PHA based denitrification: Municipal wastewater vs. acetate

Bioresource Technology 132 (2013) 28–37

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PHA based denitrification: Municipal wastewater vs. acetate Eli Krasnits, Michael Beliavsky, Sheldon Tarre, Michal Green ⇑ Faculty of Civil and Environmental Engineering, Technion, Haifa 32000, Israel

h i g h l i g h t s " PHA based denitrification (PBD) was studied with real wastewater versus acetate. " Focusing on PBD as part of two sludge system assuming chemical removal of residual P. " Highnitrate removal was observed with municipal wastewater (39–53 mg N/L). " Entrapped particulate matter contributed to the reducing power for denitrification. " Low COD/N ratios for denitrification were observed with acetate and wastewater.

a r t i c l e

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Article history: Received 28 August 2012 Received in revised form 11 November 2012 Accepted 19 November 2012 Available online 17 January 2013 Keywords: Bacterial storage polymers PHA Anaerobic storage Denitrification with PHA as the electron donor Two sludge system for wastewater treatment

a b s t r a c t Denitrification of municipal wastewater based on bacterial storage polymers–Polyhydroxyalkanoates (PHA) – was investigated in biofilm sequencing batch reactors, as a part of a two sludge system for wastewater treatment and in comparison to acetate based synthetic wastewater. The results show that PHA based denitrification (PBD) of real wastewater can be a viable alternative, especially for wastewater with low COD/N ratio, without the need for external carbon source addition. High nitrate removal capacity of about 40–50 mg N/L with a low COD/N requirement of about 4–5, were observed. It was found that entrapped particulate organic matter contributed additional reducing power, on top of the storage materials, thus allowing for the high nitrate reduction capacity. Daily removal rates were similar to those of extensive treatment systems (0.24–0.31 gr N/L reactord). Large differences in storage yield and composition between biomass grown on synthetic and municipal wastewater were observed. Ó 2012 Elsevier Ltd. All rights reserved.

1. Introduction Removal of nitrogen compounds from wastewater is usually performed by nitrification followed by denitrification. The combination of these two processes has some drawbacks, mainly: (1) oxygen is essential for nitrification and inhibits denitrification and (2) removal of organics prior to nitrification, in order to achieve efficient nitrification, results in electron donor deficiency for subsequent denitrification. Common solutions to the above difficulties are implemented in various continuous and batch configurations that include the addition of external COD sources as well as the separation of the anaerobic, anoxic and aerobic units with high recirculation rates between them, resulting in increased operational costs for energy and chemicals. In recent years, several innovative approaches have been investigated including simulta⇑ Corresponding author. Tel.: +972 54 4731028; fax: +972 48325373. E-mail address: [email protected] (M. Green). 0960-8524/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biortech.2012.11.074

neous nitrification–denitrification (SND), ANAMMOX, aerobic denitrification based technologies and others. Although some of these approaches show promising results, they are rarely implemented in commercial scale due to high operational costs, low removal rates, complex operation conditions and high sensitivity to wastewater composition and environmental conditions (Paredes et al., 2007; Ahn 2006). Intracellular carbon storage of polyhydroxyalkenoates (PHA) in heterotrophic bacteria is a well-known phenomenon and was found to be a major metabolic pathway in conventional wastewater treatment plants (Carucci et al., 2001). The concept of biodegradable COD storage was incorporated into the last activated sludge model ASM3 (Gujer et al., 1999). PHA storage in wastewater treatment facilities occurs mainly due to fluctuations in electron donor and electron acceptor availabilities. Under such conditions, bacteria capable of carbon storage gain competitive advantage over bacteria without the capacity of substrate storage (Van Loosdrecht et al., 1997). Two major groups of bacteria are capable of storage:

E. Krasnits et al. / Bioresource Technology 132 (2013) 28–37

29

Fig. 1. Simplified schematic drawing of the main treatment stages in a PBD based two-sludge system.

polyphosphate accumulating organisms – PAOs (responsible for the phosphate removal in enhanced biological phosphorus removal (EBPR) systems) and glycogen accumulating organisms – GAOs. In GAO’s the main energy source used for substrate uptake is glycogen through glycolysis, while in PAO’s it is polyphosphate cleavage. The co-existence of the two groups in wastewater treatment plants is well documented, and the relative abundance of each group depends on many factors including composition of organic matter, P/COD ratio, pH, temperature and others (Oehmen et al., 2005). Previous studies, in which simple organic compounds such as acetate were used as substrate for storage, have shown that intracellular polymers (mainly poly-3-hydroxybutirate – PHB and poly3-hydroxyvalerate – PHV) in GAOs and PAOs can serve as effective electron donors for denitrification (Kuba et al., 1996b; Qin et al., 2005; Bernat et al., 2008; Ciggin et al., 2007).

1.1. Two sludge systems Two sludge systems have been previously proposed for simultaneous phosphorus removal and PHA based denitrification (PBD), with usually an emphasis on phosphorus removal capability. Two sludge systems include three separate stages that can be implemented in either batch or continuous mode: I. Anaerobic storage of COD and release of orthophosphate followed by II. Nitrification by a separate nitrifying biomass and III. Phosphate uptake under denitrification conditions, with stored PHA serving as the electron donor. Two sludge systems have been studied under a wide range of operating conditions with synthetic acetate-based wastewater (Kuba et al., 1996a,b; Hughes et al., 2006; Carvalho et al., 2007; Zafriadis et al., 2009; Jiang et al., 2010), and to some extent with real wastewater from different sources such as piggery wastewater (Bortone et al., 1994), municipal wastewater (Sorm et al., 1996), and abattoir wastewater after anaerobic treatment with added VFA (Zhou et al., 2008a). These studies have provided information regarding COD, nitrogen and phosphorus removal potential with different wastewaters and under various operating conditions, however, the composition and stoichiometry of storage biopolymers of PBD have only been investigated using synthetic wastewaters such as acetate and propionate (Satoh et al., 1992; Pereira et al., 1996; Yagci et al., 2003; Zhou et al., 2008a,b; Pijuan et al., 2008; Zeng et al., 2003a; Carvalho et al., 2007; Zafriadis et al., 2009; Wang et al., 2011). Moreover, none of the previous studies investigated the nitrogen removal potential of two sludge systems with municipal wastewater. The main potential advantages of two sludge systems include efficient use of COD for denitrification and savings in aeration. Other advantages include lower cell yield with the corresponding lower sludge production, and simultaneous phosphorus and nitrogen removal. In a previous study it was estimated that PBD can reduce the COD/N requirement, O2 requirement and sludge production by 50%, 30% and 50%, respectively (Kuba et al., 1996b). The drawbacks of two sludge systems include potential presence of ammonia in the effluent due to the residual liquid volume in the

storage/denitrification reactor after the storage stage (Kuba et al., 1996b). Moreover, with regard to real wastewater, two sludge systems may suffer from reduced efficiency due to poor selection towards biomass capable of anaerobic carbon storage. Incomplete removal of biodegradable COD during anaerobic storage results in leftover electron donor available for the denitrification stage. Since microbial growth on internally stored PHA can be as much as 6 times slower than on external COD sources (Karahan et al., 2008), the presence of external COD during the denitrification stage may significantly reduce the system’s selection towards denitrifying biomass capable of anaerobic carbon storage, which in turn reduces the overall denitrification potential (Schuler and Jenkins, 2003). 1.2. Intrinsic differences between real and synthetic wastewater for PBD While soluble COD sources such as acetate are easily taken up and stored by biomass under anaerobic conditions, most of the COD in wastewater is usually found in a particulate phase and has to undergo hydrolysis prior to being taken up and stored as PHA. Moreover, not all the soluble COD found in wastewater is biodegradable and susceptible to storage under anaerobic conditions. In most of the previous studies on anaerobic COD storage using synthetic COD sources such as acetate or propionate and under suitable conditions, the COD was taken up completely during the anaerobic stage and therefore the following aerobic or anoxic stage was based exclusively on stored COD (Kuba et al., 1996a,b; Hughes et al., 2006; Carvalho et al., 2007; Zafriadis et al., 2009; Jiang et al., 2010). These conditions are unrealistic when dealing with real wastewater, mainly due to the presence of suspended solids which are trapped during the storage stage and can serve as electron donors during the following denitrification stage, and also due to incomplete drainage at the end of the storage stage (residual volume) which leads to the transfer of COD that is not susceptible to storage under anaerobic conditions but yet bioavailable to the following denitrification stage. In contrast to common EBPR systems which are usually aimed at phosphorous removal mainly, this research study is aimed at highly efficient nitrogen removal system, with the idea that if necessary residual phosphorous can be chemically removed. The experimental system in this research consisted of biofilm reactors that allow for better drainage than suspended sludge reactors, thus minimizing both ammonia concentration in the effluent and external COD in the denitrification stage. Nitrogen removal capacity, as well as storage polymers stoichiometry, composition and kinetics are emphasized. 2. Methods 2.1. Biofilm reactors Two identical reactors (R1 and R2) with working volume each of 2.1 L were seeded with biomass from the Haifa municipal activated sludge WWTP. Each reactor contained 283 HDPE carriers (AqwiseÒ

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E. Krasnits et al. / Bioresource Technology 132 (2013) 28–37

Three separate batch tests were performed in duplicates with sludge from R1, sludge from R2 during period 1 and sludge from R2 during period 2. Sludge was removed from 20 plastic carriers taken from a biofilm reactor (either R1 or R2) at the end of the storage stage. The sludge was suspended in a 500 mL hermetically sealed and constantly stirred vessel containing nitrate solution (50 mg NO3–N/L, 10 mg PO4–P/L). The vessel’s headspace was filled with N2 gas and the liquor was stirred under anoxic conditions for 2 h in order to simulate the denitrification stage. Biomass and liquor samples were taken in 30 min intervals and the NO3–N and NO2–N concentrations as well as the PHA and glycogen contents in the sludge were analyzed. 2.3. N2O emission tests

Fig. 2. Schematic drawing of the biofilm reactor.

L = 14 mm 600 m2/m3). A removable plastic grid was placed 10 cm below water surface in each reactor to prevent washout of carriers. Plastic balls (d = 2 cm) were used to cover the water surface in order to minimize CO2 stripping and exposure of the liquor to atmospheric oxygen. Each reactor had its own recycling pump which was operated at an average flow rate of 1 L/min during both storage and denitrification stages (Fig. 2). The reactors were kept in a temperature controlled room at 20(±2) °C. Both reactors were operated in a batch-wise mode (Fig. 3) and under the same operating conditions except for the composition of the feeding solution used in the storage stage: at the beginning of the anaerobic storage stage, R1 reactor was fed with acetate based synthetic wastewater, while R2 reactor was fed with municipal wastewater after primary sedimentation. Municipal wastewater from two sources was used in this study: wastewater collected from the Neve Sha’anan (N.S.) suburb in Haifa and wastewater from the Haifa activated sludge WWTP. Synthetic and municipal wastewater characteristics are presented in Table 1.The acetate based synthetic wastewater consisted of tap water enriched with 25 mg/L yeast extract, 200 mg/L acetate, 60 mg/L NH3–N and 5 mg/L PO4–P. The concentration of the acetate was selected to give a filtered COD (0.45 l) similar to that of the real wastewater – N.S. wastewater. Since an integrated nitrification reactor was not studied in this research project, at the end of the storage stage the reactors were drained and the effluent was discarded. At the beginning of the denitrification stage, the reactors were fed with nitrate solution. The nitrate solution consisted of tap water to which 50–60 mg/L NO3–N and 10–20 mg/L PO4–P were added (zero COD). During the experimental period, the influent and effluent of the storage and denitrification stages were monitored for COD, CODf, TSS, VSS, NO3–N, NO2–N, PO4–P, acetate and pH. 2.2. Batch tests In order to study the kinetics of PHA degradation without biofilm diffusion limitations, PHA degradation rates were measured in batch tests.

N2O emission during PBD was analyzed using a novel multiphase ATR-LP-FTIR system consisting of a horizontal ATR crystal attached to an FTIR (Tensor27, Bruker) and a long path IR cell attached to a higher resolution FTIR (Vertex 70, Bruker), according to a method described by Segal-Rosenheimer and Dubowski (2007). Biomass samples from the biofilm reactors were removed from the carriers at the end of the storage stage and suspended in 50 mL tap water in a stirred vessel. The biomass containing solution was purged with N2 gas to remove any dissolved oxygen. The pH in the vessel was adjusted to 7.0 and the solution was spiked with nitrate to achieve a final concentration of 50 mg N/L. The headspace content of the vessel was constantly recirculated through the long path IR cell during a 3 h period. IR N2O absorbance in the gaseous phase was measured in 10-min intervals and the absorption data were recorded and analyzed using MATLAB. NO3 and NO2 concentrations in the aqueous phase were measured at the end of the 3 h experimental period. 2.4. Analyses Influent, effluent and reactor concentrations for COD, CODf, TSS, VSS, NO3–N, NO2–N, PO4–P, Acetate and NH3–N, as well as pH, were tested regularly. All analyses were carried out according to standard methods (APHA, 1995). NO3–N, NO2–N, PO4–P and acetate were tested using the ion chromatography suppressed anion method. Biomass PHB, PHV and poly-3-hydroxy-2-methylvalerate – PH2MV concentrations (w/w) were determined according to (Oehmen et al., 2005) with a slight modification: For each analysis, biomass from 5 random carriers (equivalent to 60–120 mg VSS) was removed and transferred into sulfuric acid solution (pH < 2) in order to stop microbial activity. The solution was centrifuged and the supernatant was discarded. The biomass was then frozen until analysis. All biomass samples were dried at 105 °C for 2 h, weighed and placed in Pyrex tubes containing 2 mL chloroform and 2 mL methanol with sulfuric acid. For PHB and PHV determination, methanol with 3% sulfuric acid was used and the tubes were heated at 100 °C for 3.5 h. For peak calibration, R-3-hydroxybutyric acid (3HB) and R-3-hydroxyvaleric acid (3HV) copolymer (88:12) (Sigma–Aldrich) was used.

Fig. 3. Successive stages during a cycle (186 min) in R1 and R2.

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E. Krasnits et al. / Bioresource Technology 132 (2013) 28–37 Table 1 Characterization of the wastewaters and reactors used in the experiments (numbers in brackets are standard deviations). Wastewaters (concentrations in mg/L) Source

CODT

CODF*

TSS

VSS

NH3–N

PO4–P

pH

Acetate based wastewater (synthetic) Neve Sheanan wastewater Haifa WWTP wastewater

275 (6) 491 (93) 712 (201)

275 (6) 267 (49) 513 (199)

– 165 (18) 150 (53)

– 140 (18) 132 (39)

60 (2) 67 (10) 85 (11)

4.7 (0.9) 5.6 (0.5) 15.9 (1.8)

7.0 (0.1) 7.4 (0.1) 7.5 (0.1)

Reactors

R1 R2 period 1 R2 period 2 *

COD source

Treated volume per cycle (L)

Sludge concentration as TSS and (VSS) mg/L reactor

Operation days

Acetate Neve-Sha’anan suburb WW Haifa municipal WW

1.45 1.63 1.53

6070 (4990) 6170 (5470) 8040 (6620)

1–70 1–30 31–52

CODF = 0.45 l filtered COD.

For PH2MV determination, methanol with 10% sulfuric acid was used, the tubes were heated for 20 h and hydroxyhexanoic acid was used for peak calibration. After cooling, 2.5 mL deionized water were added to each tube and the tubes were shaken vigorously. After phase separation, the chloroform phase was dried over NaSO4 and filtered through a 0.2 lm syringe driven PTFE filter to remove cell debris. Benzoic acid was used as the internal standard. 1.5 lL of the chloroform phase were injected to HP 6890 GC equipped with FID detector. A 30 m DB5, 0.32 mm, 0.25 lm film column was used. For verification purposes a few samples were also injected to Thermo Scientific Focus GC equipped with ISQMS detector. A 30 m, 0.25 mm, 0.25 lm film column was used. Glycogen was determined using an enzymatic glucose assay kit (Sigma AldrichÒ GHK20). Biomass samples were dried for 2 h at 105 °C. The dried biomass samples were placed in sealed Pyrex tubes containing 0.6 M HCl solution and heated at 100 °C for 3 h. The resulting solution was neutralized using 10 M KOH solution and centrifuged to remove particulate matter. The supernatant was used to determine the glucose concentration using the enzymatic assay kit. 2.5. Calculation of storage compounds concentration (PHA and Glycogen) The major constituents of microbial PHA are considered to be PHB, PHV and poly-3-hydroxy-2-methylvalerate – PH2MV (Satoh et al., 1992). Due to the fact that the concentrations of microbial PH2MV detected were insignificant (<0.05% w/w of total VSS), the PHA concentrations reported in this paper refer to the concentrations of PHB and PHV only. PHA and glycogen concentrations are given in this paper as mg COD either per gram VSS or per liter reactor. The storage compounds concentration per liter reactor were calculated based on the measured PHA and glycogen concentrations (grPHA/grVSS or grGlycogen/grVSS) multiplied by the total VSS content in the reactor (283 carriers) and divided by the reactor volume (2.1 L). The term L reactor is used to represent a liter of the total reactor volume (including working volume and carriers) as opposed to L which represents a liter of liquid. 3. Results and discussion 3.1. Biofilm reactors characteristics and residual liquid volume after drainage Although both reactors used during the experimental period (one for acetic based wastewater and the other for real municipal wastewater) had the same total volume and contained the same amount of carriers, the treated volumes per cycle (presented in Table 1) were different, due to different TSS content and sludge water

retaining properties. The higher sludge concentration developed in R2 during feed with Haifa wastewater, is due to higher organic loads as well as increased accumulation of wastewater solids. The residual liquid volume that remained in the reactor fed with acetate based synthetic wastewater (R1) after drainage was 15% of the total liquor volume. In the reactor fed with real wastewaters (R2), the residual liquid volume fraction was 10% during period 1 (using N.S. wastewater) and 13% during period 2 (using Haifa wastewater). These residual liquor volumes are substantially lower than those of standard suspended sludge systems, which are crucial for attaining highly selected biomass capable of storage, as described in the introduction.

3.2. Storage stage 3.2.1. Concentrations profiles Typical profiles of total COD (CODT), filtered COD (CODF), phosphate, PHA, and glycogen during the storage stage are presented in Fig. 4, and average values for the entire experimental period are given in Tables 2 and 3. PHA and glycogen, extracted from the biomass are given as mg COD/L reactor (see Section 2.5). In general, the storage stage profiles show removal of COD (total and filtered), formation of PHA (PHB and PHV), release of orthophosphate (hydrolysis of poly-P as the energy source for the substrate uptake) and consumption of glycogen (glycolysis as the energy source for substrate uptake). GAOs and to some extent PAOs utilize glycogen in order to generate the ATP and reducing equivalents required for substrate uptake and storage under anaerobic conditions (Yagci et al., 2003; Zeng et al., 2003b; Zhou et al., 2008b; Pijuan et al., 2008). The pyruvate produced through glycolysis is transformed into acetyl-CoA or propionyl-CoA and stored as PHA, so that the net COD balance during storage is maintained as follows: substrate removed by storage = PHA stored  glycogen consumed (Table 3 column 6). The operational conditions in R1 and R2 were selected in order to simulate conditions similar to those that occur during treatment of real wastewater, where the presence of external COD during denitrification is unavoidable (due to entrapment of suspended solids and also COD of the residual liquid volume). In the complete two sludge system described in Fig. 1, the effluent COD after the storage stage is transferred to the nitrifying reactor (not included in this research project). At the end of the storage stage the total and filtered COD were not completely removed under daily organic loads of 300, 557 and 627 mg CODT/grVSSd in R1, R2 during period1 and R2 during period 2 respectively (Table 2). For comparison purposes, the daily filtered COD load was identical in R1 during feeding with acetate based wastewater (305 mg CODF/grVSSd) and in R2 during feeding with N.S. wastewater (303 mg CODF/ grVSSd). Based on 8 cycles per day, the specific total COD removal rate in R1 was 0.14 mg COD removed/grVSSd, while in R2 with real

E. Krasnits et al. / Bioresource Technology 132 (2013) 28–37

COD and 0.84 for filtered COD (net PHA stored from external substrate, i.e., after subtracting glycogen consumed). For N.S. wastewater (R2-period 1), the fraction of net PHA stored per COD removed was much lower when based on total COD removed – 0.48, but quite similar to R1 when based on filtered COD removed – 0.82. For Haifa wastewater (R2-period 2) the fraction of PHA stored per COD removed was much lower for both total and filtered COD: 0.38 for COD total removed and 0.4 for filtered COD removed. As can be seen from Table 3, glycogen consumption during the storage stage using real wastewaters was much higher than with the synthetic wastewater. The ratios of glycogen consumed per net PHA stored (grCOD/grCOD) were 0.37 for acetate and 0.71 and 1.53 for N.S. and Haifa wastewater respectively. This phenomenon, observed for both N.S. and Haifa wastewaters, is probably due to a higher GAO fraction developed under conditions of

wastewater, the specific removal rates were higher: 0.24 and 0.21 mg COD removed/grVSSd during periods 1 and 2 respectively. The specific filtered COD removal rate in R1 was 0.16 mg COD removed/grVSSd while in R2, the rates were 0.14 and 0.18 mg COD removed/grVSSd during periods 1 and 2 respectively. Although the specific total COD removal rates in R2 were 71% higher than in R1 during period 1 and 50% higher than in R1 during period 2, the specific filtered COD removal rates were similar in R1 and in R2 during both periods. 3.2.2. Storage yields and contents For the acetate fed reactor (R1) the fraction of net PHA stored per COD removed, as presented in Table 3, was 0.95 for total

R1 acetate based synthetic wastewater

300

14

200

10

150

8 6

100

4 50 0

2 0

0.25

0.5

0.75

1

PO4-P (mg/L)

COD, CODf (mg/L)

12

160

440

140

400

120

360 320

100

280

80

240

60

200

40

160

20

120 80

0

0 0

0.25

0.5

0.75

1

t (hr)

t (hr) 500

180

480

12

R2 N.S. wastewater

R2 N.S. wastewater

200

480

COD, CODf (mg/L)

8

300 250

6

200 4

150 100

PO4-P (mg/L)

350

2

50 0 0.25

0.5

0.75

1

0

460 160

440 420

120

400 380

80

360 340

40

320 0

0

0.25

t (hr) 700

24

0.75

1

300

R2 Haifa wastewater

200

660

600

18

550

14

500

12 10

450

8

400

6 4

350

2

300 0.25

0.5

t (hr)

0.75

1

0

PO4-P (mg/L)

16

640 160

620 600

120

580 560

80

540 520

40

Gly (mgCOD/Lreactor)

20

PHB, PHV (mgCOD /Lreactor)

22

650

COD, CODf (mg/L)

0.5

t (hr)

R2 Haifa wastewater

0

Gly (mgCOD/Lreactor)

10

400

PHB, PHV (mgCOD /Lreactor)

450

0

Gly (mgCOD /Lreactor)

250

R1 acetate based synthetic wastewater

520

16

PHB, PHV (mgCOD /Lreactor)

32

500 0

0

0.25

0.5

0.75

1

480

t (hr)

Fig. 4. Typical profiles of COD: (j), CODF (D), PO4–P: (N), PHB: (e), PHV: (h) and Glycogen: (x), during the storage stage on acetate based synthetic wastewater in R1 (day 55), N.S. wastewater in R2 (day 27) and Haifa wastewater in R2 (day 48).

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E. Krasnits et al. / Bioresource Technology 132 (2013) 28–37 Table 2 Storage stage influent and effluent characteristics – average values (numbers in brackets are standard deviations). Reactor

Storage stage influent (mg/L)

R1 acetate WW R2-1 N.S. WW R2-2 Haifa WW *

Storage stage effluent (mg/L)

CODT

CODF*

NH3–N

PO4–P

pH

CODT

CODF*

NH3–N

PO4–P

pH

275 (6) 491 (93) 712 (201)

275 (6) 267 (49) 513 (199)

60 (3.5) 62 (7.7) 105 (11.3)

4.7 (0.9) 5.6 (0.5) 15.9 (1.8)

7.0 (0.1) 7.4 (0.1) 7.5 (0.1)

152 (18) 283 (78) 470 (176)

135 (18) 145 (27) 303 (148)

44 (4.1) 45 (6.3) 85 (15.8)

11.8 (2.6) 10.7 (1.9) 22.5 (3.3)

7.6 (0.2) 7.7 (0.1) 7.9 (0.1)

CODF – filtered COD = COD after filtration through a 0.45 l filter.

Table 3 Storage yield and composition – average values. During storage stage

End of storage stage

mg COD/L reactor*

* **

mg COD/mg COD

% w/w

Units

CODT removed

CODF removed

PHA stored

Glycogen consumed

Net PHA stored from external substrate**

Net PHA stored/ CODT removed

Net PHA stored/ CODF removed

PHA/ VSS

PHV/ PHA

Gly/ VSS

R1

85 (24)

97 (27)

30 (12)

81

0.95

0.84

R2 period 1 R2period 2

161 (32)

95 (45)

55 (22)

78

0.48

0.82

176 (57)

153 (61)

111 (24) 133 (42) 157 (58)

95 (39)

62

0.36

0.41

7.17 (1.23) 3.84 (0.66) 2.10 (0.48)

17 (4.5) 50 (8.9) 57 (8.7)

2.78 (0.89) 5.91 (1.61) 6.98 (2.14)



treatedv olume COD removed/L reactor was calculated as COD remov ed ðmg=LÞ 2:1L Calculated as COD PHA stored  COD Glycogen consumed.

ðLÞ

. For treated_volume – see Table 1.

higher COD/P. However, since the ratio of P released to net PHA stored (mgP/grCOD) was quite similar with acetate and wastewater – 61 with acetate and 51 and 74 with N.S and Haifa wastewater respectively, the higher glycogen consumption with real wastewater could be due to a higher energy requirement for substrate hydrolysis and uptake. The high ratio of PHA stored per filtered COD removed in both acetate fed reactor and N.S. wastewater fed reactor indicate that more than 80% of the filtered COD were removed by the storage mechanism. In contrast, the low ratio of PHA stored per total COD removed observed for the two real wastewaters, points out that particulate COD was mainly removed by adsorption to the biofilm and not converted into PHA during the storage stage. The accumulated suspended solids served as an energy source for the following denitrification stage and also comprised a large fraction of the VSS in the real wastewater reactor, R2. The low ratios of PHA stored measured for Haifa wastewater, even when calculated based on filtered COD removed, imply a significant difference in characteristics between the Haifa and N.S. wastewaters. The Haifa wastewater, collected from the Haifa WWTP which serves the entire Haifa district, consisted of domestic as well as industrial wastewater, with a relatively long residence time (up to 24 h) in the sewerage system. On the other hand, N.S. wastewater was taken from a local wastewater collection point and had a significantly shorter residence time (2–4 h) in the sewerage system. Filtered COD, determined by 0.45 l filtration, usually contains colloidal particles and macromolecules that still have to undergo hydrolysis prior to being taken up and stored. It is reasonable to assume that the filtered COD fraction in Haifa wastewater contained a higher fraction of particles less susceptible to hydrolysis and storage in comparison to N.S. wastewater. Another explanation for the relatively low net PHA production per filtered COD removed with Haifa wastewater is a poorer selection towards a biomass capable of anaerobic storage. The higher effluent COD concentrations (both filtered and total) during period 2 in R2 in comparison to period 1 with N.S. wastewater, suggest that more external COD was present in the residual liquid volume during denitrification with Haifa wastewater than with N.S. waste-

water. Presence of external COD during denitrification would impair the growth of biomass capable of anaerobic storage. The measured PHA content within the biomass using VSS concentration as the measure for biomass concentration was 7.17% w/w at the end of the storage stage when grown on acetate based synthetic wastewater. For the biomass grown on real wastewaters the values were significantly lower: 3.84% w/w with N.S. wastewater and 2.1% w/w with Haifa wastewater (Table 3). The lower measured PHA content within the so called ‘‘biomass’’ in R2 can partially be explained by the large fraction of volatile solids originating from the wastewater rather than active biomass that obviously affected the calculated values. However, the much higher concentrations of PHA per liter reactor observed in the acetate reactor (>400 mg COD/L reactor) versus those observed in the wastewaters reactor (Fig. 4), indicate higher PHA storage with acetate based wastewater. In contrast, the glycogen concentrations per liter reactor were much higher in the wastewaters fed reactor (>400 mg/L reactor) in comparison to those observed in the acetate reactor (<200mg/ L reactor). The higher glycogen concentrations in biomass grown on wastewater together with the higher glycogen consumed/ PHA stored ratio (Table 3) may indicate higher GAO activity in the reactor fed with wastewater. However, additional research is required in order to verify this assumption. 3.2.3. PHA composition With acetate as the COD source for storage, PHB was the major PHA constituent, which is in a good agreement with previous studies (Table 4). With wastewater as the COD source, the PHV fraction was significantly higher than with acetate, and accounted for 50– 57% of the total PHA in biomass grown on real wastewater versus 19% in biomass grown on acetate based wastewater (Table 3). PH2MV was not detected in significant concentrations (<0.05% w/w) in both reactors during the research period. While acetate is easily taken up and stored as PHA with PHB comprising the major fraction, metabolism of complex organic substrate in wastewater may result in a different PHA composition. Apart from VFAs (mostly acetate), municipal wastewater contains a large fraction of soluble and particulate organic matter, the major

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E. Krasnits et al. / Bioresource Technology 132 (2013) 28–37

Table 4 Storage stoichiometry compared to reported values (model predictions and experimental results). Source

COD source, PAO/GAO

Units Pereira et al. (1996)* Zhou et al. (2008b) Yagci et al. (2003) Pijuan et al. (2008) Carvalho et al. (2007) Zeng et al. (2003b) R1 R2 period 1 R2 period 1 R2 period 2 R2 period 2 * **

PHB stored

PHV stored

PH2MV stored

PHA stored

Glycogen consumed

Net PHA stored**

– – 0.06 – – – – –

1.68 1.36 1.41 1.35 1.63 1.66 1.30 0.82 1.39 0.90 1.03

0.7 0.45 0.51 0.34 0.50 0.74 0.35 0.34 0.58 0.54 062

1.0 0.91 0.9 1.01 1.13 0.92 0.95 0.48 0.82 0.36 0.41

grCOD/grCOD removed Acetate, PAO Acetate, PAO Acetate, mixed Acetate, mixed Acetate, denit PAO Acetate, mixed Acetate WW (COD) WW (CODF) WW (COD) WW (CODF)

1.14 1.28 1.17 1.17 1.07 1.44 1.05 0.33 0.56 0.36 0.41

0.56 0.08 0.18 0.18 0.56 0.22 0.25 0.49 0.83 0.54 0.62

–

Results of model predictions. All other results are experimental. Net PHA stored = PHA stored minus glycogen consumed during storage.

constituents of which are carbohydrates, lipids and proteins. Prior to being stored as PHA, the organic matter has to undergo hydrolysis by exoenzymes. The resulting soluble low molecular weight compounds (e.g., sugars, disaccharides, amino acids, oligopeptides, glycerol, fatty acids) can then be taken up and metabolized via various metabolic pathways. Depending on the metabolic pathway used by different GAOs and PAOs, different ratio of acetyl-CoA and propionyl-CoA is obtained, resulting in a different PHA composition (Mino, 2000). Phosphate release during the storage stage was similar for all three cases: 7.1 mg PO4–P/L on acetate, 5.1 mg PO4–P/L with N.S. wastewater and 6.6 mg PO4–P/L with Haifa wastewater.

the operational conditions in this study in comparison to previous studies. While most of the previous studies were performed with operational conditions that selected for biomass capable of anaerobic storage (complete separation of electron donor and acceptor availabilities), the sludges in the present study were exposed to non perfect selection conditions. A previous study (Pijuan et al., 2008),which investigated the anaerobic storage of acetate and aerobic PHA degradation performance of sludges that were grown under non-perfect selection conditions in full scale WWTPs, reported similar COD removal rates to those found in the current study with acetate and real wastewater. 3.3. Denitrification stage

3.2.4. COD removal rates Most of the previously reported data regarding COD removal rates in anaerobic storage based systems are based on a linear average over varying periods of time while in fact the removal rates were not linear over the entire storage period. In order to compare the specific COD removal rates in this study to those found in previous studies, average specific COD removal rates during the first 30 min of the storage stage were calculated. This period of time was selected since in both of the reactors studied 80– 90% of the COD removal occurred during the first 30 min of the storage stage and the deviation from linearity during this time interval was found to be minimal. The results showed specific total COD removal rates during the storage stage of 32 mg COD per gram VSS per hour in the acetate fed reactor versus 40 mg COD per gram VSS per hour in the R2 reactor when fed with N.S. wastewater (period 1) and 50 mg COD per gram VSS per hour when fed with Haifa wastewater (period 2). These results correspond to daily removal rate of 0.14, 0.16 and 0.20 mg COD per gram VSS, respectively, based on several cycles per day. The higher specific total COD removal rate in R2 in comparison to that in R1, even though a significant fraction of the VSS in the real wastewaters reactor consisted of non-biomass solids, can be explained by the adsorption and entrapment of suspended solids. Batch tests performed with sludge grown on N.S. wastewater, taken from R2 during period 1 and with acetate as the substrate, yielded COD removal rates similar to those of sludge from R1 grown on acetate: 31 mg COD removed/grVSSh. These results imply that indeed the reason for the higher measured total COD removal rates in the reactor fed with real wastewater is suspended solids removal by trapping and adsorption. The results for both the acetate and the wastewaters reactor are in the lower range of the reported COD removal rates which vary between 32 and 130 mg COD/grVSSh (Carvalho et al., 2007; Jiang et al., 2010; Pijuan et al., 2008). These lower rates can probably be explained by

3.3.1. Concentrations profiles Since an integrated nitrification reactor was not studied in the present research project, both R1 and R2 were fed with nitrate and phosphorous solution at the beginning of the denitrification stage. During the denitrification stage nitrogen removal, phosphate uptake, PHA degradation and glycogen replenishment all occurred concurrently. Typical profiles of nitrate, nitrite, phosphate, PHA, and glycogen are presented in Fig. 5. During the denitrification stage the nitrate concentration decreased from 50 mg/L (NO3–N) to 19.5 mg/L in R1, while in R2 nitrate concentration decreased from 60 to 21 and from 60 to 7.4 with N.S. wastewater and Haifa wastewater, respectively. In R1, the nitrate removal efficiency was 61%, while in R2 the nitrate removal was 65% with N.S. wastewater and 87.7% with Haifa wastewater. A slight nitrite build up was observed in R2 during denitrification (periods 1 and 2), yet, at the end of the denitrification stage nitrite concentration decreased to almost zero (Table 5 and Fig. 5). Nitrogen removal (mg N/L reactorday), as shown in Table 5, was 44% higher with N.S. wastewater and 82% higher on Haifa wastewater than on acetate based synthetic wastewater. COD/N ratios, based on COD and N removal during a whole cycle, varied between 4.4 and 5.3 (Table 6). However, the actual values were slightly lower (around 4–4.2) due to the oxidation of stored COD by dissolved oxygen in the nitrate solution during the denitrification stage and also due to COD lost by suspended solids washout at the end of the denitrification stage. During denitrification, stored PHA served as electron donor for both anabolism and catabolism, as well as for replenishment of the glycogen reservoir. The results presented in Table 6 indicate that in R1, the PHA stored during the anaerobic stage from acetate based synthetic wastewater could account for the majority (94%) of the COD consumed for denitrification. In R2, where real wastewater was the source for COD storage, the ratio of CODPHA/NO3–N

35

E. Krasnits et al. / Bioresource Technology 132 (2013) 28–37

520

R1 acetate based synthetic wastewater

PHB, PHV (mgCOD/Lreactor)

60 50 40 30 20 10 0

0

0.5

1

1.5

160

440

140

400

120

360 320

100

280

80

240

60

200

40

160

20

120 80

2

0

0.5

1

t (hr)

2

220

R2 N.S. wastewater

480

R2 N.S. wastewater

200

PHB, PHV (mgCOD/Lreactor)

60 50 40 30 20 10

0

0.5

1

1.5

440

160

420

140 120

400

100

380

80

360

60

340

40

320

20 0

2

460

180

0

0.5

1

t (hr) 70

180

2

300

50 40 30 20 10 0 1

1.5

2

t (hr)

660

R2 Haifa wastewater

160

640

140

620

120

600

100

580

80

560

60

540

40

520

20

500

0

0

0.5

1

1.5

2

Gly (mgCOD/Lreactor)

PHB, PHV (mgCOD/Lreactor)

60

0.5

1.5

t (hr)

R2 Haifa wastewater

0

0

t (hr)

0

NO3-N, NO2-N, PO4-P (mg/L)

1.5

Gly (mgCOD/Lreactor)

NO3-N, NO2-N, PO4-P (mg/L)

70

180

R1 acetate based synthetic wastewater

480

Gly (mgCOD /Lreactor)

NO3-N, NO2-N, PO4-P (mg/L)

70

480

t (hr)

Fig. 5. Profiles of NO3–N (d), NO2–N (s), PO4–P (N), PHB (e), PHV (h) and Glycogen (x), during a typical denitrification stage in R1 on acetate based wastewater (day 55), in R2 with N.S. wastewater (day 27) and in R2 with Haifa wastewater (day 48).

removed, was lower than the theoretical ratio required for catabolism only. In R2 the COD from stored PHA accounted for about 60% of the COD required for denitrification when fed with N.S. wastewater, and only about 40% of the COD required for denitrification

when fed with Haifa wastewater. These results support the assumption that COD removed by way of adsorption and entrapment of solids during the storage stage was later on used as additional reducing power for the denitrification.

Table 5 Denitrification stage influent and effluent characteristics – average values. Denitrification stage influent NO3–N

PO4–P

Denitrification stage effluent pH

mg/L R1 R2 period 1 R2 Period 2

50 (1.7) 60 (1.0) 60 (1.2)

NO3–N

NO2–N

Daily removal of N and P PO4–P

pH

mg/l 9.8 (1.4) 9.2 (1.4) 18.5 (0.5)

7.5 (0.1) 7.5 (0.1) 7.5 (0.1)

19.5 (4.6) 21 (4.8) 7.4 (5.2)

PO4–P

NO3–N

mg/L reactorday 0 0 0.3 (0.1)

0.6 (0.5) 1.4 (1.4) 7.9 (1.3)

7.9 (0.5) 8.0 (0.2) 8.2 (0.2)

50.8 48.4 61.8

168 242 306

36

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Table 6 PHA and glycogen contents at the end of denitrification stage and PHA 1st order degradation rate constant. Units

PHA rate constant in biofilm reactors (In brackets batch tests – suspended) K (1/hr)*

PHA/VSS

R1 R2 period 1 R2 period 2

0.10 (0.21) 0.26 (0.81) 0.25 (0.72)

5.87 2.28 0.91

Glycogen/VSS

COD removed/N removed**

PHA consumed/N removed

% w/w g

COD/g

NO3–N

3.10 6.76 8.18

4.6 5.3 4.4

3.9 2.6 1.5

K is the 1st order constant from df PHA ¼ KfPHA while fPHA is the fraction (w/w) of PHA in the VSS. dt Based on COD and N removal per cycle. Actual COD/N values were slightly lower (around 4–4.2), due to the presence of dissolved oxygen and due to COD suspended solids washout. *

**

3.3.2. PHA degradation rates with biomass fed with acetate based synthetic wastewater under PBD conditions The kinetics of PHA degradation was studied in the R1 biofilm reactor and in batch tests under PBD conditions with biomass taken from R1 The results from the batch tests (Table 6) showed a first order PHA degradation rate with a reaction constant of K = 0.21 h1. These results are similar to those of Beun et al. (2000), where a first order reaction with a constant rate of 0.2 h1 was observed for PHB degradation. In the R1 biofilm reactor, the observed PHA degradation rate was much lower than in the suspended sludge batch tests, with a first order reaction rate constant of K = 0.1 h1 (based on initial and final PHA concentration in the denitrification stage). The difference between the PHA degradation rates in the biofilm reactor and the suspended sludge batch test can be attributed to diffusion limitations that resulted in lower PHA degradation rate in R1. The diffusion limitations in the biofilm reactors were exacerbated by poor mixing conditions caused by local biomass clogging of the plastic biofilm carriers used in the reactors, versus the completely mixed conditions in the batch tests. 3.3.3. PHA degradation rates with biomass fed with real wastewaters under PBD conditions Similarly to the above experiments on sludge grown on acetate based synthetic wastewater (R1), the PHA degradation kinetics of sludge grown on real wastewater under PBD conditions was studied in R2 itself and in suspended sludge batch tests with sludge taken from R2 during periods 1 and 2. The batch tests results (Table 6) showed significantly higher measured specific PHA degradation rates than those observed in the batch tests with the acetate fed biomass: 1st order PHA degradation rate with a reaction constant of K = 0.81 h1 for sludge fed with N.S. wastewater and K = 0.72 h1 for sludge fed with Haifa wastewater. The PHA degradation rate constants obtained from R2 (based on initial and final PHA concentration during the denitrification stage) were also higher than that of the R1 reactor: K = 0.26 h1 for N.S. wastewater and K = 0.25 h1 for Haifa wastewater. Comparison between the rates constants obtained directly from R2 and from batch tests with suspended sludge taken from R2, indicates that similarly to R1, diffusion limitations in R2 significantly reduced the PHA degradation rates.

Except for the size of PHA granules which influences the surface area available for enzymatic attack (Beun et al., 2000), there are many known factors influencing microbial PHA degradation rates, such as chemical composition of the polymer, level of crystallinity, dimensions of spherulites and enzyme type and activity. The higher level of heterogeneity of the PHA copolymer found in this research for the sludge grown on real wastewater (higher PHV/PHA ratio) could have a strong positive influence on PHA degradation rate (Volova, 2004). However, understanding the exact reasons for the great difference in PHA degradation rates between R1 and R2 requires additional research. 3.4. N2O emission during PBD N2O is a potent greenhouse gas. Previous studies have found that during denitrification with PHA as the electron donor, a significant fraction of the NO3–N is partially denitrified to N2O. One of the factors contributing to the emission of N2O is considered to be the relatively low PHA degradation rate, which causes competition for electrons between denitrifying enzymes. Such competition creates an advantage for NO reducing enzymes over N2O reducing enzymes (Kampschreur et al., 2009). Reported studies on PBD with synthetic wastewater have shown a wide range of N2O emission rate values, depending on operating conditions such as anaerobic reaction time and NO2 and free nitrous acid (FNA) concentrations (Wang et al., 2011). During the current research, batch tests were carried out in order to measure the fraction of denitrified NO3–N resulting in N2O–N. The results obtained from the batch experiments (Fig. 6) showed that during PBD on acetate, the fraction of denitrified NO3–N reduced to N2O–N was 14.8%, while during PBD on real

700 600 500

ugrN2O-N/L

During denitrification, phosphorus removal was observed in both reactors. The substantial phosphorus removal observed indicates that denitrifying PAOs were present and active during the denitrification stage in both R1 and R2: 9.2 mg/L PO4–P in R1 and 7.8 and 10.6 mg PO4–P/L in R2 during periods 1 and 2 respectively (Table 5). Since no controlled sludge removal was practiced during the current research, the phosphorus removal was probably possible due to natural washout of phosphorus enriched sludge (PAOs).

400 300 200 100 0

0

30

60

90

120

150

180

t (min) Fig. 6. N2O–N concentrations in the gaseous phase during PBD on N.S. wastewater in R2 (d) and acetate based synthetic wastewater in R1 (D), as measured in the ATR-LP-FTIR chamber.

E. Krasnits et al. / Bioresource Technology 132 (2013) 28–37

wastewater the fraction was 13.7%. These results are in agreement with previous findings on PBD in full scale wastewater treatment plants (Wang et al., 2011). 4. Conclusions PBD with municipal wastewater using biofilm reactors was capable of removing high nitrate concentration (39–52 mg N/L) with a low COD requirement (COD/N ratios of 4.6–5.3). It was found that entrapped particulate organic matter in the denitrification stage more than doubled the denitrification capacity in comparison to denitrification based on PHA only. However, the COD removal rates were found to be in the lower range of reported data for PBD on acetate (0.24–0.31 gr N/L reactord), due to unavoidable non – perfect selection conditions towards storage capable bacteria when using real wastewater. References Ahn, Y.H., 2006. Sustainable nitrogen elimination biotechnologies: a review. Process Biochemistry 41, 1709–1721. APHA, 1995. Standard Methods for the Examination of Water and Wastewater, nineteenth ed. APHA, AWWA, WPCF, Washington, DC. Bernat, K., Wojnowska-Baryła, I., Dobrzyn´ska, A., 2008. Denitrification with endogenous carbon source at low C/N and its effect on P(3HB) accumulation. Bioresource Technology 99, 2410–2418. Beun, J.J., Verhoef, E.V., Van Loosdrecht, M.C.M., Heijnen, J.J., 2000. Stoichiometry and kinetics of poly-b-hydroxybutyrate metabolism under denitrifying conditions in activated sludge cultures. Biotechnology and Bioengineering 68 (5), 496–507. Bortone, G., Malaspina, F., Stante, L., Tilche, A., 1994. Biological nitrogen and phosphorus removal in an anaerobic/anoxic sequencing batch reactor with separated biofilm nitrification. Water Science & Technology 30 (6), 303–313. Carucci, A., Dionusi, D., Majone, M., Rolle, E., Smurra, P., 2001. Aerobic storage by activated sludge on real wastewater. Water Research 35 (16), 3833–3844. Carvalho, G., Lemos, P.C., Oehmen, A., Reis, M.A.M., 2007. Denitrifying phosphorus removal: linking the process performance with the microbial community structure. Water Research 41, 4383–4396. Ciggin, A.S., Karahan, O., Orhon, D., 2007. Effect of biochemical storage on the denitrification potential of acetate in sequencing batch reactors. Water Science & Technology 55 (10), 261–268. Gujer, W., Henze, M., Mino, T., Van Loosdrecht, M.C.M., 1999. Activated sludge model No. 3. Water Science & Technology 39 (1), 183–193. Hughes, L.J., Lancaster, L., Cord-Ruwisch, R., 2006. Multistage wastewater treatment using separated storage driven denitrification and nitrification biofilms. Water Science & Technology 53 (6), 51–58. Jiang, X., Ma, F., Zhang, X., Yin, J., Wang, L. 2010. Characteristics of denitrifying phosphorous removal in anaerobic/anoxic SBR system, National Natural Science Foundation of China (NSFC) (No. 50778052). Kampschreur, M.J., Temmink, H., Kleerebezem, R., Jetten, M.S.M., Van Loosdrecht, M.C.M., 2009. Nitrous oxide emission during wastewater treatment. Water Research 43, 4093–4103. Karahan, O., Orhon, D., Van Loosdrecht, M.C.M., 2008. Simultaneous storage and utilization of polyhydroxyalkanoates and glycogen under aerobic conditions. Water Science & Technology 58 (4), 945–951. Kuba, T., Murnleitner, E., Van Loosdrecht, M.C.M., Heijnen, J.J., 1996a. A metabolic model for biological phosphorus removal by denitrifying organisms. Biotechnology and Bioengineering 52, 685–695. Kuba, T., Van Loosdrecht, M.C.M., Heijnen, J.J., 1996b. Phosphorus and nitrogen removal with minimal COD requirement by integration of denitrifying

37

dephosphatation and nitrification in a two-sludge system. Water Research 30 (7), 1702–1710. Mino, T., 2000. Microbial selection of polyphosphate-accumulating bacteria in activated sludge wastewater treatment processes for enhanced biological phosphate removal. Biochemistry (Moscow) 65 (3), 341–348. Oehmen, A., Keller-Lehmann, B.J., Zeng, R., Yuan, Z., Keller, J., 2005. Optimization of poly-b-hydroxyalkanoate analysis using gas chromatography for enhanced biological phosphorus removal systems. Journal of Chromatography A 1070, 131–136. Paredes, D., Kusch, P., Mbwette, T.S.A., Stange, F., Müller, R.A., Köser, H., 2007. New aspects of microbial nitrogen transformations in the context of wastewater treatment – a review. England Life Science 7 (1), 13–25. Pereira, H., Lemos, P.C., Reis, M.A.M., Crespo, J.P.S.G., Carrondo, M.J.T., Santos, H., 1996. Model for carbon metabolism in biological phosphorus removal processes based on in vivo NMR labelling experiments. Water Research 30 (9), 2128– 2138. Pijuan, M., Oehmen, A., Baeza, A., Casas, C., Yuan, Z., 2008. Characterizing the biochemical activity of full-scale enhanced biological phosphorus removal systems: a comparison with metabolic models. Biotechnology and Bioengineering 99 (1), 170–179. Qin, L., Liu, Y., Joo-Hwa, T., 2005. Denitrification on poly-b-hydroxybutyrate in microbial granular sludge sequencing batch reactor. Water Research 39, 1503– 1510. Satoh, H., Mino, T., Matsuo, T., 1992. Uptake of organic substrates and accumulation of polyhydroxyalkanoates linked with glycolysis of intracellular carbohydrates under anaerobic conditions in the biological excess phosphate removal processes. Water Science and Technology 26 (5–6), 942–993. Schuler, A.J., Jenkins, D., 2003. Enhanced biological phosphorus removal from wastewater by biomass with different phosphorus content. Part II. Anaerobic adenosine triphosphate utilization and acetate uptake rates. Water Environment Research 75 (6), 499–511. Segal-Rosenheimer, M., Dubowski, Y., 2007. Heterogeneous ozonolysis of cypermethrin using real-time monitoring FTIR techniques. Journal of Physical Chemistry C 111, 11682–11691. Sorm, R., Bortone, G., Saltarelly, R., Jenicek, P., Wanner, J., Tilche, A., 1996. Phosphate uptake under anoxic conditions and fixed film nitrification in nutrient removal activated sludge system. Water Research 30 (7), 1573–1584. Van Loosdrecht, M.C.M., Pot, M.A., Heijnen, J.J., 1997. Importance of bacterial storage polymers in bioprocess. Water Science and Technology 35, 41–47. Volova, T.G., 2004. Polyhydroxyalkenoates – Plastic Materials of the 21st Century: Production, Properties, Application. Nova Science Publishers Inc., New York, 160–161. Wang, Y., Geng, J., Ren, Z., He, W., Xing, M., Wu, M., Chen, S., 2011. Effect of anaerobic reaction time on denitrifying phosphorus removal and N2O production. Bioresource Technology 102, 5674–5684. Yagci, N., Artan, N., Cokgor, E.U., Randall, C.W., Orhon, D., 2003. Metabolic model for acetate uptake by a mixed culture of phosphate- and glycogen-accumulating organisms under anaerobic conditions. Biotechnology and Bioengineering 84 (3), 359–373. Zafriadis, I., Kapagiannidis, A.G., Aviasidis, A., 2009. Monitoring of microbial storage products and the efficiency of an activated sludge plant performing anoxic phosphorus removal under different operational conditions. In: Proceedings of the Eleventh International Conference on Environmental Science and Technology. Chaina Crete, Greece, pp. A1578–1575. Zeng, R.J., Saunders, A.M., Yuan, Z., Blackall, L.L., Keller, J., 2003a. Identification and comparison of aerobic and denitrifying polyphosphate-accumulating organisms. Biotechnology and Bioengineering 83 (2), 140–148. Zeng, R.J., Yuan, Z., Keller, J., 2003b. Model-based analysis of anaerobic acetate uptake by a mixed culture of polyphosphate-accumulating and glycogenaccumulating organisms. Biotechnology and Bioengineering 83 (3), 293–302. Zhou, Y., Pijuan, M., Yuan, Z., 2008a. Development of a 2-sludge, 3-stage system for nitrogen and phosphorous removal from nutrient-rich wastewater using granular sludge and biofilms. Water Research 42, 3207–3217. Zhou, Y., Pijuan, M., Zeng, R.J., Lu, H., Yuan, Z., 2008b. Could polyphosphateaccumulating organisms (PAOs) be glycogen-accumulating organisms (GAOs)? Water Research 42, 2361–2368.