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Comparative evaluation of anaerobic digestion for sewage sludge and various organic wastes with simple modeling
Waste Management xxx (2015) xxx–xxx
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Comparative evaluation of anaerobic digestion for sewage sludge and various organic wastes with simple modeling Taira Hidaka ⇑, Feng Wang, Jun Tsumori Recycling Research Team, Materials and Resources Research Group, Public Works Research Institute, 1-6, Minamihara, Tsukuba, Ibaraki 305 8516, Japan
a r t i c l e
i n f o
Article history: Received 20 August 2014 Accepted 20 April 2015 Available online xxxx Keywords: Anaerobic digestion Sewage sludge Organic waste Digested sludge Mathematical model
a b s t r a c t Anaerobic co-digestion of sewage sludge and other organic wastes, such as kitchen garbage, food waste, and agricultural waste, at a wastewater treatment plant (WWTP) is a promising method for both energy and material recovery. Substrate characteristics and the anaerobic digestion performance of sewage sludge and various organic wastes were compared using experiments and modeling. Co-digestion improved the value of digested sewage sludge as a fertilizer. The relationship between total and soluble elemental concentrations was correlated with the periodic table: most Na and K (alkali metals) were soluble, and around 20–40% of Mg and around 10–20% of Ca (alkaline earth metals) were soluble. The ratio of biodegradable chemical oxygen demand of organic wastes was 65–90%. The methane conversion ratio and methane production rate under mesophilic conditions were evaluated using a simplified mathematical model. There was reasonably close agreement between the model simulations and the experimental results in terms of methane production and nitrogen concentration. These results provide valuable information and indicate that the model can be used as a pre-evaluation tool to facilitate the introduction of co-digestion at WWTPs. Ó 2015 Elsevier Ltd. All rights reserved.
1. Introduction Anaerobic digestion of organic wastes with high moisture content, such as sewage sludge, kitchen garbage, food waste, and agricultural waste, is a promising method for both energy and material recovery. A wastewater treatment plant (WWTP) that can accept many kinds of organic waste for co-digestion could become a regional energy hub if the anaerobic digester produces more methane than the digestion of sewage sludge alone (Cabbai et al., 2013; Navaneethan et al., 2011; Di Maria et al., 2014). In 2011, the Sewerage and Wastewater Management Department of the Ministry of Land, Infrastructure, Transport and Tourism, Japan (2013) initiated the nationwide project called Breakthrough by Dynamic Approach in Sewage High Technology (B-DASH Project) for the development of innovative technologies in sewage treatment. Demonstration experiments were performed, including the co-digestion of sewage sludge and regional organic waste (NILIM, 2013a,b). However, it is likely that the characteristics of sewage sludge and regional organic waste are different in each WWTP. The properties of substrates such as total solids (TS), volatile solids (VS), chemical oxygen demand (COD), proteins, lipids, and nutrient ⇑ Corresponding author. Tel.: +81 29 879 6765; fax: +81 29 879 6797. E-mail address: [email protected] (T. Hidaka).
and trace elements can provide basic information for understanding anaerobic co-digestion characteristics. Co-digestion could potentially cause reaction inhibition by such as ammonia, metals and organics (Chen et al., 2008), increase loading to the water treatment process, and worsen the quality of effluent water from WWTPs. Some trace elements are required for methanogenesis, and some elements are important components of fertilizer. Basic information and a simple pre-evaluation tool for co-digestion are required to promote these technologies. Batch experiments are widely used to investigate methane conversion performance. Alzate et al. (2012) used batch experiments to compare anaerobic digestion of microalgae with substrate-toinoculum (S/I) ratios of 0.5, 1, and 3 (gVS/gVS), and reported that the methane production was higher at lower S/I ratios. The S/I ratios used in batch experiments are often higher than 0.5. Furthermore, the biodegradation profile in batch experiments affects the methane production quantity in continuous operation, especially in the case of a shorter hydraulic retention time (HRT). Assuming a first-order reaction and a continuously stirred tank reactor (Astals et al., 2013), theoretically not all of the degradable substrate is degraded, whereas almost all of the substrate is ultimately degraded in batch experiments. The sum of the methane conversion ratio multiplied by the feeding quantity is commonly used to design co-digestion systems, but this summation cannot accurately estimate the total methane production quantity in continuous operations.
http://dx.doi.org/10.1016/j.wasman.2015.04.026 0956-053X/Ó 2015 Elsevier Ltd. All rights reserved.
Please cite this article in press as: Hidaka, T., et al. Comparative evaluation of anaerobic digestion for sewage sludge and various organic wastes with simple modeling. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.04.026
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Mathematical models have been used to evaluate methane fermentation performance, combining operational parameters such as HRT, and substrate characteristics that differ among WWTPs. The International Water Association anaerobic digestion model No. 1 (ADM1) (Batstone et al., 2002) is an example of an effective general model that can be applied to many kinds of organic substrates. However, characterization of the biodegradability of substrates is required before using the ADM1. Yasui et al. (2008) and Astals et al. (2013) evaluated the sewage sludge from different WWTPs using anaerobic respirometry and mathematical modeling, but did not include other types of organic waste. Mottet et al. (2010) analyzed the thermophilic anaerobic biodegradability of waste activated sludge (WAS) from different WWTPs. Girault et al. (2012) proposed a COD fraction characterization method for the ADM1, but only two types of organic waste (pig slurry and WAS) were evaluated. Souza et al. (2013) proposed a simplified COD fractioning criterion for the ADM1, but it applied only to raw and pre-treated WAS. The ADM1 could be too complicated for use by designers and operational managers on site. Donoso-Bravo et al. (2009) proposed a two-population, three-reaction model, which was used to evaluate the thermophilic anaerobic digestion of sewage sludge (Donoso-Bravo et al., 2012). However, this simplified model did not include the self-degradation of microorganisms, which is important in batch experiments. Although extensive research on co-digestion using experiments and modeling has been reported (Derbal et al., 2009; Esposito et al., 2011), there have been few comparative evaluations using a wide range of organic wastes together with sewage sludge. In the present study, the substrate characteristics and anaerobic digestion performance of various organic wastes were compared using experiments and simplified mathematical modeling. The primary purpose was to provide municipal employees and designers on site with basic information and a simple pre-evaluation method to facilitate the introduction of co-digestion at WWTPs. 2. Materials and methods 2.1. Anaerobic digestion characteristics of various organic wastes Various organic wastes were compared in the present study. Glucose (GL) was used as a substrate to investigate the activity of the digested sludge. Mixed sewage sludge (MS), which is a mixture of primary sludge and WAS, was obtained from WWTP-A. Separated primary sludge (PS) from WWTP-B was obtained using an effective separation system demonstrated by the B-DASH Project (NILIM, 2013a), instead of a traditional primary
sedimentation system. Kitchen garbage (KG), which was a mixture of PS and raw KG, was obtained from the B-DASH Project (NILIM, 2013a). Cow manure (CM) and swine manure (SM) were obtained from an agricultural research institute. Cabbage (CB), potato (PT), and carrot (CR) were evaluated as representatives of waste leaf and root vegetables. All parts of the vegetables, including edible parts and peels, were crushed together and used for the experiments. As part of the B-DASH Project, imperfect vegetables from markets were planned to be co-digested with sewage sludge. Digested sludge adapted to various organic wastes was cultivated as inoculum sludge for the batch experiments. Two semi-continuous complete mix reactors, Reactor 1 and Reactor 2, each with a working volume of 100 L, were operated under mesophilic (35 °C) and thermophilic (55 °C) conditions, respectively. When the present study started, Reactor 1 and Reactor 2 had been operated for three months with MS. At the beginning of the present study, 3 L of the mesophilic digested sludge from the B-DASH Project demonstration anaerobic digester (NILIM, 2013b) was added to each reactor as an inoculum sludge adapted to MS and other waste biomass. A mixture of MS, PS, CM, and SM with average VS of around 37 g/L was fed to Reactor 1 and Reactor 2. The temperature was controlled using a hot water bath. The reactors were fed three times a week (on Monday, Wednesday, and Friday). The HRT was set at 120 d, and it was gradually decreased to 47 d, which corresponds to an organic loading rate (OLR) of around 0.8 kg VS/(m3 d). The main purpose was to cultivate inoculum sludge adapted to various organic wastes for the following batch experiments. Therefore, the HRT was set to be longer than the HRT of conventional anaerobic digesters (JSWA, 2003). Four series of batch experiments (B1, B2, B3, and B4) were performed using a commercially available experimental kit with 600 mL-vessels stirred at around 100 rpm (AMPTS II; Bioprocess Control AB, Sweden) (Table 1). Four hundred mL of the digested sludge and substrate were added to each vial. No microelements were added. The S/I ratios were set lower than for typical batch experiments, to determine rate constants under similar conditions to continuous operation. Before sealing, the vials were flushed with nitrogen gas. The produced biogas was passed through an alkali liquid (3 M NaOH) to remove CO2, and the quantity of produced methane gas was monitored using the gas meter (AMPTS II). Before starting each batch experiment, substrate feeding was suspended for about one week in Reactor 1 and Reactor 2 to reduce the background methane production. The digested sludge cultivated in Reactor 1 and Reactor 2 was used for B1 and B3 (mesophilic), and B2 and B4 (thermophilic), respectively. The substrate addition ratio was designed using the
Table 1 Batch experimental conditions.
a
Mesophilic
Thermophilic
B1
B2
Substrate
B1-1
B2-1
Blank
B1-2
B2-2
GL
B1-3
B2-3
GL
B1-4
B2-4
MS
B1-5
B2-5
B1-6
S/I a
Substrate
S/I a
Mesophilic
Thermophilic
B3
B4
−
B3-1
B4-1
Blank
−
0.1
B3-2
B4-2
GL
0.1
0.2
B3-3
B4-3
GL
0.2
0.1
B3-4
B4-4
PS
0.1
MS
0.2
B3-5
B4-5
PS
0.2
B2-6
KG
0.1
B3-6
B4-6
CM
0.1
B1-7
B2-7
KG
0.2
B3-7
B4-7
CM
0.2
B1-8
B2-8
KG
0.4
B3-8
B4-8
CM
0.4
B3-9
B4-9
SM
0.1
B3-10
B4-10
SM
0.2
B3-11
B4-11
SM
0.4
B3-12
B4-12
CB
0.2
B3-13
B4-13
PT
0.2
B3-14
B4-14
CR
0.2
substrate/inoculum (digested sludge) ratio on VS base
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Moderately biodegradable particulate COD (Xm, NXm)
Readily biodegradable particulate COD (Xr, NXr) Solubilization R1r
R1m
R1s
Slowly biodegradable particulate COD (Xs, NXs)
Inert particulate COD (Xi, NXi)
Biodegradable soluble COD (Cb, NCb)
(Y4) (1᧩Y4)
Inert soluble COD (Ci, NCi)
Acidogenesis R2 Acid fermenter (Xsu, NXsu)
(Y2) (1᧩Y2)
Self-degradation R4Xsu
Ammonia (Nam)
Acetate and hydrogen (Cac) Methanogenesis R3
Methanogens (Xac, NXac)
(Y3) (1᧩Y3)
Self-degradation R4Xac
Methane State variables: X, particulate COD (gCOD/L) C, soluble COD (gCOD/L) N, Nitrogen (gN/L)
Reaction rates: R1i᧹k1i×Xi (i=r, m, s) R2᧹k2×Cb/(K2+Cb)×Xsu R3᧹k3× Cac/(Kac+Cac)× Xac R4Xi᧹k4i× Xi (i=su, ac)
Fig. 1. Schematic diagram of the developed model.
VS ratio of substrate to inoculum digested sludge. Vegetable samples were crushed using an electric blender (ABS-V; Osaka Chemical Co., Ltd., Japan). Each experiment included blank experiments using only digested sludge. Additional blank experiments (B3-X1, -X2 and B4-X1, -X2) were performed using 500-mL conical flasks, and the quantity of produced biogas was measured to ensure experimental accuracy. 2.2. Mathematical model Fig. 1 shows a schematic diagram of the developed model. This model is based on a similar concept to that of the ADM1, but is simplified. In the ADM1, COD fractions are characterized by carbohydrates, proteins, and lipids. However, because accurate measurement of these items is time-consuming and costly, estimated ratios are often used for the ADM1 (Souza et al., 2013). In the present study, the substrate COD fraction was determined in terms of degradation speed rather than components. The substrate for methanogenesis is acetate and hydrogen, but in this case, acetate and hydrogen were considered as one state variable. Organic waste is divided into the following categories: readily biodegradable particulate COD (Xr), mildly biodegradable particulate COD (Xm), slowly biodegradable particulate COD (Xs), inert particulate COD (Xi), biodegradable soluble COD (Cb), and inert soluble COD (Ci). Particulate COD components are solubilized into Cb (R1, solubilization), Cb is converted into acetate and hydrogen (Cac) (R2, acidogenesis), and Cac is converted into methane (R3, methanogenesis). There are two types of microorganisms: acid fermenters (Xsu), which produce acetate and hydrogen from soluble COD such as sugar, and methanogens (Xac), which produce methane from acetate and hydrogen. Xsu and Xac are degraded into Xs and Xi (R4, self-degradation). Each COD fraction contains nitrogen (NXr, NXm, NXs, NXi, NCb, NCi, NXsu, and Nac), and ammonia (Nam) is produced corresponding to the degradation of
soluble COD into acetate and hydrogen. The mass balance of each state variable was calculated considering yield coefficients (Y2–Y4). Each reaction rate was expressed as the multiplication of a rate constant, the Michaelis–Menten kinetic, and a microorganism concentration. The effect of ammonia on microbial activity was expressed with a half saturation inhibition constant. In the present mesophilic experiments, pH was 7.3–7.7, and the total ammonia concentration was <1.1 g N/L, which would not lead to inhibition from ammonia accumulation (Rajagopal et al., 2013). The time course was calculated using the Runge–Kutta–Gill method. The program was developed using Visual Basic for Applications (VBA) in Microsoft Excel 2013, which could be used by municipal employees or designers on site. 2.3. Chemical analysis The pH, electrical conductivity (EC), TS, VS, and alkalinity were analyzed using standard methods (APHA, 1995). Total COD (TCOD) and soluble COD (SCOD) were analyzed using a spectrophotometer (DR3900; Hach Company, USA) and COD Digestion Vials (High Range, Hach Company). Particle COD (PCOD) was calculated as TCOD SCOD. Ammonia and total nitrogen were assayed using an auto analyzer (TRAACS2000; Bran Luebbe, Germany). Dissolved samples were prepared by filtration through a cellulose acetate filter (pore size = 0.45 lm). The carbon hydrogen nitrogen oxygen (CHNO) ratio was measured using an elemental analyzer (FLASH EA1112; Amco Inc., Japan). The elements Al, Ca, Fe, K, Mg, Na, and P, and B, Cr, Mn, Co, Ni, Cu, Zn, As, Se, Cd, Sb, and Pb were measured using inductively coupled plasma (ICP) atomic emission spectrometry (ICPS-8000E; Shimadzu, Japan), and ICP mass spectrometry (ICP-MS X series; Thermo Fisher, USA), respectively. Protein was measured using the Lowry method (Lowry et al., 1951), and lipid was measured by direct-heat extraction (SF-6; Sanshin Industrial Co. Ltd., Japan).
Please cite this article in press as: Hidaka, T., et al. Comparative evaluation of anaerobic digestion for sewage sludge and various organic wastes with simple modeling. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.04.026
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T. Hidaka et al. / Waste Management xxx (2015) xxx–xxx
and Mg compared with MS, and SM has higher containing ratios of Ca, K, Mg, Na, P, Co, Cu, and Zn. These elements are valuable in terms of the production of fertilizer. Co-digestion can therefore improve the value of digested sewage sludge as a fertilizer.
3. Results and discussion 3.1. Comparison of substrate characteristics The basic characteristics of MS and organic wastes are summarized in Table 2. MS was measured twice for batch experiments B1 and B2. The TS, VS, and TCOD were much higher in organic wastes than in MS. The C- and H-containing ratios were similar, except for KG. The C/N ratio of organic wastes was similar or higher than that of MS, which was round 9 ( ). Co-digestion can be used to increase the C/N ratio, by feeding with various organic wastes. The average ratio of g protein/g nitrogen measured using the elemental analyzer was around 5.7. The KG and PT had more lipids than other organic wastes. Elemental concentrations of MS and organic wastes are summarized in Table 3. The containing ratio of the organic wastes was generally lower than that of MS. Among these elements, Fe, Ni, Co, and Zn are typical nutrients required for methanogenesis (Takashima et al., 2011). While some organic wastes had relatively higher containing ratios of these elements compared with MS, most organic wastes had lower containing ratios. When organic wastes that do not contain sufficient nutrients are digested, the addition of these elements is recommended. However, receiving such organic wastes at WWTPs generally results in reduced addition of nutrients, because sewage sludge contains these elements. This is another potential advantage of co-digestion utilizing anaerobic digesters at WWTPs. CM has higher containing ratios of Ca, K,
3.2. Effect of co-digestion on quality of the digested sludge The relationship between the substrate C/N ratio of organic waste and increased ammonia per produced methane is shown in Fig. 2. Organic wastes with a C/N ratio <50 are shown. The quantity of methane produced was used as an index of degraded COD. Increased ammonia contributed by added organic wastes is estimated by extracting produced ammonia from the inoculum sludge. The MS had a C/N ratio of around 9, and organic wastes with higher C/N ratios produced less ammonia per produced methane, although there was some variation resulting from the different degradation characteristics of COD and nitrogen components. An ammonia concentration of around 3 g N/L is acceptable for mesophilic anaerobic digestion of MS, although it is not acceptable for thermophilic operation (Hidaka et al., 2013). Ammonia inhibition might not be a problem in anaerobic digestion of sewage sludge only, but it should be considered when organic wastes containing high N components are co-digested. The EC in the mesophilic and thermophilic batch experiments was 7.3–9.5 and 7.2–8.7 mS/cm, respectively. Alkalinity in the mesophilic and thermophilic batch experiments was 3.6–6.2 and 4.5–5.9 gCaCO3/L, respectively. The EC and alkalinity were markedly affected by an increase in ammonia.
Table 2 Substrate characteristics.
MSa MSb PS KG CM SM CB PT CR a b c
TS (g/kg)
VS (g/kg)
TCOD (g/kg)
SCOD (g/kg)
C (%-TS)
H (%-TS)
N (%-TS)
O (%-TS)
C/N (-)
Protein (%-TS)
23.1 17.3 11.9 131 161 241 95.3 181 97.6
19.3 14.1 10.5 111 144 194 87.5 171 89.7
28.5 23.0 17.8 166 155 268 116 179 114
1.1 1.8 2.4 48.1
41.1 41.0 44.1 53.4 44.2 40.2 39.9 41.9 40.0
6.1 6.2 6.7 7.8 6.3 6.2 6.2 6.6 6.3
4.3 4.8 2.3 3.4 1.9 4.1 2.0 0.9 0.2
33.7 30.5 32.4 31.9 33.0 29.5 41.5 42.3 42.5
9.6 8.6 19.5 15.6 23.8 9.9 20.0 45.1 194
24.1
2.8
17.3 12.7 16.5 24.4 7.0 8.7 3.9
4.5 21.3 2.3 3.0 4.2 9.7 0.8
c c
76.4 58.3 79.2
Lipid (%-TS)
Measured for B1. Measured for B2. Cannot be measured due to high TS.
Table 3 Elemental concentrations of substrate per dry solids (DS), mg E/kg DS.
Al Ca Fe K Mg Na P B Cr Mn Co Ni Cu Zn As Se Cd Sb Pb a
MS
PS
FW
FT
KG
CM
SM
CB
PT
CR
7.29 103 1.01 104 1.80 104 3.43 103 2.72 103 2.99 103 1.40 104 1.75 10 6.20 5.90 102 1.98 5.82 1.04 102 1.80 102 4.42 2.51 3.52 10 1 7.20 10 1 1.19 10
4.24 103 8.77 103 6.34 103 3.25 103 1.32 103 4.36 103 7.33 103 1.11 10 7.18 10 1.09 102 2.24 6.09 10 1.66 102 6.56 102 1.36 5.64 10 1 6.91 10 1 1.24 1.41 10
2.67 104 1.38 103 7.17 102 ND 2.05 102 5.63 102 5.02 103 1.80 4.69 1.06 10 ND 1.46 6.89 1.39 10 ND ND ND ND 4.15 10 1
5.46 10 1.97 103 1.09 102 1.09 103 2.73 102 ND ND 4.54 3.97 1.78 102 5.19 10 1 1.97 1.76 1.49 10 ND ND 2.35 10 1 ND 1.81
4.80 102 2.51 103 8.00 102 3.52 103 6.40 102 7.41 103 2.45 103 4.06 1.02 10 1.67 10 3.15 10 1 8.05 2.35 10 6.28 10 5.41 10 1 ND ND 1.65 10 1 3.35
3.24 102 1.53 104 4.86 102 7.57 103 6.11 103 1.73 103 5.95 103 1.80 10 1.64 1.31 102 3.60 10 1 2.20 1.93 10 6.79 10 2.11 10 1 ND 1.65 10 1 ND 2.80 10 1
3.57 102 3.18 104 2.34 103 1.31 104 9.07 103 4.89 103 1.97 104 1.91 10 4.25 3.52 102 2.68 5.22 1.90 102 3.66 102 7.56 10 1 9.47 10 1 3.23 10 1 ND 5.94
NDa 5.25 103 ND 2.51 104 1.52 103 ND 3.19 103 1.69 10 ND 1.03 10 ND 1.55 10 1 2.03 1.12 10 ND ND ND 4.27 10 1 ND
1.53 102 1.53 102 1.02 102 1.81 104 9.20 102 ND 1.48 103 5.88 ND 6.91 ND 2.67 10 1 2.67 8.59 ND ND ND ND ND
ND 2.33 103 ND 2.57 104 6.90 102 8.49 102 1.59 103 1.64 10 ND 7.61 ND 2.07 10 1 3.27 7.82 ND ND ND ND ND
ND: not detected.
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The relationship between total elemental concentrations and soluble elemental concentrations measured by ICP after each batch experiment is summarized in Fig. 3. The pH in the mesophilic and thermophilic batch experiments was 7.3–7.7 and 7.9–8.5, respectively. The focus is the ratio of soluble to total elemental concentrations of the digested sludge, considering filtrate from digested sludge as liquid fertilizer or returning to the water treatment. The relationship was correlated with the periodic table. Alkali metals (Na and K) showed a similar tendency and were soluble, except in the thermophilic experiments fed with SM and 0.1
(gN-Increased ammonia/gCOD-produced methane)
mesophilic(MS) mesophilic(except MS) thermophilic(MS)
0.08
thermophilic(except MS) 0.06
0.04
0.02
0 0
10
20
30
40
50
substrate C/N (-) Fig. 2. Relationship between a substrate C/N ratio of organic wastes and increased ammonia per produced methane.
(A) 1000
M-Na
vegetables. The alkaline earth metals (Mg and Ca) followed an analogous trend. Around 20–40% of Mg and around 10–20% of Ca were soluble, regardless of the temperature conditions. Sui et al. (2011) performed ozonation of activated sludge to reduce sludge generation, and reported that K and Mg were solubilized at a higher ratio than COD solubilization, while the solubilized ratio of Ca was lower. Nagare et al. (2012) showed that more than 80% of P and K and around 60% of Ca and Mg were extracted from powdered mature corn using distilled water at 20 °C and 80 °C, regardless of the temperature difference. Solubilization of these elements might be caused by a simple physical phenomenon, rather than by biochemical digestion reactions. For the other elements, the ratio of soluble elements under thermophilic conditions was higher than under mesophilic conditions. Metalloids (B, As, and Sb) and nonmetallic elements (P and Se) had relatively higher ratios of soluble elements. Among the transition elements, the VIIIB elements (Co and Ni), which are essential for methanogenesis, had relatively higher ratios of soluble elements. The total elemental concentrations of Cr and Ni under thermophilic conditions were much higher than under mesophilic conditions. Digested sludge from Reactor 1 and Reactor 2 showed this concentration difference before the batch experiments. In Reactor 2, the concentrations were higher than the concentrations of substrate MS. This was probably not the result of co-digestion of some specific wastes, but Cr and Ni were possibly eluted from the reactor walls, particularly under thermophilic conditions. Naka et al. (2012) reported that P was absorbed onto Al(OH)3 at a ratio of 0.05 mmol P/mmol Al in mesophilic anaerobic digesters of sewage sludge where Al was present. Dabrowska and Rosinska (2012) reported that high temperatures could affect the release of metal ions despite the possibility of heavy metal precipitation in the form of sulfides in thermophilic anaerobic digestion of sewage sludge. The behavior of each element is affected by both the intake of each element by microorganisms and thermodynamic conditions including pH, temperature, coexisting materials, and ionic strength.
(B)
1000 M-Al
M-K T-K M-Mg
100
T-Mg M-Ca T-Ca y=x
T-Al
Soluble Conc. (mgE/L)
Soluble Conc. (mgE/L)
T-Na
10
M-Fe
100
T-Fe M-P T-P
10
y=x
1 10
100
1000
10
(C)
100
1000
Total Conc. (mgE/L)
1 M-B
(D)
10
Soluble Conc. (mgE/L)
Total Conc. (mgE/L)
1
M-Co T-Co
Soluble Conc. (mgE/L)
T-B
0.1
M-As T-As M-Sb
0.01
T-Sb M-Se
0.001
T-Se y=x
0.0001 0.001
0.01
0.1
Total Conc. (mgE/L)
1
M-Ni T-Ni M-Mn
0.1
T-Mn M-Zn
0.01
T-Zn y=x
0.001 0.01
0.1
1
10
100
Total Conc. (mgE/L)
Fig. 3. Relationship between total and soluble elemental concentrations measured by ICP after each batch experiment. M = mesophilic, T = thermophilic.
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3.3. Methane production and model application In the thermophilic batch experiments, the methane production rate, particularly at the beginning of the experiments, was not stable, and the results were not suitable for model application, although the final methane conversion performance was almost similar to that in the mesophilic experiments. Generally, thermophilic digestion has a higher reaction speed and a low viscosity (easy-mixing), while it is vulnerable to ammonia inhibition (Hidaka et al., 2013). In the present study, methane production performance in the mesophilic batch experiments was comparatively evaluated using the developed model. The parameters for the model application are summarized in Table 4. Most parameters were selected within the range of values summarized in the ADM1. In the present study, the multi-acid fermentation steps in the ADM1 were combined. Therefore, some parameters related to this part were determined by the experimental results. The remaining biodegradable COD in the inoculum sludge was considered to be Xs, and degradation of Xs was determined from the blank experiments. Initial Xsu and Xac concentrations were determined from the glucose experiments. The anaerobic degradation behavior of glucose is actually very complex with various products such as hydrogen, ethanol, and organic acids (Rodríguez et al., 2006). However, in the present batch experiments, the initial S/I ratio was lower, and 89.1 ± 3.0% of the added glucose was converted into methane on the basis of COD. Considering that some glucose is used for microbial synthesis, glucose can be used as an indicator of methane production from easily biodegradable soluble COD. No data were available on the proportion of Xsu to Xac, so the same concentrations were used. The methane conversion ratio of organic wastes was expressed as the ratio of Xr + Xm + Xs (convertible) to Xr + Xm + Xs + Xi (total). The methane production results in the batch experiments (gCOD-methane produced/gCOD-substrate) are slightly different from the ratio of Xr + Xm + Xs to Xr + Xm + Xs + Xi, because some of the Xr + Xm + Xs is used for Table 4 Parameters for the mathematical model.
Rate constant
Half saturation constant Yield coefficient
Nitrogen content
Symbol
Unit
Value
Remark
k1r k1m k1s k2 k3 k4su k4ac K2 Kac Y2 Y3 Y4 NXsu/Xsu, NXac/Xac
1/d 1/d 1/d L/(gCODd) L/(gCODd) 1/d 1/d gCOD/L gCOD/L
1.2 0.6 0.1 50 8 0.01 0.01 5 0.15 0.1 0.1 0.5 0.0875
Experiment Experiment Experiment Experiment ADM1 ADM1 ADM1 Experiment Experiment ADM1 ADM1 ADM1 ADM1
gN/gCOD
Xsu and Xac growth, part of which again becomes the substrate for methanogenesis after self-degradation of Xsu and Xac. The degradation constants for Xr and Xm were determined by referring to the fastest (at the beginning of the batch experiment, around days 0–2) and the mid-stage (at the middle of the batch experiment, around days 2–7) degradation results. The biodegradability of each organic waste is different. Inputting a different degradation constant for each waste is complicated when many types of wastes are co-digested. The developed model expresses the different degradation speeds by setting the ratios of Xr, Xm, and Xs (Table 5). Each organic waste includes not only particulate COD, but also soluble COD including volatile fatty acids (VFAs). In the present study, all substrate components were input as particulate COD except for glucose, although soluble COD was measured. Fast degradation of VFAs was observed at the beginning of the batch experiment, and the detailed methane production profile indicates more fractionation of substrate COD, which has a different degradation speed. However, the effect of such fast degradation is limited (Girault et al., 2012). To evaluate the methane production characteristics in practical continuous operation using the daily methane production profile, this simplification is acceptable, because the estimated value is expected to be slightly lower. The verification results of the developed model in the batch experiments are shown in Fig. 4. Results of S/I = 0.1 or 0.2 for the methane production profiles (accumulated methane and anaerobic respirometry) were used. When S/I = 0.4, the methane production speed seemed to be inhibited, which could be the result of an accumulation of VFAs or a decrease in pH, although the methane conversion ratio at the end of the experiments was not affected. Differences in methane production speeds among organic wastes were accurately calculated by dividing the substrate COD into Xr, Xm, Xs, and Xi. The ammonia concentration was also calculated accurately. The verification results of the developed model in the continuous experiment (Reactor 1) are shown in Fig. 5. The data for days 111–157, when MS, PS, CM, and SM were co-digested, were used. The methane production and the nitrogen concentrations were simulated well, and the applicability of the model to continuous operation was confirmed. Although VS loading rates of 1–3 kg VS/(m3 d) (JSWA, 2003) and 1.9–2.5 kg VS/(m3 d) (WEF and ASCE, 2009) have been suggested, the operational OLR in most WWTPs in Japan is as low as 1 kg VS/(m3 d) (JSWA, 2012). This is also the case in European Union (Di Maria et al., 2014). Furthermore, because of decreasing populations, particularly in small- and medium-sized cities, the operational OLR is decreasing. There is some room to accept more organic waste at many WWTPs. In practical operation, the VS loading rate is set lower than the suggested value, even if additional organic waste is fed, and digestion failure caused by overloading is not of great concern. When installing co-digestion systems, the estimated methane production, affected by the methane conversion ratio, degradation speed, and the operational HRT of anaerobic digesters, is required. For this purpose, the developed model can
Table 5 Substrate fraction for mathematical model.
MS PS KG CM SM CB PT CR
Ratio of biodegradable COD (Xr + Xm + Xs)/(Xr + Xm + Xs + Xi)
Ratio of Xr Xr/(Xr + Xm + Xs)
Ratio of Xm Xm/(Xr + Xm + Xs)
Ratio of Xs Xs/(Xr + Xm + Xs)
0.7 0.8 0.9 0.85 0.65 0.9 0.9 0.9
0.7 0.8 0.9 0.5 0.4 1 1 1
0.3 0.2 0.1 0 0.6 0 0 0
0 0 0 0.5 0 0 0 0
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T. Hidaka et al. / Waste Management xxx (2015) xxx–xxx
(A)
1.2
3
2.5
1
2.5
2
0.8
2
0.6
1.5
B1-1(exp.)
0.4 B1-1(calc.)
0 㻜
㻝㻜
㻞㻜
0.5 0
㻜
㻟㻜
㻝㻜
㻞㻜
2.5 2
4
4
1.5
3
3
B3-1(exp.)
2
0.5
B3-1(calc.)
1
B3-2(exp.) B3-2(calc.) B3-3(exp.) B3-3(calc.)
0
0 㻜
㻝㻜
㻞㻜
㻝㻜
㻞㻜
㻜
㻝㻜
㻞㻜
4
4
3
3 B3-9(exp.) B3-9(calc.) B3-10(exp.) B3-10(calc.)
2 1 㻝㻜
㻞㻜
㻞㻜
1 0.6 0.4
1 㻜
㻝㻜
㻞㻜
㻜
㻝㻜
㻞㻜
0.6 0.4
㻝㻜
1 0.8 0.6 0.4
㻞㻜
㻟㻜
2.5
B3-1(calc.)
B3-2(exp.) B3-2(calc.) B3-3(exp.) B3-3(calc.)
1.5
0.2
0 㻜
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B1-6(exp.) B1-6(calc.) B1-7(exp.) B1-7(calc.)
1.2
0.2 㻜
㻟㻜
B3-1(exp.)
㻟㻜
1.4 B1-4(exp.) B1-4(calc.) B1-5(exp.) B1-5(calc.)
0
3
0.4
㻞㻜
㻟㻜
0.2
0
0.5
㻝㻜
B3-12(exp.) B3-12(calc.) B3-13(exp.) B3-13(calc.) B3-14(exp.) B3-14(calc.)
0.2 0
㻜
㻟㻜
0.8
0.8
0.05
㻝㻜
2
㻟㻜
B1-2(exp.) B1-2(calc.) B1-3(exp.) B1-3(calc.)
1.2
B1-1(calc.) 0.1
㻟㻜
0 㻜
0.15
0.5
B3-4(exp.) B3-4(calc.) B3-5(exp.) B3-5(calc.) 㻜
㻟㻜
1.4 B1-1(exp.)
㻞㻜
1
5
㻟㻜
0.2
㻝㻜
2
5
0 㻜
B1-6(exp.) B1-6(calc.) B1-7(exp.) B1-7(calc.)
1
0
㻟㻜
B3-6(exp.) B3-6(calc.) B3-7(exp.) B3-7(calc.)
1.5
0 㻜
㻟㻜 5
1
2 B1-4(exp.) B1-4(calc.) B1-5(exp.) B1-5(calc.)
1
5
4 3.5 3 2.5 2 1.5 1 0.5 0
(B)
1.5 B1-2(exp.) B1-2(calc.) B1-3(exp.) B1-3(calc.)
0.5
0
3 2.5
1
0.2
3.5
B3-4(exp.) B3-4(calc.) B3-5(exp.) B3-5(calc.)
1.5 1
1 0.5
0.1
0.5
0
0 㻜
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0 㻜
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1
1.2
B3-6(exp.) B3-6(calc.) B3-7(exp.) B3-7(calc.)
0.8 0.6 0.4
㻜
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B3-9(exp.) B3-9(calc.) B3-10(exp.) B3-10(calc.)
1 0.8 0.6
1.5 1 0.5
0.2
0
0 㻜
㻝㻜
㻞㻜
0 㻜
㻟㻜
㻟㻜
B3-12(exp.) B3-12(calc.) B3-13(exp.) B3-13(calc.) B3-14(exp.) B3-14(calc.)
2
0.4 0.2
㻞㻜
2.5
㻝㻜
㻞㻜
㻟㻜
㻜
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㻟㻜
(C) NH4-N(calc.)(gN/L)
1.5
1.0
0.5 0.5
1.0 NH4-N(exp.)(gN/L)
1.5
Fig. 4. Verification results of the developed model in the batch experiment; (A) accumulated methane production, x-axis = time (d), y-axis = accumulated methane production (gCOD/L-reactor); (B) anaerobic respirometry, x-axis = time (d), y-axis = methane production rate (gCOD/(L-reactord)); (C) ammonia release.
Please cite this article in press as: Hidaka, T., et al. Comparative evaluation of anaerobic digestion for sewage sludge and various organic wastes with simple modeling. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.04.026
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T. Hidaka et al. / Waste Management xxx (2015) xxx–xxx
(m3-CH4/(m3-reactor.d))
(A)
CH4(exp.) CH4(calc.)
0.2 0.15 0.1 0.05 0 110
120
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140
150
160
time (d)
N conc. (gN/L)
(B)
3
NH4(exp.)
TN(exp.)
NH4(calc.)
TN(calc.)
2
1
0 110
120
130
140
150
160
time (d) Fig. 5. Verification results of the developed model in the continuous experiment (Reactor 1); (A) methane production; (B) ammonia and total nitrogen concentrations.
combine the batch experimental results and can be used effectively for the simple pre-estimation of co-digestion. 4. Conclusions In this study, the characteristics and anaerobic digestion performance of mixed sewage sludge and various organic wastes were compared, to develop a pre-evaluation tool to facilitate the introduction of co-digestion at WWTPs. The relationship between total and soluble elemental concentrations was correlated with the periodic table. The simple mathematical model was successfully applied to both batch and continuous experiments under mesophilic conditions. Co-digestion was shown to be a promising method to utilize various types of organic wastes at WWTPs, and the results of this study can be used as a pre-evaluation tool to introduce co-digestion. Acknowledgements We express our appreciation to members of the local governments and the B-DASH Project for their kind support. We are also grateful to the Shizuoka Prefectural Research Institute of Animal Industry, Swine and Poultry Research Center for providing the agricultural samples. This work was mainly conducted through a research fund for sewerage and sewage purification programs, the Ministry of Land, Infrastructure, Transport and Tourism, Japan. References APHA (American Public Health Association Publication), 1995. Standard Methods for the Examination of Water and Wastewater, 19th ed. Washington, DC, USA. Alzate, M.E., Muñoz, R., Rogalla, F., Fdz-Polanco, F., Pérez-Elvira, S.I., 2012. Biochemical methane potential of microalgae: influence of substrate to inoculum ratio, biomass concentration and pretreatment. Bioresour. Technol. 123, 488–494. Astals, S., Esteban-Gutiérrez, M., Fernández-Arévalo, T., Aymerich, E., García-Heras, J.L., Mata-Alvarez, J., 2013. Anaerobic digestion of seven different sewage sludges: a biodegradability and modelling study. Water Res. 47, 6033–6043. Batstone, D.J., Keller, J., Angelidaki, I., Kalyuzhnyi, S., Pavlostathis, S.G., Rozzi, A., Sanders, W., Siegrist, H., Vavilin, V., 2002. Anaerobic Digestion Model No. 1 (ADM1). IWA Publishing, London, UK.
Cabbai, V., Ballico, M., Aneggi, E., Goi, D., 2013. BMP tests of source selected OFMSW to evaluate anaerobic codigestion with sewage sludge. Waste Manage. 33, 1626–1632. Chen, Y., Cheng, J.J., Creamer, K.S., 2008. Inhibition of anaerobic digestion process: a review. Bioresour. Technol. 99, 4044–4064. Dabrowska, L., Rosinska, A., 2012. Change of PCBs and forms of heavy metals in sewage sludge during thermophilic anaerobic digestion. Chemosphere 88, 168– 173. Derbal, K., Bencheikh-lehocine, M., Cecchi, F., Meniai, A., Pavan, P., 2009. Application of the IWA ADM1 model to simulate anaerobic co-digestion of organic waste with waste activated sludge in mesophilic condition. Bioresour. Technol. 100, 1539–1543. Di Maria, F., Sordi, A., Cirulli, G., Gigliotti, G., Massaccesi, L., Cucina, M., 2014. Cotreatment of fruit and vegetable waste in sludge digesters. An analysis of the relationship among bio-methane generation, process stability and digestate phytotoxicity. Waste Manage. 34, 1603–1608. Donoso-Bravo, A., Perez-Elvira, S., Wouwer, A..Vande., Fdz-Polanco, F., 2012. Longterm hydrolytic capacity evaluation of a thermophilic anaerobic digester treating sewage sludge. Water Sci. Technol. 66, 2378–2384. Donoso-Bravo, A., Retamal, C., Carballa, M., Ruiz-Filippi, G., Chamy, R., 2009. Influence of temperature on the hydrolysis, acidogenesis and methanogenesis in mesophilic anaerobic digestion: parameter identification and modeling application. Water Sci. Technol. 60, 9–17. Esposito, G., Frunzo, L., Panico, A., Pirozzi, F., 2011. Model calibration and validation for OFMSW and sewage sludge co-digestion reactors. Waste Manage. 31, 2527– 2535. Girault, R., Bridoux, G., Nauleau, F., Poullain, C., Buffet, J., Steyer, J.-P., Sadowski, A.G., Béline, F., 2012. A waste characterisation procedure for ADM1 implementation based on degradation kinetics. Water Res. 46, 4099–4110. Hidaka, T., Wang, F., Togari, T., Uchida, T., Suzuki, Y., 2013. Comparative performance of mesophilic and thermophilic anaerobic digestion for highsolid sewage sludge. Bioresour. Technol. 149, 177–183. JSWA (Japan Sewage Works Association), 2003. Guideline for sewerage system operation, Tokyo, Japan. (in Japanese). JSWA (Japan Sewage Works Association), 2012. Annual report on sewerage system in Japan, 67, Tokyo, Japan. (in Japanese). Lowry, O.H., Rosebrough, N.J., Farr, L., Randall, R.J., Randall, R.J., 1951. Protein measurement with the folin phenol reagent. J. Biol. Chem. 193, 265–275. Mottet, A., François, E., Latrille, E., Steyer, J.P., Déléris, S., Vedrenne, F., Carrère, H., 2010. Estimating anaerobic biodegradability indicators for waste activated sludge. Chem. Eng. J. 160, 488–496. Nagare, H., Fujiwara, T., Inoue, T., Akao, S., Inoue, K., Maeda, M., Yamane, S., Takaoka, M., Oshita, K., Sun, X., 2012. Nutrient recovery from biomass cultivated as catch crop for removing accumulated fertilizer in farm soil. Water Sci. Technol. 66, 1110–1116. Naka, D., Goel, R., Yasui, H., 2012. Thermodynamic prediction model for inorganic composition of anaerobically digested sludge. Jpn. J. Water Treat. Biol. 48 (3), 81–88. Navaneethan, N., Topczewski, P., Royer, S., Zitomer, D., 2011. Blending anaerobic codigestates: synergism and economics. Water Sci. Technol. 63, 2916–2922. NILIM (National Institute for Land and Infrastructure Management), Ministry of Land, Infrastructure, Transport and Tourism, Japan, 2013a. B-DASH Project No.1, Guideline for introducing an energy management system using intensive solid– liquid separation technology, Technical Note of NILIM, No.736. (ISSN13467328), Tsukuba, Japan. (in Japanese). NILIM (National Institute for Land and Infrastructure Management), Ministry of Land, Infrastructure, Transport and Tourism, Japan, 2013b. B-DASH Project No.2, Guideline for introducing an Innovative Biogas Production System, Technical Note of NILIM, No.737. (ISSN1346-7328), Tsukuba, Japan. (in Japanese). Rajagopal, R., Massé, D.I., Singh, G., 2013. A critical review on inhibition of anaerobic digestion process by excess ammonia. Bioresour. Technol. 143, 632–641. Rodríguez, J., Lema, J.M., van Loosdrecht, M.C.M., Kleerebezem, R., 2006. Variable stoichiometry with thermodynamic control in ADM1. Water Sci. Technol. 54, 101–110. Sewerage and Wastewater Management Department, Ministry of Land, Infrastructure, Transport and Tourism, Japan, 2013. Efforts for use of methane gas in wastewater treatment of Japan, Municipal Wastewater Proceedings, Methane Expo 2013, Global Methane Initiative.. Souza, T.S.O., Carvajal, A., Donoso-Bravo, A., Peña, M., Fdz-Polanco, F., 2013. ADM1 calibration using BMP tests for modeling the effect of autohydrolysis pretreatment on the performance of continuous sludge digesters. Water Res. 47, 3244–3254. Sui, P., Nishimura, F., Nagare, H., Hidaka, T., Nakagawa, Y., Tsuno, H., 2011. Behavior of inorganic elements during sludge ozonation and their effects on sludge solubilization. Water Res. 45, 2029–2037. Takashima, M., Shimada, K., Speece, R.E., 2011. Minimum requirements for trace metals (iron, nickel, cobalt, and zinc) in thermophilic and mesophilic methane fermentation from glucose. Water Environ. Res. 83, 339–346. WEF (Water Environment Federation) and ASCE (American Society of Civil Engineers), 2009. Design of Municipal Wastewater Treatment Plants, Fifth Edition, McGraw-Hill Professional, USA. Yasui, H., Goel, R., Li, Y.Y., Noike, T., 2008. Modified ADM1 structure for modelling municipal primary sludge hydrolysis. Water Res. 42, 249–259.
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Waste Management journal homepage: www.elsevier.com/locate/wasman
Comparative evaluation of anaerobic digestion for sewage sludge and various organic wastes with simple modeling Taira Hidaka ⇑, Feng Wang, Jun Tsumori Recycling Research Team, Materials and Resources Research Group, Public Works Research Institute, 1-6, Minamihara, Tsukuba, Ibaraki 305 8516, Japan
a r t i c l e
i n f o
Article history: Received 20 August 2014 Accepted 20 April 2015 Available online xxxx Keywords: Anaerobic digestion Sewage sludge Organic waste Digested sludge Mathematical model
a b s t r a c t Anaerobic co-digestion of sewage sludge and other organic wastes, such as kitchen garbage, food waste, and agricultural waste, at a wastewater treatment plant (WWTP) is a promising method for both energy and material recovery. Substrate characteristics and the anaerobic digestion performance of sewage sludge and various organic wastes were compared using experiments and modeling. Co-digestion improved the value of digested sewage sludge as a fertilizer. The relationship between total and soluble elemental concentrations was correlated with the periodic table: most Na and K (alkali metals) were soluble, and around 20–40% of Mg and around 10–20% of Ca (alkaline earth metals) were soluble. The ratio of biodegradable chemical oxygen demand of organic wastes was 65–90%. The methane conversion ratio and methane production rate under mesophilic conditions were evaluated using a simplified mathematical model. There was reasonably close agreement between the model simulations and the experimental results in terms of methane production and nitrogen concentration. These results provide valuable information and indicate that the model can be used as a pre-evaluation tool to facilitate the introduction of co-digestion at WWTPs. Ó 2015 Elsevier Ltd. All rights reserved.
1. Introduction Anaerobic digestion of organic wastes with high moisture content, such as sewage sludge, kitchen garbage, food waste, and agricultural waste, is a promising method for both energy and material recovery. A wastewater treatment plant (WWTP) that can accept many kinds of organic waste for co-digestion could become a regional energy hub if the anaerobic digester produces more methane than the digestion of sewage sludge alone (Cabbai et al., 2013; Navaneethan et al., 2011; Di Maria et al., 2014). In 2011, the Sewerage and Wastewater Management Department of the Ministry of Land, Infrastructure, Transport and Tourism, Japan (2013) initiated the nationwide project called Breakthrough by Dynamic Approach in Sewage High Technology (B-DASH Project) for the development of innovative technologies in sewage treatment. Demonstration experiments were performed, including the co-digestion of sewage sludge and regional organic waste (NILIM, 2013a,b). However, it is likely that the characteristics of sewage sludge and regional organic waste are different in each WWTP. The properties of substrates such as total solids (TS), volatile solids (VS), chemical oxygen demand (COD), proteins, lipids, and nutrient ⇑ Corresponding author. Tel.: +81 29 879 6765; fax: +81 29 879 6797. E-mail address: [email protected] (T. Hidaka).
and trace elements can provide basic information for understanding anaerobic co-digestion characteristics. Co-digestion could potentially cause reaction inhibition by such as ammonia, metals and organics (Chen et al., 2008), increase loading to the water treatment process, and worsen the quality of effluent water from WWTPs. Some trace elements are required for methanogenesis, and some elements are important components of fertilizer. Basic information and a simple pre-evaluation tool for co-digestion are required to promote these technologies. Batch experiments are widely used to investigate methane conversion performance. Alzate et al. (2012) used batch experiments to compare anaerobic digestion of microalgae with substrate-toinoculum (S/I) ratios of 0.5, 1, and 3 (gVS/gVS), and reported that the methane production was higher at lower S/I ratios. The S/I ratios used in batch experiments are often higher than 0.5. Furthermore, the biodegradation profile in batch experiments affects the methane production quantity in continuous operation, especially in the case of a shorter hydraulic retention time (HRT). Assuming a first-order reaction and a continuously stirred tank reactor (Astals et al., 2013), theoretically not all of the degradable substrate is degraded, whereas almost all of the substrate is ultimately degraded in batch experiments. The sum of the methane conversion ratio multiplied by the feeding quantity is commonly used to design co-digestion systems, but this summation cannot accurately estimate the total methane production quantity in continuous operations.
http://dx.doi.org/10.1016/j.wasman.2015.04.026 0956-053X/Ó 2015 Elsevier Ltd. All rights reserved.
Please cite this article in press as: Hidaka, T., et al. Comparative evaluation of anaerobic digestion for sewage sludge and various organic wastes with simple modeling. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.04.026
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Mathematical models have been used to evaluate methane fermentation performance, combining operational parameters such as HRT, and substrate characteristics that differ among WWTPs. The International Water Association anaerobic digestion model No. 1 (ADM1) (Batstone et al., 2002) is an example of an effective general model that can be applied to many kinds of organic substrates. However, characterization of the biodegradability of substrates is required before using the ADM1. Yasui et al. (2008) and Astals et al. (2013) evaluated the sewage sludge from different WWTPs using anaerobic respirometry and mathematical modeling, but did not include other types of organic waste. Mottet et al. (2010) analyzed the thermophilic anaerobic biodegradability of waste activated sludge (WAS) from different WWTPs. Girault et al. (2012) proposed a COD fraction characterization method for the ADM1, but only two types of organic waste (pig slurry and WAS) were evaluated. Souza et al. (2013) proposed a simplified COD fractioning criterion for the ADM1, but it applied only to raw and pre-treated WAS. The ADM1 could be too complicated for use by designers and operational managers on site. Donoso-Bravo et al. (2009) proposed a two-population, three-reaction model, which was used to evaluate the thermophilic anaerobic digestion of sewage sludge (Donoso-Bravo et al., 2012). However, this simplified model did not include the self-degradation of microorganisms, which is important in batch experiments. Although extensive research on co-digestion using experiments and modeling has been reported (Derbal et al., 2009; Esposito et al., 2011), there have been few comparative evaluations using a wide range of organic wastes together with sewage sludge. In the present study, the substrate characteristics and anaerobic digestion performance of various organic wastes were compared using experiments and simplified mathematical modeling. The primary purpose was to provide municipal employees and designers on site with basic information and a simple pre-evaluation method to facilitate the introduction of co-digestion at WWTPs. 2. Materials and methods 2.1. Anaerobic digestion characteristics of various organic wastes Various organic wastes were compared in the present study. Glucose (GL) was used as a substrate to investigate the activity of the digested sludge. Mixed sewage sludge (MS), which is a mixture of primary sludge and WAS, was obtained from WWTP-A. Separated primary sludge (PS) from WWTP-B was obtained using an effective separation system demonstrated by the B-DASH Project (NILIM, 2013a), instead of a traditional primary
sedimentation system. Kitchen garbage (KG), which was a mixture of PS and raw KG, was obtained from the B-DASH Project (NILIM, 2013a). Cow manure (CM) and swine manure (SM) were obtained from an agricultural research institute. Cabbage (CB), potato (PT), and carrot (CR) were evaluated as representatives of waste leaf and root vegetables. All parts of the vegetables, including edible parts and peels, were crushed together and used for the experiments. As part of the B-DASH Project, imperfect vegetables from markets were planned to be co-digested with sewage sludge. Digested sludge adapted to various organic wastes was cultivated as inoculum sludge for the batch experiments. Two semi-continuous complete mix reactors, Reactor 1 and Reactor 2, each with a working volume of 100 L, were operated under mesophilic (35 °C) and thermophilic (55 °C) conditions, respectively. When the present study started, Reactor 1 and Reactor 2 had been operated for three months with MS. At the beginning of the present study, 3 L of the mesophilic digested sludge from the B-DASH Project demonstration anaerobic digester (NILIM, 2013b) was added to each reactor as an inoculum sludge adapted to MS and other waste biomass. A mixture of MS, PS, CM, and SM with average VS of around 37 g/L was fed to Reactor 1 and Reactor 2. The temperature was controlled using a hot water bath. The reactors were fed three times a week (on Monday, Wednesday, and Friday). The HRT was set at 120 d, and it was gradually decreased to 47 d, which corresponds to an organic loading rate (OLR) of around 0.8 kg VS/(m3 d). The main purpose was to cultivate inoculum sludge adapted to various organic wastes for the following batch experiments. Therefore, the HRT was set to be longer than the HRT of conventional anaerobic digesters (JSWA, 2003). Four series of batch experiments (B1, B2, B3, and B4) were performed using a commercially available experimental kit with 600 mL-vessels stirred at around 100 rpm (AMPTS II; Bioprocess Control AB, Sweden) (Table 1). Four hundred mL of the digested sludge and substrate were added to each vial. No microelements were added. The S/I ratios were set lower than for typical batch experiments, to determine rate constants under similar conditions to continuous operation. Before sealing, the vials were flushed with nitrogen gas. The produced biogas was passed through an alkali liquid (3 M NaOH) to remove CO2, and the quantity of produced methane gas was monitored using the gas meter (AMPTS II). Before starting each batch experiment, substrate feeding was suspended for about one week in Reactor 1 and Reactor 2 to reduce the background methane production. The digested sludge cultivated in Reactor 1 and Reactor 2 was used for B1 and B3 (mesophilic), and B2 and B4 (thermophilic), respectively. The substrate addition ratio was designed using the
Table 1 Batch experimental conditions.
a
Mesophilic
Thermophilic
B1
B2
Substrate
B1-1
B2-1
Blank
B1-2
B2-2
GL
B1-3
B2-3
GL
B1-4
B2-4
MS
B1-5
B2-5
B1-6
S/I a
Substrate
S/I a
Mesophilic
Thermophilic
B3
B4
−
B3-1
B4-1
Blank
−
0.1
B3-2
B4-2
GL
0.1
0.2
B3-3
B4-3
GL
0.2
0.1
B3-4
B4-4
PS
0.1
MS
0.2
B3-5
B4-5
PS
0.2
B2-6
KG
0.1
B3-6
B4-6
CM
0.1
B1-7
B2-7
KG
0.2
B3-7
B4-7
CM
0.2
B1-8
B2-8
KG
0.4
B3-8
B4-8
CM
0.4
B3-9
B4-9
SM
0.1
B3-10
B4-10
SM
0.2
B3-11
B4-11
SM
0.4
B3-12
B4-12
CB
0.2
B3-13
B4-13
PT
0.2
B3-14
B4-14
CR
0.2
substrate/inoculum (digested sludge) ratio on VS base
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Moderately biodegradable particulate COD (Xm, NXm)
Readily biodegradable particulate COD (Xr, NXr) Solubilization R1r
R1m
R1s
Slowly biodegradable particulate COD (Xs, NXs)
Inert particulate COD (Xi, NXi)
Biodegradable soluble COD (Cb, NCb)
(Y4) (1᧩Y4)
Inert soluble COD (Ci, NCi)
Acidogenesis R2 Acid fermenter (Xsu, NXsu)
(Y2) (1᧩Y2)
Self-degradation R4Xsu
Ammonia (Nam)
Acetate and hydrogen (Cac) Methanogenesis R3
Methanogens (Xac, NXac)
(Y3) (1᧩Y3)
Self-degradation R4Xac
Methane State variables: X, particulate COD (gCOD/L) C, soluble COD (gCOD/L) N, Nitrogen (gN/L)
Reaction rates: R1i᧹k1i×Xi (i=r, m, s) R2᧹k2×Cb/(K2+Cb)×Xsu R3᧹k3× Cac/(Kac+Cac)× Xac R4Xi᧹k4i× Xi (i=su, ac)
Fig. 1. Schematic diagram of the developed model.
VS ratio of substrate to inoculum digested sludge. Vegetable samples were crushed using an electric blender (ABS-V; Osaka Chemical Co., Ltd., Japan). Each experiment included blank experiments using only digested sludge. Additional blank experiments (B3-X1, -X2 and B4-X1, -X2) were performed using 500-mL conical flasks, and the quantity of produced biogas was measured to ensure experimental accuracy. 2.2. Mathematical model Fig. 1 shows a schematic diagram of the developed model. This model is based on a similar concept to that of the ADM1, but is simplified. In the ADM1, COD fractions are characterized by carbohydrates, proteins, and lipids. However, because accurate measurement of these items is time-consuming and costly, estimated ratios are often used for the ADM1 (Souza et al., 2013). In the present study, the substrate COD fraction was determined in terms of degradation speed rather than components. The substrate for methanogenesis is acetate and hydrogen, but in this case, acetate and hydrogen were considered as one state variable. Organic waste is divided into the following categories: readily biodegradable particulate COD (Xr), mildly biodegradable particulate COD (Xm), slowly biodegradable particulate COD (Xs), inert particulate COD (Xi), biodegradable soluble COD (Cb), and inert soluble COD (Ci). Particulate COD components are solubilized into Cb (R1, solubilization), Cb is converted into acetate and hydrogen (Cac) (R2, acidogenesis), and Cac is converted into methane (R3, methanogenesis). There are two types of microorganisms: acid fermenters (Xsu), which produce acetate and hydrogen from soluble COD such as sugar, and methanogens (Xac), which produce methane from acetate and hydrogen. Xsu and Xac are degraded into Xs and Xi (R4, self-degradation). Each COD fraction contains nitrogen (NXr, NXm, NXs, NXi, NCb, NCi, NXsu, and Nac), and ammonia (Nam) is produced corresponding to the degradation of
soluble COD into acetate and hydrogen. The mass balance of each state variable was calculated considering yield coefficients (Y2–Y4). Each reaction rate was expressed as the multiplication of a rate constant, the Michaelis–Menten kinetic, and a microorganism concentration. The effect of ammonia on microbial activity was expressed with a half saturation inhibition constant. In the present mesophilic experiments, pH was 7.3–7.7, and the total ammonia concentration was <1.1 g N/L, which would not lead to inhibition from ammonia accumulation (Rajagopal et al., 2013). The time course was calculated using the Runge–Kutta–Gill method. The program was developed using Visual Basic for Applications (VBA) in Microsoft Excel 2013, which could be used by municipal employees or designers on site. 2.3. Chemical analysis The pH, electrical conductivity (EC), TS, VS, and alkalinity were analyzed using standard methods (APHA, 1995). Total COD (TCOD) and soluble COD (SCOD) were analyzed using a spectrophotometer (DR3900; Hach Company, USA) and COD Digestion Vials (High Range, Hach Company). Particle COD (PCOD) was calculated as TCOD SCOD. Ammonia and total nitrogen were assayed using an auto analyzer (TRAACS2000; Bran Luebbe, Germany). Dissolved samples were prepared by filtration through a cellulose acetate filter (pore size = 0.45 lm). The carbon hydrogen nitrogen oxygen (CHNO) ratio was measured using an elemental analyzer (FLASH EA1112; Amco Inc., Japan). The elements Al, Ca, Fe, K, Mg, Na, and P, and B, Cr, Mn, Co, Ni, Cu, Zn, As, Se, Cd, Sb, and Pb were measured using inductively coupled plasma (ICP) atomic emission spectrometry (ICPS-8000E; Shimadzu, Japan), and ICP mass spectrometry (ICP-MS X series; Thermo Fisher, USA), respectively. Protein was measured using the Lowry method (Lowry et al., 1951), and lipid was measured by direct-heat extraction (SF-6; Sanshin Industrial Co. Ltd., Japan).
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and Mg compared with MS, and SM has higher containing ratios of Ca, K, Mg, Na, P, Co, Cu, and Zn. These elements are valuable in terms of the production of fertilizer. Co-digestion can therefore improve the value of digested sewage sludge as a fertilizer.
3. Results and discussion 3.1. Comparison of substrate characteristics The basic characteristics of MS and organic wastes are summarized in Table 2. MS was measured twice for batch experiments B1 and B2. The TS, VS, and TCOD were much higher in organic wastes than in MS. The C- and H-containing ratios were similar, except for KG. The C/N ratio of organic wastes was similar or higher than that of MS, which was round 9 ( ). Co-digestion can be used to increase the C/N ratio, by feeding with various organic wastes. The average ratio of g protein/g nitrogen measured using the elemental analyzer was around 5.7. The KG and PT had more lipids than other organic wastes. Elemental concentrations of MS and organic wastes are summarized in Table 3. The containing ratio of the organic wastes was generally lower than that of MS. Among these elements, Fe, Ni, Co, and Zn are typical nutrients required for methanogenesis (Takashima et al., 2011). While some organic wastes had relatively higher containing ratios of these elements compared with MS, most organic wastes had lower containing ratios. When organic wastes that do not contain sufficient nutrients are digested, the addition of these elements is recommended. However, receiving such organic wastes at WWTPs generally results in reduced addition of nutrients, because sewage sludge contains these elements. This is another potential advantage of co-digestion utilizing anaerobic digesters at WWTPs. CM has higher containing ratios of Ca, K,
3.2. Effect of co-digestion on quality of the digested sludge The relationship between the substrate C/N ratio of organic waste and increased ammonia per produced methane is shown in Fig. 2. Organic wastes with a C/N ratio <50 are shown. The quantity of methane produced was used as an index of degraded COD. Increased ammonia contributed by added organic wastes is estimated by extracting produced ammonia from the inoculum sludge. The MS had a C/N ratio of around 9, and organic wastes with higher C/N ratios produced less ammonia per produced methane, although there was some variation resulting from the different degradation characteristics of COD and nitrogen components. An ammonia concentration of around 3 g N/L is acceptable for mesophilic anaerobic digestion of MS, although it is not acceptable for thermophilic operation (Hidaka et al., 2013). Ammonia inhibition might not be a problem in anaerobic digestion of sewage sludge only, but it should be considered when organic wastes containing high N components are co-digested. The EC in the mesophilic and thermophilic batch experiments was 7.3–9.5 and 7.2–8.7 mS/cm, respectively. Alkalinity in the mesophilic and thermophilic batch experiments was 3.6–6.2 and 4.5–5.9 gCaCO3/L, respectively. The EC and alkalinity were markedly affected by an increase in ammonia.
Table 2 Substrate characteristics.
MSa MSb PS KG CM SM CB PT CR a b c
TS (g/kg)
VS (g/kg)
TCOD (g/kg)
SCOD (g/kg)
C (%-TS)
H (%-TS)
N (%-TS)
O (%-TS)
C/N (-)
Protein (%-TS)
23.1 17.3 11.9 131 161 241 95.3 181 97.6
19.3 14.1 10.5 111 144 194 87.5 171 89.7
28.5 23.0 17.8 166 155 268 116 179 114
1.1 1.8 2.4 48.1
41.1 41.0 44.1 53.4 44.2 40.2 39.9 41.9 40.0
6.1 6.2 6.7 7.8 6.3 6.2 6.2 6.6 6.3
4.3 4.8 2.3 3.4 1.9 4.1 2.0 0.9 0.2
33.7 30.5 32.4 31.9 33.0 29.5 41.5 42.3 42.5
9.6 8.6 19.5 15.6 23.8 9.9 20.0 45.1 194
24.1
2.8
17.3 12.7 16.5 24.4 7.0 8.7 3.9
4.5 21.3 2.3 3.0 4.2 9.7 0.8
c c
76.4 58.3 79.2
Lipid (%-TS)
Measured for B1. Measured for B2. Cannot be measured due to high TS.
Table 3 Elemental concentrations of substrate per dry solids (DS), mg E/kg DS.
Al Ca Fe K Mg Na P B Cr Mn Co Ni Cu Zn As Se Cd Sb Pb a
MS
PS
FW
FT
KG
CM
SM
CB
PT
CR
7.29 103 1.01 104 1.80 104 3.43 103 2.72 103 2.99 103 1.40 104 1.75 10 6.20 5.90 102 1.98 5.82 1.04 102 1.80 102 4.42 2.51 3.52 10 1 7.20 10 1 1.19 10
4.24 103 8.77 103 6.34 103 3.25 103 1.32 103 4.36 103 7.33 103 1.11 10 7.18 10 1.09 102 2.24 6.09 10 1.66 102 6.56 102 1.36 5.64 10 1 6.91 10 1 1.24 1.41 10
2.67 104 1.38 103 7.17 102 ND 2.05 102 5.63 102 5.02 103 1.80 4.69 1.06 10 ND 1.46 6.89 1.39 10 ND ND ND ND 4.15 10 1
5.46 10 1.97 103 1.09 102 1.09 103 2.73 102 ND ND 4.54 3.97 1.78 102 5.19 10 1 1.97 1.76 1.49 10 ND ND 2.35 10 1 ND 1.81
4.80 102 2.51 103 8.00 102 3.52 103 6.40 102 7.41 103 2.45 103 4.06 1.02 10 1.67 10 3.15 10 1 8.05 2.35 10 6.28 10 5.41 10 1 ND ND 1.65 10 1 3.35
3.24 102 1.53 104 4.86 102 7.57 103 6.11 103 1.73 103 5.95 103 1.80 10 1.64 1.31 102 3.60 10 1 2.20 1.93 10 6.79 10 2.11 10 1 ND 1.65 10 1 ND 2.80 10 1
3.57 102 3.18 104 2.34 103 1.31 104 9.07 103 4.89 103 1.97 104 1.91 10 4.25 3.52 102 2.68 5.22 1.90 102 3.66 102 7.56 10 1 9.47 10 1 3.23 10 1 ND 5.94
NDa 5.25 103 ND 2.51 104 1.52 103 ND 3.19 103 1.69 10 ND 1.03 10 ND 1.55 10 1 2.03 1.12 10 ND ND ND 4.27 10 1 ND
1.53 102 1.53 102 1.02 102 1.81 104 9.20 102 ND 1.48 103 5.88 ND 6.91 ND 2.67 10 1 2.67 8.59 ND ND ND ND ND
ND 2.33 103 ND 2.57 104 6.90 102 8.49 102 1.59 103 1.64 10 ND 7.61 ND 2.07 10 1 3.27 7.82 ND ND ND ND ND
ND: not detected.
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The relationship between total elemental concentrations and soluble elemental concentrations measured by ICP after each batch experiment is summarized in Fig. 3. The pH in the mesophilic and thermophilic batch experiments was 7.3–7.7 and 7.9–8.5, respectively. The focus is the ratio of soluble to total elemental concentrations of the digested sludge, considering filtrate from digested sludge as liquid fertilizer or returning to the water treatment. The relationship was correlated with the periodic table. Alkali metals (Na and K) showed a similar tendency and were soluble, except in the thermophilic experiments fed with SM and 0.1
(gN-Increased ammonia/gCOD-produced methane)
mesophilic(MS) mesophilic(except MS) thermophilic(MS)
0.08
thermophilic(except MS) 0.06
0.04
0.02
0 0
10
20
30
40
50
substrate C/N (-) Fig. 2. Relationship between a substrate C/N ratio of organic wastes and increased ammonia per produced methane.
(A) 1000
M-Na
vegetables. The alkaline earth metals (Mg and Ca) followed an analogous trend. Around 20–40% of Mg and around 10–20% of Ca were soluble, regardless of the temperature conditions. Sui et al. (2011) performed ozonation of activated sludge to reduce sludge generation, and reported that K and Mg were solubilized at a higher ratio than COD solubilization, while the solubilized ratio of Ca was lower. Nagare et al. (2012) showed that more than 80% of P and K and around 60% of Ca and Mg were extracted from powdered mature corn using distilled water at 20 °C and 80 °C, regardless of the temperature difference. Solubilization of these elements might be caused by a simple physical phenomenon, rather than by biochemical digestion reactions. For the other elements, the ratio of soluble elements under thermophilic conditions was higher than under mesophilic conditions. Metalloids (B, As, and Sb) and nonmetallic elements (P and Se) had relatively higher ratios of soluble elements. Among the transition elements, the VIIIB elements (Co and Ni), which are essential for methanogenesis, had relatively higher ratios of soluble elements. The total elemental concentrations of Cr and Ni under thermophilic conditions were much higher than under mesophilic conditions. Digested sludge from Reactor 1 and Reactor 2 showed this concentration difference before the batch experiments. In Reactor 2, the concentrations were higher than the concentrations of substrate MS. This was probably not the result of co-digestion of some specific wastes, but Cr and Ni were possibly eluted from the reactor walls, particularly under thermophilic conditions. Naka et al. (2012) reported that P was absorbed onto Al(OH)3 at a ratio of 0.05 mmol P/mmol Al in mesophilic anaerobic digesters of sewage sludge where Al was present. Dabrowska and Rosinska (2012) reported that high temperatures could affect the release of metal ions despite the possibility of heavy metal precipitation in the form of sulfides in thermophilic anaerobic digestion of sewage sludge. The behavior of each element is affected by both the intake of each element by microorganisms and thermodynamic conditions including pH, temperature, coexisting materials, and ionic strength.
(B)
1000 M-Al
M-K T-K M-Mg
100
T-Mg M-Ca T-Ca y=x
T-Al
Soluble Conc. (mgE/L)
Soluble Conc. (mgE/L)
T-Na
10
M-Fe
100
T-Fe M-P T-P
10
y=x
1 10
100
1000
10
(C)
100
1000
Total Conc. (mgE/L)
1 M-B
(D)
10
Soluble Conc. (mgE/L)
Total Conc. (mgE/L)
1
M-Co T-Co
Soluble Conc. (mgE/L)
T-B
0.1
M-As T-As M-Sb
0.01
T-Sb M-Se
0.001
T-Se y=x
0.0001 0.001
0.01
0.1
Total Conc. (mgE/L)
1
M-Ni T-Ni M-Mn
0.1
T-Mn M-Zn
0.01
T-Zn y=x
0.001 0.01
0.1
1
10
100
Total Conc. (mgE/L)
Fig. 3. Relationship between total and soluble elemental concentrations measured by ICP after each batch experiment. M = mesophilic, T = thermophilic.
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3.3. Methane production and model application In the thermophilic batch experiments, the methane production rate, particularly at the beginning of the experiments, was not stable, and the results were not suitable for model application, although the final methane conversion performance was almost similar to that in the mesophilic experiments. Generally, thermophilic digestion has a higher reaction speed and a low viscosity (easy-mixing), while it is vulnerable to ammonia inhibition (Hidaka et al., 2013). In the present study, methane production performance in the mesophilic batch experiments was comparatively evaluated using the developed model. The parameters for the model application are summarized in Table 4. Most parameters were selected within the range of values summarized in the ADM1. In the present study, the multi-acid fermentation steps in the ADM1 were combined. Therefore, some parameters related to this part were determined by the experimental results. The remaining biodegradable COD in the inoculum sludge was considered to be Xs, and degradation of Xs was determined from the blank experiments. Initial Xsu and Xac concentrations were determined from the glucose experiments. The anaerobic degradation behavior of glucose is actually very complex with various products such as hydrogen, ethanol, and organic acids (Rodríguez et al., 2006). However, in the present batch experiments, the initial S/I ratio was lower, and 89.1 ± 3.0% of the added glucose was converted into methane on the basis of COD. Considering that some glucose is used for microbial synthesis, glucose can be used as an indicator of methane production from easily biodegradable soluble COD. No data were available on the proportion of Xsu to Xac, so the same concentrations were used. The methane conversion ratio of organic wastes was expressed as the ratio of Xr + Xm + Xs (convertible) to Xr + Xm + Xs + Xi (total). The methane production results in the batch experiments (gCOD-methane produced/gCOD-substrate) are slightly different from the ratio of Xr + Xm + Xs to Xr + Xm + Xs + Xi, because some of the Xr + Xm + Xs is used for Table 4 Parameters for the mathematical model.
Rate constant
Half saturation constant Yield coefficient
Nitrogen content
Symbol
Unit
Value
Remark
k1r k1m k1s k2 k3 k4su k4ac K2 Kac Y2 Y3 Y4 NXsu/Xsu, NXac/Xac
1/d 1/d 1/d L/(gCODd) L/(gCODd) 1/d 1/d gCOD/L gCOD/L
1.2 0.6 0.1 50 8 0.01 0.01 5 0.15 0.1 0.1 0.5 0.0875
Experiment Experiment Experiment Experiment ADM1 ADM1 ADM1 Experiment Experiment ADM1 ADM1 ADM1 ADM1
gN/gCOD
Xsu and Xac growth, part of which again becomes the substrate for methanogenesis after self-degradation of Xsu and Xac. The degradation constants for Xr and Xm were determined by referring to the fastest (at the beginning of the batch experiment, around days 0–2) and the mid-stage (at the middle of the batch experiment, around days 2–7) degradation results. The biodegradability of each organic waste is different. Inputting a different degradation constant for each waste is complicated when many types of wastes are co-digested. The developed model expresses the different degradation speeds by setting the ratios of Xr, Xm, and Xs (Table 5). Each organic waste includes not only particulate COD, but also soluble COD including volatile fatty acids (VFAs). In the present study, all substrate components were input as particulate COD except for glucose, although soluble COD was measured. Fast degradation of VFAs was observed at the beginning of the batch experiment, and the detailed methane production profile indicates more fractionation of substrate COD, which has a different degradation speed. However, the effect of such fast degradation is limited (Girault et al., 2012). To evaluate the methane production characteristics in practical continuous operation using the daily methane production profile, this simplification is acceptable, because the estimated value is expected to be slightly lower. The verification results of the developed model in the batch experiments are shown in Fig. 4. Results of S/I = 0.1 or 0.2 for the methane production profiles (accumulated methane and anaerobic respirometry) were used. When S/I = 0.4, the methane production speed seemed to be inhibited, which could be the result of an accumulation of VFAs or a decrease in pH, although the methane conversion ratio at the end of the experiments was not affected. Differences in methane production speeds among organic wastes were accurately calculated by dividing the substrate COD into Xr, Xm, Xs, and Xi. The ammonia concentration was also calculated accurately. The verification results of the developed model in the continuous experiment (Reactor 1) are shown in Fig. 5. The data for days 111–157, when MS, PS, CM, and SM were co-digested, were used. The methane production and the nitrogen concentrations were simulated well, and the applicability of the model to continuous operation was confirmed. Although VS loading rates of 1–3 kg VS/(m3 d) (JSWA, 2003) and 1.9–2.5 kg VS/(m3 d) (WEF and ASCE, 2009) have been suggested, the operational OLR in most WWTPs in Japan is as low as 1 kg VS/(m3 d) (JSWA, 2012). This is also the case in European Union (Di Maria et al., 2014). Furthermore, because of decreasing populations, particularly in small- and medium-sized cities, the operational OLR is decreasing. There is some room to accept more organic waste at many WWTPs. In practical operation, the VS loading rate is set lower than the suggested value, even if additional organic waste is fed, and digestion failure caused by overloading is not of great concern. When installing co-digestion systems, the estimated methane production, affected by the methane conversion ratio, degradation speed, and the operational HRT of anaerobic digesters, is required. For this purpose, the developed model can
Table 5 Substrate fraction for mathematical model.
MS PS KG CM SM CB PT CR
Ratio of biodegradable COD (Xr + Xm + Xs)/(Xr + Xm + Xs + Xi)
Ratio of Xr Xr/(Xr + Xm + Xs)
Ratio of Xm Xm/(Xr + Xm + Xs)
Ratio of Xs Xs/(Xr + Xm + Xs)
0.7 0.8 0.9 0.85 0.65 0.9 0.9 0.9
0.7 0.8 0.9 0.5 0.4 1 1 1
0.3 0.2 0.1 0 0.6 0 0 0
0 0 0 0.5 0 0 0 0
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Fig. 4. Verification results of the developed model in the batch experiment; (A) accumulated methane production, x-axis = time (d), y-axis = accumulated methane production (gCOD/L-reactor); (B) anaerobic respirometry, x-axis = time (d), y-axis = methane production rate (gCOD/(L-reactord)); (C) ammonia release.
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(m3-CH4/(m3-reactor.d))
(A)
CH4(exp.) CH4(calc.)
0.2 0.15 0.1 0.05 0 110
120
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time (d)
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(B)
3
NH4(exp.)
TN(exp.)
NH4(calc.)
TN(calc.)
2
1
0 110
120
130
140
150
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time (d) Fig. 5. Verification results of the developed model in the continuous experiment (Reactor 1); (A) methane production; (B) ammonia and total nitrogen concentrations.
combine the batch experimental results and can be used effectively for the simple pre-estimation of co-digestion. 4. Conclusions In this study, the characteristics and anaerobic digestion performance of mixed sewage sludge and various organic wastes were compared, to develop a pre-evaluation tool to facilitate the introduction of co-digestion at WWTPs. The relationship between total and soluble elemental concentrations was correlated with the periodic table. The simple mathematical model was successfully applied to both batch and continuous experiments under mesophilic conditions. Co-digestion was shown to be a promising method to utilize various types of organic wastes at WWTPs, and the results of this study can be used as a pre-evaluation tool to introduce co-digestion. Acknowledgements We express our appreciation to members of the local governments and the B-DASH Project for their kind support. We are also grateful to the Shizuoka Prefectural Research Institute of Animal Industry, Swine and Poultry Research Center for providing the agricultural samples. This work was mainly conducted through a research fund for sewerage and sewage purification programs, the Ministry of Land, Infrastructure, Transport and Tourism, Japan. References APHA (American Public Health Association Publication), 1995. Standard Methods for the Examination of Water and Wastewater, 19th ed. Washington, DC, USA. Alzate, M.E., Muñoz, R., Rogalla, F., Fdz-Polanco, F., Pérez-Elvira, S.I., 2012. Biochemical methane potential of microalgae: influence of substrate to inoculum ratio, biomass concentration and pretreatment. Bioresour. Technol. 123, 488–494. Astals, S., Esteban-Gutiérrez, M., Fernández-Arévalo, T., Aymerich, E., García-Heras, J.L., Mata-Alvarez, J., 2013. Anaerobic digestion of seven different sewage sludges: a biodegradability and modelling study. Water Res. 47, 6033–6043. Batstone, D.J., Keller, J., Angelidaki, I., Kalyuzhnyi, S., Pavlostathis, S.G., Rozzi, A., Sanders, W., Siegrist, H., Vavilin, V., 2002. Anaerobic Digestion Model No. 1 (ADM1). IWA Publishing, London, UK.
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Please cite this article in press as: Hidaka, T., et al. Comparative evaluation of anaerobic digestion for sewage sludge and various organic wastes with simple modeling. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.04.026