASV is sensitive to all quickly dissociating metal, including free metal ion, inorganic complexed metal and some weak organic complexed metals. Although the relationship between the free ion, [M 2+ ], and [M'] can be determined for inorganic ion pairs with simple equilibrium speciation modeling, it is not possible to correct [M'] for quickly dissociating organics, as their concentrations or side-reaction binding strengths are not known. Therefore, in this study K' values, determined from ASV without any corrections of [M'], were used. Journal of Water and Environment Technology, Vol. 8, No.4, 2010 - 411 - K' [L T ] [L T ] Slope = –K' [ML]/[M'] [ML] K 1 ' [L 1,T ] + K 2 ' [L 2,T ] Slope 1 = – K 1 ' [ML]/[M'] [ML] Slope 2 = – K 2 ' [L 1,T ][L 2,T ] K 2 ' [L T ] [L T ] K 2 ' [L 2,T ] K 1 ' [L 1,T ] K' [L T ] [L T ] Slope = –K' [ML]/[M'] [ML] K 1 ' [L 1,T ] + K 2 ' [L 2,T ] Slope 1 = – K 1 ' [ML]/[M'] [ML] Slope 2 = – K 2 ' [L 1,T ][L 2,T ] K 2 ' [L T ] [L T ] K 2 ' [L 2,T ] K 1 ' [L 1,T ] (a) (b) Fig. A1 - Interpretation of titration data with Scatchard linear transformation (a) one- ligand case (b) two-ligand case; K 1 ' and K 2 ' represent the conditional stability constants of strong and weak ligand, respectively; Dotted line shows the resolved component of each ligand in the two-ligand case . After 2 weeks 6 .8 ND 1.0 ND 1.0 74 ± 17 ND 74 ± 17 River water DOM After 6 weeks 6.7 ND 1.1 ND 1.1 72 ± 8 ND 72 ± 8 Original sample 7.2 6.3 0 .85 1.5 2.4 350. 350 ± 7 610 ± 12 960 ± 14 After 4 days 7.2 6.3 0 .84 1.4 2.2 340 ± 5 550 ± 8 890 ± 9 After 7 days 7.2 6.3 0 .86 1.4 2.3 350 ± 7 550 ± 12 900 ± 14 DOM of WWTP