Anaerobic degradation pathway and kinetics of domestic wastewater at low temperatures

Bioresource Technology 100 (2009) 6155–6162

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Anaerobic degradation pathway and kinetics of domestic wastewater at low temperatures Beni Lew a,*, Sheldon Tarre b, Michael Beliavski b, Michal Green b a b

Agricultural Research Organization, P.O. Box 6, 50250 Bet Dagan, Israel Faculty of Civil and Environmental Engineering, Technion 32000 – Haifa, Israel

a r t i c l e

i n f o

Article history: Received 13 January 2009 Received in revised form 21 June 2009 Accepted 22 June 2009 Available online 14 August 2009 Keywords: Anaerobic treatment Degradation kinetics Degradation pathway Domestic wastewater Temperature effect

a b s t r a c t The effect of temperatures below 20 °C (20, 15 and 10 °C) on the anaerobic degradation pathway and kinetics of domestic wastewater fractionated at different sizes was studied in a fluidized-bed batch reactor. The overall degradation pathway was characterized by a soluble fraction degrading according to zero-order kinetics and a colloidal fraction (between 0.45 and 4.5 lm) that first disintegrates into a particulate fraction smaller than 0.45 lm before finally degrading. The colloidal degradation processes follow a first-order kinetic. In contrast, suspended solids (bigger than 4.5 lm) degrade to soluble and colloidal fractions according to first-order kinetics. The colloidal fraction originating from suspended solids further degrades into soluble fraction. These soluble fractions have the same degradation kinetics as the original soluble fraction. The suspended solids degradation was highly affected by temperature, whereas the soluble fraction slightly affected and the colloidal fraction was not affected at all. On the other hand, the colloidal non-degradable fraction increased significantly with the decrease in temperature while the suspended solids slowly increased. The soluble non-degradable fraction was little affected by temperatures changes. Ó 2009 Elsevier Ltd. All rights reserved.

1. Introduction Anaerobic treatment of domestic wastewater is an attractive option for secondary wastewater treatment. The high costs of aeration and sludge handling associated with aerobic sewage treatment are lower. However, domestic wastewater has typically low concentrations of COD, resulting in relatively small methane production that is insufficient to heat the reactor to more favorable mesophilic temperatures. The relatively higher concentration of particulate matter in domestic wastewater has low degradation rates at psychrophilic temperatures (Elmitwalli et al., 2001; Lew et al., 2003; Singh and Viraraghavan, 2002). Therefore, anaerobic process efficiency is dependent on local ambient temperatures. At tropical temperatures, intensive anaerobic systems have been successfully applied and found wide acceptance for domestic wastewater treatment. COD removals above 70% have been observed by several authors (Chernicharo and Cardoso, 1999; Kalogo and Verstraete, 2000; Lew et al., 2003). At lower ambient temperatures, the application of anaerobic reactors for domestic wastewater treatment has been studied only in lab and pilot scale plants (Lettinga et al., 1981; Lew et al., 2004; Seghezzo et al., 1998). Dete-

* Corresponding author. Tel.: +972 39683453; fax: +972 34604704. E-mail address: [email protected] (B. Lew). 0960-8524/$ - see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.biortech.2009.06.073

rioration in overall performance was observed, with COD removals of 65% at 20 °C and 55–65% at 13–17 °C. Henze and Harremoes (1983) observed that temperature decrease affects not only the reaction rate, but also the yield coefficient, the decay rate and the specific growth rate of anaerobic bacteria. At low temperatures, the lower extracellular degradation rate and/or the decrease in the organic matter degradation promotes the accumulation of suspended matter in anaerobic reactor treating domestic wastewater (Elmitwalli et al., 2001; Lew et al., 2003; Singh and Viraraghavan, 2002). Moreover, the degradation of colloidal particles (20–30% of the total COD for domestic wastewater) was reported to be the rate limiting step in anaerobic digestion at lower temperatures (Mergaert et al., 1992; Wang, 1994). Understanding the anaerobic degradation processes (pathway, biodegradability and kinetics) at different temperatures is fundamental to the evaluation of the potential of intensive domestic wastewater anaerobic treatment. Bergamo et al. (2009) observed first-order degradation kinetics in a sequencing batch bioreactor with immobilized microorganisms, which decreased with temperature between 30 and 15 °C. Measurement of the anaerobic degradation process for domestic wastewater at different temperatures has also been carried out in attached biofilm experiments (Alderman et al., 1998; Lew et al., 2003; van der Last and Lettinga, 1992). In general, a decrease in total COD removal was observed with the decrease in temperature for the same retention time. In

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addition, at each temperature a major increase in the total COD removal was observed when increasing retention times shorter than 15 h. For retention times longer than 15 h, further increases in retention time led to a minor increase in the total COD removal. Based on the data available on degradation and degradability of wastewater under anaerobic conditions various digestion models have been built and can be found in the literature, each of them having its own potential and worth. No unified modeling framework for the anaerobic digestion process exists, however, the model mostly used today for anaerobic digestion processes is the ADM1 (Anaerobic Digestion Model No. 1), developed by the IWA anaerobic digestion modeling task-group (Batstone et al., 2002). The ADM1 provides an excellent generic model structure and parameterization to characterize anaerobic process dynamics under various reactor configurations and feeding protocols and can be employed to investigate the relationship between digester stability and acetoclastic methanogenic population dynamics. However, the model assumes a single type of particulate matter and relatively high temperatures. In this study, the anaerobic degradation (pathway, biodegradability and kinetics constants) of domestic wastewater at different low temperatures is conducted in a fluidized-bed reactor. An arbitrary division of the particulate matter into two fractions based on size is made, suspended solids (greater than 4.5 lm) and colloidal particles (from 0.45 to 4.5 lm). This division is applied in order to study the effect of colloidal particles on anaerobic degradation kinetics and develop a more comprehensive model. A single domestic wastewater soluble fraction is assumed (smaller than 0.45 lm) and its degradation was also studied.

2. Methods 2.1. Fluidized-bed reactor (FBR) Two batch fluidized-bed reactors (FBR) were constructed of plexiglass with a working volume of 3.1 l (5.0 cm diameter, 150.0 cm height). The reactors were filled with 400 ml sinter glass carrier (1.0–2.0 mm diameter) and fed with domestic wastewater after primary sedimentation from the Neve Sha’anan neighborhood located in Haifa, at the given constant experimental temperature. The wastewater can be classified as a medium strength domestic wastewater and its characteristics after primary sedimentation are given in Table 1. A pump provided for constant wastewater recirculation and fluidization of the biofilm covered sinter glass particles at 20 m/h upflow velocity. For each experimental temperature (20, 15 and 10 °C) a start-up period of at least three months was conducted in order for the Table 1 Characteristics of domestic wastewater from Neve Sha’anan neighborhood after primary sedimentation. Parameter

Unit

Average

St. Dev.

BOD Total COD Suspended solids COD Colloidal COD Soluble COD TSS VSS Total protein Suspended solids protein Colloidal protein Soluble protein Total carbohydrate Suspended solids carbohydrate Colloidal carbohydrate Soluble carbohydrate Ammonia

mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l

290.0 566.0 326.6 94.7 214.0 211.2 189.3 276.8 153.2 42.9 86.1 65.9 23.3 20.0 22.6 39.2

±110.0 ±242.8 ±80.8 ±16.8 ±48.0 ±115.9 ±90.3 ±17.3 ±10.8 ±26.7 ±34.2 ±49.9 ±27.1 ±40.0 ±23.5 ±7.8

reactors to reach steady-state conditions (van der Last and Lettinga, 1992; Zakkour et al., 2001). During the start-up period the FB reactors were continuously operated at a retention time of 6 h and fed with domestic wastewater. 2.2. The experimental procedure At the end of the start-up period (3 months), experiments were carried out in duplicate with the reactors operated in batch mode for 24 h, with continuous recirculation of wastewater at 20 m/h upflow velocity. Before the beginning of each experiment, the reactor was filled three times with the appropriate substrate at intervals of 15 min between each filling. The purpose of this procedure was to provide maximum adsorption of suspended solids and substrate to the biofilm before starting the experiment. For each temperature, the experiments were conducted with the same wastewater filtered at different sizes: (a) with the soluble fraction only (domestic wastewater filtered at 0.45 lm); (b) with soluble and colloidal fractions (domestic wastewater filtered at 4.5 lm); and (c) with the raw original domestic wastewater. In this way, the first experiment with the soluble fraction is carried out without the interference of bigger fractions. The soluble degradation rate was calculated assuming zero-order kinetics based on the constant decrease in the soluble concentration with time:

dS ¼ kt Xdt

ð1Þ

where S is the soluble concentration (mg COD/l), k is the zero-order constant, X is the microorganisms concentration (mg TSS/l) and t is the time (in hours). In the second experiment (colloidal and soluble fractions), the colloidal fraction degradation kinetics was determined assuming that the colloidal fraction is particulate matter that must undergo extracellular degradation to be transformed into soluble COD. Extracellular degradation is considered to be a single combined temperature dependent process with first-order kinetics (Batstone et al., 2002; Sanders et al., 2000) according to:

dP ¼ ktðP  Pn Þ dt

ð2Þ

where P is the particulate matter concentration (mg COD/l), k is the first-order constant, Pn is the non-degradable particulate fraction and t is the time (in hours). Moreover, the soluble fraction is assumed to degrade with the same degradation kinetics as in the first experiment and, differences between observed and assumed soluble fraction degradation are caused by the colloidal contribution to the soluble fraction. Two approaches to analyze the degradation of the soluble material originating from the colloidal fraction were taken: (1) assuming that part of it degrades with the same degradation kinetics as the soluble fraction from the first experiment and part is non-degradable; and (2) assuming first-order degradation kinetics (Eq. (2)). In the third experiment the FB batch reactor was operated using unseparated domestic wastewater. The suspended solids fraction was also considered to be particulate matter composed of a nondegradable fraction and one that must pass an extracellular degradation step that obeys first-order kinetics (Eq. (2)). The suspended solids matter that undergoes the extracellular degradation step can become wholly colloidal, wholly soluble or part colloidal and part soluble fraction. The suspended solids degradation pathway can be very complex and multiple assumptions can be made. In order to simplify the determination of the overall suspended solids degradation pathway, the contribution to the colloidal fraction was modeled first. Colloidal originating from suspended solids degrades according to first-order kinetics (Eq. (2)) with the following possible

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alternative pathways: A) degradation constant similar to the original degradable colloidal; (B) degradation rate constant different to the original degradable colloidal; (C) part of it is non-degradable; and (D) suspended solids degradation leads to soluble formation (directly). Each assumption can be considered by itself or coupled with others. The colloidal fraction originating from suspended solids can undergo a new extracellular step and also becoming soluble fraction (indirectly). Four different assumptions for the degradation of soluble originating from suspended solids (direct and indirect) were undertaken: similar to that of original soluble; similar to that of soluble fraction originating from colloidal degraded; part of the soluble is non-degradable; and different than that of the original soluble and the soluble originating from colloidal degraded. Each assumption was considered by itself or coupled with others, for direct and indirect soluble fractions originating from suspended solids. At all temperatures studied a similar microorganism concentration was used, about 21.8 g VSS/l sinter glass, or 2.8 g VSS/l reactor. 2.3. Analysis Collected samples were tested on a regular basis for COD, TSS and VSS. All samples were filtered and diluted whenever necessary and all were performed according to standard methods for the examination of water and wastewater (APHA, 1995). Filtered sample was measured from samples after filtration through 4.5 lm filter paper (Schleicher & Schuell 595). Soluble sample was measured from samples after filtration in a 0.45 lm membrane filter. The suspended solids and colloidal concentration were calculated by the differences between total and filtered samples; filtered and soluble samples, respectively. Reactor biomass concentration (TSS and VSS) and fluorescent in situ hybridization (FISH) were carried out at the end of each experiment at each temperature on the sinter glass. The FISH protocol described by Amann et al. (1990) was used. Microorganism communities in the reactor were investigated by using Cy3 or Cy5 labeled r-RNA targeted oligonucleotide probes purchased from Proligo for Eubacteria (EUB338 – GCTGCCTCCCGTAGGAGT) and Archaea (ARC915 – GTGCTCCCCCGCCAATTCCT) domains and; for acetoclastic methanogenic: methanosaeta (MX825 – TCGCACCGTGGCCGACACCTAGC) and methanosarcina species (SARCI551 – ACCAATAATCACGATCAC). The NON338 probe (ACTCCTACGGGAGGCAGC) was used as a control probe for EUB338. Specific microorganism populations were quantified by cell area measurement in CLSM images and results were expressed as percentage share of the total microorganism cell area, which was determined after 40 ,6-diamidino-2-phenylindole (DAPI) stained cells. Both DAPI and hybridization results were examined by an epifluorescence microscope Zeiss Axioskop. 10 microscope fields

200

Soluble COD (mg/l)

180 160 140 120

10 °C

100 80

15 °C

60

20 °C

40 20 0

0

5

10

Time (h)

15

Fig. 1. Soluble fraction degradation with time at 20, 15 and 10 °C.

20

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(corresponding to 500–1000 DAPI-stained cells) were counted to determine the average number of cells per sample (Daims et al., 2001). 3. Results and discussion At each temperature studied three sets of batch experiments were carried out in two FB reactors working in parallel using the same domestic wastewater separated into different sizes: (1) less than 0.45 lm (soluble fraction); (2) less than 4.5 lm (colloidal and soluble fraction); and (3) unseparated wastewater. The results of each fraction’s degradation at each temperature studied are shown in Figs. 1–3, where the lines joining the experimental data points were generated by MS Excel. 3.1. Soluble fraction only A constant soluble fraction degradation rate was observed at all temperatures in the first 5 h of the experiment (Fig. 1) and calculated according to a zero-order model (Eq. (1)) to be 0.26, 0.18 and 0.12 g COD g VSS1 d1 for 20, 15 and 10 °C, respectively. The soluble fraction was assumed to contain substrate ready for oxidation by microorganisms, a single step process. At the end of the degradation period, the soluble fraction remained constant. The remaining COD observed can be assumed to be the non-degradable fraction and it was calculated to be around 13% of the initial concentration for all temperatures studied with no temperature dependency. A similar non-degradable value was also observed by others (Bergamo et al., 2009; Elmitwalli et al., 2001). 3.2. Colloidal and soluble fractions The second experiment in the FBR reactor was conducted using domestic water screened to contain soluble and colloidal fractions. The colloidal fraction concentration remained constant in the first 4 h of the experiment at 20 °C (Fig. 2). Afterwards, the colloidal fraction concentration decreased in a first-order like fashion until the end of the experiment. A first-order like decrease was also observed at 15 and 10 °C, however, from the beginning of the experiments. The lag time in the colloidal degradation seen only at 20 °C was probably due to the different wastewaters used at different temperatures and not a biofilm condition. The first-order kinetic degradation constants and colloidal nondegradable fraction were determined assuming different values in Eq. (2) and comparing the results with experimental data. The values that gave the highest R2 were proposed to be the correct ones. A colloidal non-degradable fraction of 10%, 22% and 44% and a firstorder extracellular degradation constant rate of 4.65, 4.59 and 4.10 d1 were calculated for 20, 15 and 10 °C, respectively. The similarity in the first-order degradation rates obtained at different temperatures together with the sharp increase in the colloidal non-degradable fraction with the decrease in temperature suggests that the observed degradation of colloidal matter should not be assumed to be a simple one-step process. Colloidal degradation can be better described as a multistage process including particle disintegration and hydrolysis steps. The ‘‘degradation” observed at lower temperatures is probably colloidal size particles disintegrating into smaller soluble size particles rather than true biodegradation. The initial soluble concentration was considered to have a similar degradation rate as in Fig. 1; however, the change in the soluble fraction with time (Fig. 2) showed a different pattern. These differences were caused by the contribution from the colloidal fraction which became part of the soluble fraction. The two

B. Lew et al. / Bioresource Technology 100 (2009) 6155–6162

COD (mg/l)

6158 160 140 120 100 80 60 40 20 0

20 °C

0

2

4

6

8

10

12

Time (h)

COD (mg/l)

200

15 °C

150 100 50 0 0

5

10

Time (h)

15

20

COD (mg/l)

200

25

10 °C

150 100 50 0 0

5

10

15

20

25

Time (h) Fig. 2. Soluble (dashed line with open circles) and colloidal (full line with solid diamond) fractions degradation with time at 20, 15 and 10 °C.

300

20 °C

COD (mg/l)

250 200 150 100 50 0 0

5

10

15

20

25

Time (h) 250

15 °C

COD (mg/l)

200 150 100 50 0 0

5

10

15

20

25

Time (h)

COD (mg/l)

200

10 °C

150 100 50 0 0

5

10

Time (h)

15

20

25

Fig. 3. Soluble (dashed line with open circles), colloidal (full line with solid diamond) and suspended solids (dashed line with open squares) fractions degradation with time at 20, 15 and 10 °C.

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approaches to analyze the kinetics of the degradable soluble material originating from the colloidal fraction were: (1) assuming the same degradation kinetics as the soluble fraction from Fig. 1 (original soluble material); and (2) assuming a different degradation kinetic than the original soluble material. In the first approach to take in consideration the soluble concentration rise observed at 20 °C after 5 h (Fig. 2), a high fraction (62%) of non-degradable soluble originating from colloidal must be assumed, however, this value does not fit other experimental data at this temperature. The second approach (first-order degradation) is supported by the heterogeneous nature of domestic wastewater (Table 1). The soluble fraction is mainly composed of proteins, while the colloidal fraction is composed of carbohydrates and proteins of similar concentrations. Colloidal particles disintegrate in to particulate fractions smaller than 0.45 lm that are not actually soluble and still need to be further hydrolyzed and degraded according to different first-order kinetic constants. Based on this assumption, the highest R2 (higher than 0.98) between model and experimental data was observed for a first-order degradation constant of 4.92, 2.04 and 1.20 d1 for the soluble originating from colloidal material at 20, 15 and 10 °C, respectively. Similar first-order constants for extracellular degradation were observed by others (Eastman and Ferguson, 1981; Feng et al., 2005; Straub et al., 2005). Due to the high R2, the logic of the basic assumption and the suitability to other experiments with the original domestic wastewater and other domestic wastewater samples, this approach was assumed to be the correct biodegradation pathway. 3.3. Unseparated domestic wastewater – suspended solids, colloidal and soluble fractions The third experiment in the FBR reactor was conducted using unseparated domestic wastewater. In the experiments conducted at 20 °C (Fig. 3), the concentration of the suspended solids fraction remained constant in the first 2 h. After the first 2 h, the concentration decreased in a first-order like fashion until the end of the experiment. A first-order decrease was also observed at 15 and 10 °C, but from the beginning of the experiments. The suspended solids fraction, like the colloidal one, is considered to be particulate matter composed of a non-degradable fraction and one that must pass an extracellular degradation step that obeys first-order kinetics. A suspended solids non-degradable fraction of 13%, 26% and 38% and first-order extracellular process rate constant of 4.97, 3.00 and 1.00 d1 were calculated for 20, 15 and 10 °C, respectively. Similar non-degradable fractions and first-order extracellu-

lar process rate constants were observed by others at 20 °C (Elmitwalli et al., 2001; Feng et al., 2005). In the previous experiment for the colloidal material only (Fig. 2), the non-degradable fraction was found to be temperature dependent. However, for suspended solids the non-degradable and the degradable fractions were found to be temperature dependent as shown by the increase in non-degradable fraction and the decrease in the first-order extracellular process rate constant with the decrease in temperature. Four viable alternative pathways for the degradation of colloidal size particles originating from suspended solids degradation are shown in Table 2. In all the alternatives, the colloidal concentration at the beginning of the experiments was considered to degrade with the degradation kinetics calculated for original colloidal fraction (Fig. 2). For each alternative pathway tested, the first-order degradation rate and the percentage of each degradation product that gave the best R2 is given in Table 2. Alternatives I and II do not give a good curve fit (low R2) at 20 and 15 °C and were therefore eliminated. Alternatives III and IV gave a good curve fit at all temperatures studied. For repeated domestic wastewater batch experiments conducted (data not shown), alternative IV showed the best correlation for all temperatures and was assumed to be the correct biodegradation pathway. According to alternative IV, a decrease in the first-order constant degradation rate was observed for colloidal material originating from suspended solids with a decrease in temperature (Table 2), unlike the colloidal fraction from Fig. 2, indicating that the degradation of colloidal material originating from suspended solids is a temperature dependent process. Moreover, 56% of the suspended solids degraded becomes colloidal material and 44% becomes soluble material. At 15 and 10 °C, a higher percentage of the suspended solids becomes colloidal material, 70%. This shows the sensitivity of hydrolysis (degradation of suspended solids to soluble material) to temperature as opposed to physical disintegration processes (degradation of suspended solids to colloidal fraction). Based on the previous assumptions and results, the questions remaining for the suspended solids pathway are: (I) the kinetics of the soluble fraction originating from the colloidal fraction (from now on labeled soluble indirect from suspended solids); and (II) the kinetics of the soluble fraction originating directly from suspended solids degradation (from now on labeled soluble direct from suspended solids). Seven viable alternative pathways are presented for the degradation kinetics of soluble originating from suspended solids (direct and indirect): (I) partly similar to original soluble and partly non-

Table 2 Kinetics parameters calculated from the different degradation pathways of colloidal originating from suspended solids. T (°C)

Pathway

Assumptions

R2

k (d1)

Colloidal non-degradable

Colloidal degradable (%)

Sol

20

I II III IV

A, C B, C A, C, D B, D

0.206 0.498 0.898 0.968

4.65a 2.40 4.65a 4.60

0 60% 5% 0

100 40 55 56

0 0 40% 44%

15

I II III IV

A, C B, C A, C, D B, D

0.659 0.630 0.913 0.984

4.59a 2.4 4.59a 2.00

23% 0 29% 0

77 100 51 70

0 0 20% 30%

10

I II III IV

A, C B, C A, C, D B, D

0.790 0.775 0.934 0.996

4.10a 1.20 4.10a 1.30

20% 30% 35% 0

80 70 30 70

0 0 35% 30%

a

Previously calculated from the degradation of the original colloidal fraction.

From SS to

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degradable; (II) partly similar to soluble originating from colloidal and partly non-degradable; (III) part of the direct soluble similar to original soluble and part non-degradable; and part of the indirect soluble similar to soluble originating from colloidal and part nondegradable; (IV) partly of the direct soluble similar to soluble originating from colloidal and partly non-degradable; and part of the indirect soluble similar to original soluble and part non-degradable; (V) direct soluble with unique degradation kinetics and indirect soluble similar to soluble originating from colloidal; (VI) partly with unique degradation kinetics and partly non-degradable; and

(VII) direct soluble with unique degradation kinetics and indirect soluble also with unique degradation kinetics. Moreover, in all the alternatives the soluble concentration in the beginning of the experiments was considered to be the original soluble with the degradation kinetics previous calculated (Fig. 1). For each alternative the degradation kinetics rate and the percentage of non-degradable soluble fraction that gave the best R2 are shown in Table 3. Alternative I, with the highest R2 value observed (0.97) at all temperatures studied, is considered to be the correct pathway. According to alternative I, the degradable soluble

Table 3 Kinetics parameters calculated from the different degradation pathways of soluble originating from suspended solids – direct and indirect. T (°C)

Pathway

R2

K(direct) a

K(indirect) a

Non-degradable (direct)

Non-degradable (indirect)

20

I II III IV V VI VII

0.973 0.432 0.946 0.939 0.906 0.922 0.927

0.26 4.92 d1,b 0.26a 4.92 d1,b 26.4 d1 19.2 d1 18.24 d1

0.26 4.92 d-1,b 4.92 d1,b 0.26a 4.92 d1,b 19.2 d1 9.12 d1

2% 0 0 7% 0 7.5% 0

20% 0 0 3% 0 7.5% 0

15

I II III IV V VI VII

0.969 0.927 0.902 0.918 0.964 0.903 0.954

0.18a 2.04 d1,b 0.18a 2.04 d1,b 6.24 d1 4.8 d1 2.4 d1

0.18a 2.04 d1,b 2.04 d1,b 0.18a 2.04 d1,b 4.8 d1 7.2 d1

3% 0 0 8% 0 5% 0

22% 0 0 0 0 5% 0

10

I II III IV V VI VII

0.974 0.932 0.948 0.874 0.893 0.923 0.905

0.12a 1.20 d1,b 0.12a 1.20 d1,b 0.72 d1 0.72 d1 0.72 d1

0.12a 1.20 d1,b 1.20 d1,b 0.12a 1.20 d1,b 0.72 d1 0.72 d1

4% 0 0 8% 0 0 0

28% 0 0 0 0 0 0

a b

Previously calculated from the degradation of the original soluble fraction-unit: g COD gVSS1 d1. Previously calculated from the degradation of the soluble originating from colloidal.

SSdegradable

0.9%

20 °C

56.0%

Colldegradable

Colldegradable from SS 43.1% 21.0%

Soldegradalbe from Coll

79.0%

Solnon-degradalbe

Soldegradable from SS

Soldegradable

Degradation SSdegradable

1.5%

15 °C

70.0%

Colldegradable

Colldegradable from SS 28.5% 20.0%

Soldegradalbe from Coll

80.0%

Solnon-degradalbe

Soldegradable from SS

Soldegradable

Degradation SSdegradable

1.8%

10 °C

70.0%

Colldegradable

Colldegradable from SS 21.0% 30.0%

Soldegradalbe from Coll

70.0%

Solnon-degradalbe

Soldegradable from SS

Soldegradable

Degradation Fig. 4. Fate of the different wastewater fractions at 20, 15 and 10 °C.

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B. Lew et al. / Bioresource Technology 100 (2009) 6155–6162 Table 4 Kinetic degradation parameters and non-biodegradable percentage (%) of the different fractions at different temperatures. Fraction

Unit

20 °C

15 °C

10 °C

SS

d1 % d1 % d1 % g COD g VSS1 d1 % g COD g VSS1 d1 % g COD g VSS1 d1 % d1 %

4.63(±0.53) 18(±6) 4.65(±0.78) 13(±3) 4.65(±0.78) 0(±0) 0.30(±0.04) 13(±4) 0.30(±0.04) 2(±2) 0.30(±0.04) 21(±7) 4.92(±0.94) 0(±0)

3.01(±0.02) 29(±2) 4.57(±0.13) 27(±6) 2.00(±0.42) 0(±0) 0.16(±0.04) 13(±4) 0.16(±0.04) 5(±3) 0.16(±0.04) 20(±1) 2.04(±0.57) 0(±0)

1.06(±0.06) 39(±10) 4.08(±0.37) 45(±6) 1.32(±0.36) 0(±0) 0.13(±0.03) 16(±4) 0.13(±0.03) 6(±3) 0.13(±0.03) 30(±1) 1.20(±0.20) 0(±0)

Coll original Coll from SS Sol original Sol direct from SS Sol indirect from SS Sol from Coll

SS – Suspended Solids fraction; Coll – Colloidal fraction; Sol – Soluble fraction.

fraction originating directly from suspended solids is high (97%) at all temperatures. In contrast, the non-degradable fraction originating indirectly from suspended solids is significantly higher at 20 °C (20%) and increases to 28% with a drop in temperature to 10 °C. Similar observations and results were observed by others (Batstone et al., 2002; Elmitwalli et al., 2001; Yasui et al., 2005). Alternative number II did not give a good curve fit at 20 °C and is considered to be incorrect. Alternatives number III, IV, V, VI and VII gave good curve fit values at all temperatures studied, however, a good correlation was not observed for repeated domestic wastewater batch experiments conducted. To verify the parameters based on the above described experiments, five more experiments were conducted with complete domestic wastewater from Neve Sha’anan. The pathway that best suited wastewater anaerobic degradation at all temperatures studied is summarized in Fig. 4, and the kinetics parameters and the non-degradable fractions are shown in Table 4. For each temperature studied the biodegradation pathway is the same, the only difference is the percentage of material going at each pathway. 3.4. Temperature effect on observed parameters In general the overall pathway pictured in Fig. 4 qualitatively describes the degradation pathway for the temperature range investigated. In order to calculate the kinetic parameters for any given temperature (between 10 and 20 °C), the temperature activity coefficients were calculated based on the Arrhenius equation and are shown in Table 5. The temperature activity coefficient of the

Table 5 Temperature activity coefficient of the degradation constant, the equations derived from curve fitting for the non-degradable fraction and for the percentage pathway for the different domestic wastewater fractions at temperatures between 10 and 20 °C.

original colloidal fraction is not listed in Table 5 because no temperature dependence was found. The highest temperature coefficient was observed for the degradation rate of suspended solids fraction, 1.160, indicating that suspended solids degradation was the most sensitive process to temperature change. Moreover, the temperature coefficient observed for degradation of the three different particulates (suspended solids, colloidal originating from suspended solids and soluble originating from colloidal) are higher than the soluble fraction (1.079), the substrate assumed for methanogenesis. These results are in accordance with others (Lew et al., 2003 and Gujer and Zehnder, 1983). The influence of temperature on anaerobic processes is not limited to the reaction rate, but also reflects the extent of the anaerobic degradation process. The temperature effect on kinetic parameters is governed by the free energy law, however, the temperature effect on the non-degradable fraction itself and on each pathway’s percentage are calculated based on the best curve fitting. The particulate matter non-degradable fraction increases with the decrease in temperature, with the colloidal fraction being the most affected by temperature changes. The increase in the nondegradable colloidal fraction was about twice the increase in the non-degradable suspended solids fraction. In contrast, experimental results show that the original soluble non-degradable fraction stayed the same. 3.5. Temperature influence on microorganism population Quantification of microbial groups expressed as percentages in the FBR at different temperatures (20, 15 and 10 °C) is listed in Table 6. Fluorescent in situ hybridization (FISH) and DAPI results showed a decrease in microorganism viability (total archaea and eubacteria counted by FISH as opposed to the DAPI count) with the decrease in temperature, from 40% at 20 °C to 18% at 10 °C. Although the results are qualitative, they suggest that at lower temperature a lower concentration of viable microorganisms is present in the reactor. Probably a higher amount of microorganisms is present in the reactor at lower temperatures. FISH results show that archaeal cells dominated at all temperatures studied, around 70% of the total active microorganisms. A slight decrease in archaeal cells with the decrease in temperature was observed, from 76% at 20 °C to 68% at 10 °C. Similar results were observed by Gomec et al. (2004). Probably temperature changes between 20 to 10 °C have little affect on microorganisms ratio in an anaerobic reactor treating domestic wastewater. It was also observed that acetoclastic methanosaeta species (MX825) were the major methanogenic archaea present at all temperatures studied, around 84% of ARC915, with almost no change with temperature decrease. This high acetoclastic species concentration can be explained by the fact that seventy percent of the methane produced in anaerobic processes comes from the decarboxylation of acetate. Similar results of methanosaeta species being the major archaea present in anaerobic sludge treating domestic wastewater has been observed by others (Araujo et al., 2001; Sanz et al., 2002). No acetoclastic methanosarcinaceae species (SARCI551) were observed at all temperatures studied.

Fraction

Temperature activity coefficient

Nondegradable fraction

% Pathway

Suspended solids Original colloidal Colloidal from suspended solids Original soluble Soluble direct from suspended solids Soluble indirect from suspended solids Soluble from colloidal

1.160 – 1.124

2.1  T + 60.2 3.2  T + 76.3 –

– – 1.4  T + 86.3

1.079 1.079

0.3  T + 18.5 0.4  T + 10.3

– 2.2  T  2.3

Temperature

1.079

0.9  T + 37.2

0.5  T + 56.9

1.141

–

–

10 °C 15 °C 20 °C

Table 6 FISH results at different temperatures. Viability

40.4 (±2.0) 29.4 (±5.5) 17.7 (±5.5)

Viable microorganisms % Eubacteria

% Archaea

25.4 (±3.5) 28.9 (±3.3) 32.2 (±2.7)

74.6 (±3.5) 71.1 (±3.3) 67.8 (±2.7)

Methanosaeta % of Archaea 88.0 (±5.3) 81.5 (±3.1) 85.7 (±4.2)

6162

B. Lew et al. / Bioresource Technology 100 (2009) 6155–6162

4. Conclusions The importance of considering multiple types of particulate matter at low temperatures is shown. The colloidal fraction first disintegrates (a physical process not affected by temperature) into a particulate fraction smaller than 4.5 lm before finally being degraded, both processes in a first-order manner. Suspended solids partly degrades into soluble fraction and partly to colloidal degradable fraction, which further degrades to soluble. Soluble originating from suspended solids degrades with degradation kinetics similar to the original soluble fraction. The same overall degradation pathway was found to fit experimental data for all temperatures studied with the only difference being the percentage of material going to each pathway and the degradation kinetic constants.

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