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Strengthen effects of dominant strains on aerobic digestion and stabilization of the residual sludge
Bioresource Technology 235 (2017) 202–210
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Bioresource Technology journal homepage: www.elsevier.com/locate/biortech
Strengthen effects of dominant strains on aerobic digestion and stabilization of the residual sludge Yongjun Liu ⇑, Min Gao, Aining Zhang, Zhe Liu Key Lab of Northwest Water Resource, Ecology and Environment, Ministry of Education, Xi’an University of Architecture and Technology, Xi’an 710055, PR China
h i g h l i g h t s Strains which have obvious tolerance to the dissolved oxygen were isolated. Addition of dominant strains has a significant role on the sludge reduction. The stabilization time of sludge was shortened considerably. Accumulation of nitrogen and phosphorus in sludge supernatant was increased.
a r t i c l e
i n f o
Article history: Received 4 January 2017 Received in revised form 4 March 2017 Accepted 8 March 2017 Available online 11 March 2017 Keywords: Aerobic digestion of sludge Stabilization Dominant strains Strengthen
a b s t r a c t In order to strengthen the aerobic digestion of residual sludge, shorten the time of sludge stabilization and further reduce operating costs, 3 dominant strains identified as Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 were isolated from long-term aerobic digestion sludge. Results showed that the sludge stabilization time were reduced by 3–4 days compared with the control when the dominant strains were added to the process of sludge aerobic digestion. The addition of dominant strains accelerated the accumulation of TOC, nitrate nitrogen and ammonia nitrogen in the digestive solution at different levels, and it was beneficial to the dissolution of phosphorus. Controlling DO 3–5 mg/L, pH 6.5, the strains of Pseudomonas sp. L3 and Bacillus sp. L19 were combined dosing with the dosage of 2% in the process of sludge aerobic digestion, compared with the control, digestion rates of TOC and MLSS were increased about 19% and 16%, respectively. Ó 2017 Elsevier Ltd. All rights reserved.
1. Introduction As the activated sludge processes were widely used in various industries of sewage treatment, a large number of the residual sludges were also produced (Du and Li, 2017). The residual sludge contains lots of refractory organic matters, heavy metals and salts, as well as pathogenic microorganism, parasite eggs and other toxic and harmful substances, if the residual sludges were not stabilized properly, it would cause new pollution to the environment, and the problem of sludge pollution has posed serious challenge to human living environment (Zhu et al., 2017; Jin et al., 2016a,b). At present, the main disposal methods for residual sludge were sanitary landfill, sludge incineration, sludge compost and sludge for agricultural use, etc (Martínez-García et al., 2016). However, due to the site restrictions, high infrastructure and operating man-
⇑ Corresponding author. E-mail address: [email protected] (Y.J. Liu). http://dx.doi.org/10.1016/j.biortech.2017.03.060 0960-8524/Ó 2017 Elsevier Ltd. All rights reserved.
agement costs and secondary pollution, sludge disposal problem has not been fundamentally resolved (Wang et al., 2016). Reduction and stabilization were the main targets for sludge treatment. Sludge digestion including aerobic digestion and anaerobic digestion, its essence was that the microbial cells and organic matters were degraded into small molecular matters by the processes of endogenous respiration and microbial biodegradation, and then the small molecular matters were used again as substrates by the microorganisms to achieve the purpose of sludge stabilization and reduction (Zhang et al., 2016; Rafieenia et al., 2017). The sludge volume after digestion could be reduced by 30%-50%, and the dewatering performance of sludge was greatly improved, it was one of the feasible methods for sludge reduction and stabilization (Fall et al., 2014; Bahar and Ciggin, 2016). But the existed problems were that aerobic digestion required long time aeration, the operation costs were large, and infrastructure costs for anaerobic digestion were high, at the same time, operation management was complex (Anjum et al., 2016). Therefore, it was
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imperative to find an efficient and cost-effective sludge treatment technology. The knowledge of microbial population told us that in a given system, only the most suitable microorganisms could adapt to the environment and survive (Jin et al., 2016a, b). Based on this point, researchers in the world screened many efficient strains from the environment and added it to the specific system, the efficiency of metabolic decomposition of pollutants were enhanced by increasing the dominant microbial concentration and optimizing the microbial population structures (Liu et al., 2011; Manaia et al., 2003). The use of microbial enhanced technology could improve the reduction and stabilization of sludge, which provided a new way for sludge treatment (Li et al., 2009; Kim et al., 2002). The aerobic digestion processes of residual sludge were simulated by long time aeration in this study, the dominant microorganisms were screened and used to study on the effects of sludge stabilization in order to shorten the aerobic digestion time, and further reducing the operation costs, which has very important theoretical and practical significance for the efficient treatment of residual sludge.
2. Materials and methods 2.1. Experiment equipment and the residual sludge Four sets of experimental devices were used in this study, each device was 45 cm high, and the inner diameter was 10 cm (Fig. 1). Aeration header was arranged at the bottom of the device. The sampling ports were set at 5 cm, 20 cm and 35 cm distance from the bottom of the device, respectively, sampling from 3 sampling
1
ports simultaneously, and then the samples were mixed for detection. The residual sludge used in this study were taken from the second sedimentation tank of the Fourth sewage treatment plant of Xi’an city, China, the properties of sludge were shown in Table 1. 2.2. Domestication and isolation of superior strains for sludge aerobic digestion 1 mL Excess sludge taken from aerobic digestion reactor were diluted to 101, 102, 103, 104 and 105 five concentration gradients in the bacteria-free environment, and the dilution plate method was used for separation of the microorganisms from different concentration gradients, cultural conditions were as follows: Lucia-Bertani solid culture medium (peptone 10 g, yeast extract 10 g, NaCl 5 g, agar 18 g, distilled water 1000 mL), pH 6.8–7.2, 30 °C, for 24 h. The single strain which has different colony morphology and microscopic structure was picked out respectively and saved. Method of acclimation was divided into 5 steps. The first step, the isolated strains mentioned above were cultured with LB liquid culture medium (no agar) at 30 °C, for 24 h, respectively. 30 mL culture liquid of different strains were taken, respectively, then centrifuged under 4000 rpm for 20 min. Remove supernatant, and the sediments were rinsed two times with pure water. The obtained different single strain cells were mixed, then added to 100 mL residual sludge, and acclimated under aerobic condition for 10 days. The second step, 30 mL residual sludge after acclimation was taken, then added to 70 mL fresh excess sludge, and acclimated under aerobic condition for 10 days. The third to the fifth step, repeat the second step. 1 mL sludge taken from the last acclimation step were diluted to 101, 102, 103, 104 and 105 five concentration gradients in the bacteria-free environment, then cultured in LB solid medium at 30 °C, for 24 h, the single strain which has different colony morphology and microscopic structure was picked out and named as L1-L25, respectively. 2.3. Screening of dominant strains for aerobic digestion The isolated strains named as L1-L25 mentioned above were cultured in the liquid LB medium at 30 °C, for 24 h, respectively. 30 mL culture liquid of different strains were taken, then centrifuged under 4000 rpm for 20 min. Remove supernatant, and the sediments were rinsed two times with pure water. The obtained different single strain cells were added to 100 mL new residual sludge, respectively, then cultured under aerobic condition. TOC degradation rate of each test group was measured 6 days later, the dominant strains for aerobic digestion were screened out.
2
6 3
2.4. Identification of the dominant strains
4
5
Fig. 1. Diagram of the aerobic digestion reactor. 1: Main device of the reactor, 2: Sampling holes; 3: Aeration plate, 4: Gas flow meter, 5: Aeration pump, 6: Temperature controller.
16s rDNA fragments of the screened dominant strains of aerobic digestion were amplified with the primers 27F (50 -AGAGTTT GATCMTGGCTCAG-30 ) and 1492R (50 -CGGYTACCTTGTTACGACTT30 ). PCR reaction was performed according to Cai et al. (2003). The PCR products were cloned into the pMD 18-T vector (TaKaRa Biotechnology Company) following the manufacturer’s instruction.
Table 1 Properties of the residual sludge in this study. The residual sludge flocs
The sludge supernatant
TOC (mg/L)
MLSS (mg/L)
SVI (mL/g)
pH
TOC (mg/L)
NO 3 (mg/L)
NH+4 (mg/L)
TP (mg/L)
3535.3 ± 270.1
9770.4 ± 328.6
95.8 ± 6.6
6.8 ± 0.2
19.3 ± 3.2
24.6 ± 2.2
0.5 ± 0.1
26.5 ± 2.3
Notes: TOC: Total organic carbon; MLSS: Mixed liquid suspended solid; SVI: Sludge volume index; TP: Total phosphorus.
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16s rDNA fragments were sequenced by Shanghai Biological Engineering Technology and Services Co., Ltd. BLAST program was used to search the similar sequences of 16S rRNA database, and the phylogenetic trees were constructed with the software ClustalX Version1.81 by using the neighbor-joining method. The 16S rRNA gene fragments of strains named as L3, L16 and L19 were deposited in the GenBank data library with the ID no. 1999166, 1999172 and 1999175, respectively. 2.5. Test of enhanced effect of dominant strains on residual sludge stabilization Dominant strains named as L3, L16, L19 were cultured in the LB medium at 30 °C for 24 h, respectively. 200 mL culture liquid of different strains were taken, then centrifuged under 4000 rpm for 20 min. Remove supernatant, and the sediments were rinsed two times with pure water. The obtained different microbial cells were suspended with pure water to about 10 101°Cfu/mL, the cell suspension was added to the reactor with 2 L fresh residual sludge at the volume ratio (bacteria/residual sludge) of 1%, with no addition of bacteria cells as control, DO in reactors were controlled to 3– 5 mg/L. Sampling every 24 h, TOC, MLSS, SVI of the residual sludge flocs and TOC, nitrate nitrogen, ammonia nitrogen, TP in the sludge supernatant were measured. The standard for the sludge stabilization of aerobic digestion was that the TOC degradation rate attained to 40%. 2.6. Detection methods In this study, TOC, MLSS and SVI of the residual sludge flocs, the total phosphorus, nitrate nitrogen and ammonia nitrogen in the sludge supernatant were measured by using the standard determination methods from ‘‘water and wastewater monitoring analysis method” (SEPAC, 2002a). 3. Results and discussion 3.1. Screening and identification of dominant strains for sludge aerobic digestion In the course of microbial growth, some microorganisms could utilize the substrates of the second time which formed by the death cell to grow, this growth pattern was called invisible growth (Yan et al., 2008), and the dissolution of dead cells was the key step in this course. Therefore, aerobic digestion of excess sludge might exist some flora which conducive to the dissolution of microbial cells, and the sludge stabilization by aerobic digestion was a kind of excess sludge treatment technology based on invisible growth (Bomio et al., 1989; Jang et al., 2014). 25 dominant strains with good invisible growth were isolated by domestication for many times in this study, the strains were named as L1-L25, respectively. After further sludge digestion test, TOC aerobic degradation rates (average value) of residual sludge by single strain L3, L16 and L19 were 31.2%, 28.8% and 32.7% respectively, within 6 days were used for further study. The TOC degradation rates for other strains were relatively lower from 16.1% to 23.6% within 6 days and not involved in this study. 16s rRNA gene fragments of strains L3, L16 and L19 were obtained respectively, by PCR amplification, and then submitted for sequencing. The blast search of the sequences indicated that the strain L3 was closely related to the species in genus of Pseudomonas, it also exhibited the highest similarity (99.35%) to Pseudomonas sp. and thus tentatively classified strain L3 as Pseudomonas sp. L3 (ID1999166), Fig. 2(A). Strain L16 was matched at 97.68% to Acinetobacter sp. Thus, identified as Acinetobacter sp.
L16 (ID1999172), Fig. 2(B). Strain L19 was matched at 98.39% to Bacillus sp., which was identified as Bacillus sp. L19 (ID1999175), Fig. 2(C). 3.2. Variations of TOC, MLSS and pH during the process of long term residual sludge aerobic digestion The variations of TOC and MLSS in the process of aerobic digestion were related to the length of aeration time, and the changes of pH directly affected the activity of microorganisms, further affected the aerobic digestion efficiency. Determination of digestion time was very important for saving energy and reducing operating costs. As shown in Fig. 3, after 80 days aerobic digestion, MLSS were decreased from 8900 mg/L to 3200 mg/L, and reduced by 64.0%. TOC were decreased from 3800 mg/L to 1250 mg/L, and reduced by 67.1%. During the aerobic digestion, MLSS and TOC were decreased faster in the initial 20 days, and began to enter the plateau after 20 days. When the digestion time was estimated by empirical method, usually activated sludge were 15–20 days, mixed sludge were 20–25 days (Martins et al., 2003), which were basically consistent with the experimental results of this study. During aerobic digestion of sludge, pH gradually decreased from pH 6.67 to pH 4.46. On the 70th day, the pH reached a minimum pH 3.08, and then increased gradually from pH 3.08 to pH 4.46 on the 80th day (Fig. 3). In the process of aerobic digestion, there was a significant negative correlation between pH and nitrate content, the decrease of pH was mainly due to nitrification (Jang et al., 2014). 70 days later, pH began to rise in this study, indicating a gradual decrease of nitrate content. Studies have shown that in aerobic digestion process, with the digestion of microbial cells, the intracellular substances were continuously dissolved, and the content of total phosphorus was also accumulated in the solution (Bi et al., 2013). Under the condition of large amount of nitrate and total phosphorus in sludge, the denitrifying phosphorus removal reaction with nitrate as electron acceptor could be induced, denitrifying phosphorus removal decreased nitrate content and pH increased gradually (Jin et al., 2015). 3.3. Enhanced effects of dominant strains on aerobic digestion and stabilization of the residual sludge 3.3.1. The time of residual sludge stabilization completion and the enhancement effects of the dominant strains According to the pollutant discharge standard of the municipal wastewater treatment plant, the degradation rate of organic matter in sludge aerobic digestion reached 40%, which was regarded as the stabilization of excess sludge had completed (SEPAC, 2002b). Results of the enhancement effects by the dominant strains on aerobic digestion of the residual sludge were shown as Table 2 and Fig. 4. Variations of TOC during sludge aerobic digestion were shown as Table 2, the control group had reached the sludge TOC 40% degradation rate on the 13th day, the groups of addition strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 had completed the sludge stabilization on the 9th day, 10th day and 9th day, respectively. Compared with the control group, completion time of sludge stabilization were advanced for 4d, 3d and 4d, respectively, indicated that added of the dominant strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 could obviously improve the efficiency of sludge stabilization (Fig. 4A). Variations of MLSS values during sludge aerobic digestion by adding dominant strains were shown as Table 2. Because of the completion time of sludge stabilization for the control group was on the 13th day in this study, the sludge aerobic digestion time was determined for 2 weeks. The degradation rate of MLSS after
Y.J. Liu et al. / Bioresource Technology 235 (2017) 202–210
205
Fig. 2. Phylogenetic trees based on 16S rDNA sequences of strains named as L3, L16, L19 and their related species. A: Strain L3 (ID1999166) closely related to the species in genus of Pseudomonas and thus identified as Pseudomonas sp. L3; B: Strain L16 (ID1999172) closely related to the species in genus of Acinetobacter and thus identified as Acinetobacter sp. L16; C: Strain L19 (ID1999175) closely related to the species in genus of Bacillus and thus identified as Bacillus sp. L19.
bacterial addition was increased with the passage of time, and higher evidently than that the control at the end of aerobic digestion. When the digestion reactions were carried out to 13th day, the MLSS degradation rates of adding strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 were 40.1%, 36.8% and 41.3%, respectively, compared with the control (33.1%) increased by 7.0%, 3.7% and 8.2%, respectively, indicated that the dominant microorganism could promote the residual sludge reduction (Fig. 4B). The results of SVI variations during sludge stabilization by adding strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19, respectively were showed as Table 2. When all test groups reached the sludge stabilization state (on the 13th day), compared with the control group, SVI of Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 strengthen groups decreased
by 14.1%, 9.0% and 11.1%, respectively, indicated that the addition of the dominant strains were beneficial to sludge dewatering treatment. The operation cost of residual sludge treatment could often account for 50%–60% of sewage treatment plant operation costs, and the energy expenditure of aeration in aerobic digestion accounted for most of the aerobic digestion costs (Abe et al., 2011). In the process of sludge aerobic digestion, adding dominant microorganisms could significantly improve sludge digestion efficiency and shorten sludge stabilization time in this study. Strengthen effects of strains Pseudomonas sp. L3 and Bacillus sp. L19 could shorten the aeration time by 31% in theory, which could reduce the operating costs of sewage treatment plant effectively. Adding dominant strains could shorten the time required for
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9000
decomposition and utilization by other microorganisms (Du and Li, 2017; Yan et al., 2008).
7
MLSS
7000
TOC
6
pH
6000
5
3.3.2. Variations of dissolved substances during the process of sludge stabilization Variations of dissolved substances in the supernatant of the different test groups during the sludge stabilization process were shown in Fig. 5. When the digestive reaction proceeded to 13th day, TOC in the supernatant of different reactors which were added the dominant strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 respectively, were increased to 60 mg/L, 42 mg/L and 53.4 mg/L, respectively. Compared with the control (40 mg/L), TOC cumulative volume in strengthen reactors increased by 50%, 5% and 33.5%, respectively (Fig. 5A). Nitrate nitrogen, ammonia nitrogen and dissolved phosphorus all had different degrees of accumulation in the digestive juice. On the 13th day, the amount of nitrate nitrogen in the supernatants of strengthen groups by adding strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 increased to 105.6 mg/L, 81.8 mg/L and 93.0 mg/L, respectively (Fig. 5B), compared with the control (69.2 mg/L), increased by 52.6%, 18.2% and 34.4%, respectively. The amount of ammonia nitrogen increased to 112.7 mg/L, 99.64 mg/L and
5000 4 4000 3000
pH
MLSS and TOC (mg/L)
8000
3
2000 2
1000 0
1 0
5 10 15 20 25 30 35 40 45 50 55 60 65 70 75 80 Aerobic digestion time (d)
Fig. 3. Variations of the mixed liquid suspended solid (MLSS), total organic carbon (TOC) and pH during long term aerobic digestion of the residual sludge at 25 °C.
sludge stabilization, the main reason might be that the dominant microorganisms could secrete more extracellular enzymes, these hydrolases could accelerate the hydrolysis of the dead cells, improve the efficiency of sludge hydrolysis, promote macromolecular organic matter degradation into small molecules, further
Table 2 Variations of the TOC, MLSS and SVI during the process of the residual sludge aerobic digestion by adding dominant strains.
TOC (mg/L)
Time (day)
MLSS (mg/L)
SVI (mL/g)
control
L3
L16
L19
control
L3
L16
L19
control
L3
L16
L19
0
3794.3
3796.8
3798.5
3794.2
8900.2
8896.5
8892.8
8898.0
95.3
94.8
95.1
94.6
1
3705.1
3699.3
3701.2
3696.6
8677.5
8659.7
8658.9
8657.9
96.2
98.2
100.6
97.2
2
3610.3
3572.1
3496.2
3534.7
8321.5
8205.8
8232.5
8232.5
100.6
104.1
102.5
100.1
3
3477.6
3344.3
3306.4
3325.0
8099.6
7786.5
7832.8
7743.3
103.4
106.3
106.3
105.4
4
3268.5
3154.3
3116.3
3135.4
7787.5
7342.5
7565.5
7253.5
105.1
108.2
110.7
109.7
5
3161.6
2850.0
2888.3
2812.7
7431.5
6897.5
7342.5
7075.5
106.3
105.7
110.5
104.8
6
3040.0
2625.6
2698.0
2565.0
7120.0
6452.5
7031.3
6897.5
110.0
103.6
110.3
100.3
7
2850.0
2470.0
2621.8
2432.4
6897.5
6230.0
6675.7
6452.5
110.3
95.0
100.2
90.5
8
2622.6
2321.8
2470.7
2394.5
6586.2
6007.5
6230.0
6006.5
110.6
90.8
90.0
85.2
9
2546.3
2280.5
2394.6
2276.8
6408.2
5785.6
6007.5
5783.4
100.7
83.7
85.6
81.7
10
2470.0
2223.8
2279.3
2192.6
6230.0
5696.7
5874.6
5740.5
90.5
79.4
80.5
76.3
11
2432.2
2166.7
2204.3
2090.7
6096.5
5607.4
5829.5
5607.8
86.4
71.7
74.4
72.4
12
2394.5
2090.0
2158.4
2054.3
6052.5
5518.2
5785.4
5473.5
80.0
67.6
70.6
69.2
13
2295.2
2052.0
2128.2
2014.6
5951.6
5329.3
5613.6
5224.0
75.7
65.0
68.8
67.2
14
2268.6
2013.3
2105.2
1968.4
5918.5
5286.6
5624.8
5188.7
73.7
63.4
66.7
65.8
Notes: (1) TOC: Total organic carbon; MLSS: Mixed liquid suspended solid; SVI: Sludge volume index. (2) Conditions of the aerobic digestion: inoculation amount of the dominant strains was 1% (volume ratio); dissolved oxygen 3–5 mg/L, 25 °C and natural pH. (3) Control: indicates no dominant strain addition; L3: means inoculation strain Pseudomonas sp. L3; L16: stands for strain Acinetobacter sp. L16; L19: stands for strain Bacillus sp. L19. (4) The values shown in the table were the average values of the 3 parallel samples.
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50
60
TOC digestion rate (%)
MLSS elimination rate (%)
45
50
A
40 30
control L3 L16 L19
20 10
40
B
35 30 25
control L3 L16 L19
20
15 10 5
0
0 0
1
2
3
4
5
6
7
8
9
10 11 12 13 14
0
1
2
3
Aerobic digestion time (d)
4
5
6
7
8
9
10 11 12 13 14
Aerobic digestion time (d)
Fig. 4. Effects of the different dominant strains on the mixed liquid suspended solid (MLSS) and total organic carbon (TOC) in the process of the residual sludge aerobic digestion under the condition of dissolved oxygen 3–5 mg/L, 25 °C and natural pH. A: Effects of the different dominant strains on the TOC. Control: indicates no dominant strain addition; L3: means inoculation strain Pseudomonas sp. L3 at the rate of 1% (volume ratio), the same below; L16: stands for strain Acinetobacter sp. L16; L19: stands for strain Bacillus sp. L19. B: Effects of the different dominant strains on the MLSS.
120
A
60
NO3 --N in supernatant (mg/L)
TOC in supernatant (mg/L)
70
50 40 30
20 control
10
L3
L16
L19
0
B
100 80 60 40
control
20
L16
L19
0 0
2
4
6
8
10
12
14
0
2
4
6
8
10
12
14
Aerobic digestion time (d)
Aerobic digestion time (d)
80
140 120
70
C
TP in supernatant (mg/L)
NH 3 -N in supernatant (mg/L)
L3
100
80 60 40 control
20
L3
L16
L19
D
60
50 40
30 20 control
10
L3
L16
L19
8
10
12
0
0
0
2
4
6
8
10
12
14
Aerobic digestion time (d)
0
2
4
6
14
Aerobic digestion time (d)
Fig. 5. Variations of dissolved substances in the sludge supernatant during the process of enhanced sludge stabilization under the condition of natural pH and 25 °C. A: Total organic carbon in the sludge supernatant. Control indicates no dominant strain addition, L3 means inoculation strain Pseudomonas sp. L3 at the rate of 1% (volume ratio), the same below, L16 stands for strain Acinetobacter sp. L16, L19 stands for strain Bacillus sp. L19. B: Nitrate nitrogen in the sludge supernatant. C: Ammonia nitrogen in the sludge supernatant. D: Total phosphorus in the sludge supernatant.
118.4 mg/L, respectively (Fig. 5C), compared with the control (88.5 mg/L), increased by 27.3%, 12.5% and 33.8%, respectively. The amount of dissolved phosphorus increased to 74.8 mg/L, 69.4 mg/L and 70.6 mg/L, respectively (Fig. 5D), compared with the control (58.5 mg/L), increased by 27.9%, 18.6% and 20.7%, respectively. The accumulation of the TOC, nitrate nitrogen, ammonia nitrogen and dissolved phosphorus in the bacterial strengthen
group were all higher than that the control. The contents of nitrate and total phosphorus in digestive juice were positively correlated with the degradation rate of MLSS, indicated that the accumulation of nitrogen and phosphorus in digestive juice were mainly due to the degradation of MLSS. The source of nitrogenous substances during aerobic digestion of sludge was mainly the ammonia nitrogen which was produced
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60
60
under natural pH
pH 6.5
50 40 30
A
20
50
MLSS elimination rate (%)
TOC digestion rate (%)
70
10
under natural pH
40 30 20
B
10
0
0 0
2
4
6
8
10
12
0
14
2
4
Aerobic digestion time (d)
70 1%
2%
5%
MLSS elimination rate (%)
TOC digestion rate (%)
0.50%
50 40 30
C
20 10 0
8
10
12
14
0.50%
50
1%
2%
5%
40
30 20
D
10 0
0
2
4
6
8
10
12
14
0
2
4
Aerobic digestion time (d)
6
8
10
12
14
Aerobic digestion time (d)
60
60
DO 4mg/L DO 7mg/L
MLSS elimination rate (%)
DO 2mg/L DO 6mg/L
50 TOC digestion rate (%)
6
Aerobic digestion time (d)
60
60
40 30 20
E
10
0
DO 2mg/L DO 6mg/L
50
DO 4mg/L DO 7mg/L
40 30 20
F
10
0 0
2
4
6
8
10
12
14
0
Aerobic digestion time (d)
50
MLSS elimination rate (%)
L19 L19+L3 L19+L16 L19+L16+L3
60
40 30
G
20
2
4
10
6
8
10
12
14
Aerobic digestion time (d)
60
70
TOC digestion rate (%)
pH 6.5
L19 L19+L3 L19+L16 L19+L16+L3
50 40 30 20
H
10 0
0
0
2
4
6
8
10
Aerobic digestion time (d)
12
14
0
2
4
6
8
10
12
14
Aerobic digestion time (d)
Fig. 6. Analysis of influencing factors of sludge aerobic digestion. A: Effect of pH on the total organic carbon (TOC); B: Effect of pH on the mixed liquid suspended solid (MLSS); C: Effect of inoculation amount (volume ratio) on the TOC; D: Effect of inoculation amount on the MLSS; E: Effect of dissolved oxygen (DO) on the TOC; F: Effect of DO on the MLSS; G: Effect of different dominant strains combination on the TOC, L3, L16, L19 stand for strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19, respectively; H: Effect of different strains combination on the MLSS.
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by the sludge hydrolysis and ammonization continuously, and then converted to nitrate nitrogen by nitrification. Nitrate nitrogen and ammonia nitrogen were accumulated from beginning to end in this study, the results were differ from those obtained by other researchers (Li et al., 2014; Lee et al., 2004) whose results showed that ammonia accumulated to a certain amount would no longer accumulate. The main reason was that the sludge digestion time was shorter in this study, and the number of nitrifying bacteria was less, which led to a slower conversion rate from ammonia nitrogen to nitrate nitrogen, and the ammonia nitrogen had been always accumulated in the digestive juice. The dissolution rate of phosphorus in sludge depended on the degree of hydrolysis of sludge microbial cells (Bi et al., 2013). The dissolution rate of phosphorus in the bacterial addition group was higher than that in the blank control in this study, indicated that the added microbes were favorable for sludge hydrolysis. However, the trend of phosphorus change curves were similar, indicated that the added microorganisms did not change the dissolution mechanism of phosphorus in the process of sludge aerobic digestion. 3.4. Analysis of influencing factors in the process of enhanced sludge aerobic digestion During short-term aerobic digestion (two weeks in this study), pH decreased gradually. 1 mol/L NaOH was added to the digestion reactor every 12 h to regulate pH 6.5, dissolved oxygen (DO) was controlled to 3-5 mg/L, and strain Bacillus sp. L19 was added as the test strain. On the 13th day, degradation rates of TOC (Fig. 6A) and MLSS (Fig. 6B) were 56.3% and 48.4% respectively, when the pH value was adjusted to pH6.5. Degradation rates of TOC and MLSS were 48.7% and 40.8%, respectively, in the reactor without adjusting pH, indicated that controlling pH value during sludge aerobic digestion had a significant effect on sludge stabilization. In fact, sludge aerobic digestion process was the role of microbial endogenous respiration, and the appropriate pH was favorable for microorganism to maintain its activity. When pH value was too low, it would make the changes of microbial cell permeability, thus changing the microbial enzyme catalytic capacity, and ultimately affect the microbial mass transport and metabolism process (Ichinari et al., 2008; Yan et al., 2008). DO was controlled to 3-5 mg/L, pH value was controlled to pH 6.5, and strain Bacillus sp. L19 was added as the test strain. On the 13th day, when the dosing ratios of test strain Bacillus sp. L19 were 0.5%, 1%, 2% and 5% (volume ratio), respectively, TOC degradation rates were 45.3%, 55.6%, 58.2%, 59.1% (Fig. 6C), and MLSS removal rates were 33.9%, 42.3%, 46.4%, 49.8% (Fig. 6D), respectively. The results indicated that the higher dosing ratio could improve the efficiency of sludge stabilization. Dosing ratio increased from 0.5% to 1%, the digestion rates of TOC and MLSS were increased by 10.3% and 8.4%, respectively. When the dosing ratio increased from 1% to 5%, On the 13th day, the improvement of TOC and MLSS removal rates were not significant. But for the reactor with higher initial dosing ratio, removal rates and the removal speeds of TOC and MLSS were higher, the main reason was that with a higher initial dosing ratio, the dominant microorganisms would be able to reach a larger number soon. However, because of the type and structure of dominant microorganisms were unchanged, the final removal effects of TOC and MLSS were not very different with the passage of time (Park et al., 2014). The dosing ratio of the dominant strain in the actual project could not be too low, otherwise, it was difficult to form the dominant population, and could not enhance the stabilization effect of residual sludge. Thinking about the fungus cultivation costs, suitable dosing ratio was 1% in the practical engineering. The digestion rates of TOC and MLSS in aerobic digestion reactors under different DO concentrations were shown in Fig. 6.
209
When DO concentrations were 2 mg/L, 4 mg/L, 6 mg/L and 7 mg/ L, respectively, On the 13th day, TOC digestion rates were 43.3%, 50.6%, 51.6%, 52.5% (Fig. 6E), and MLSS removal rates were 33.9%, 41.8%, 43.78%, 45.0% (Fig. 6F), respectively. With the increase of DO concentration in the reactor, TOC and MLSS removal rates were improved to some extent. When DO was 4 mg/L, TOC and MLSS removal rate reached a relatively ideal state, DO increased further, the sludge digestion effect was not improved obviously in this study. DO was one of the key factors affecting sludge aerobic digestion efficiency, meanwhile, the aeration cost was the main part of aerobic digestion costs (Kim et al., 2001). In practical engineering, increasing the amount of aeration could improve aerobic digestion efficiency and shorten aerobic digestion time, but it would increase operating cost at the same time. While reducing the amount of aeration, aerobic digestion efficiency would be reduced, the required aerobic digestion time would become longer, there might not reach the corresponding treatment effects (Barbusin´ski and Filipek, 2003; Zhou, 2003). Choosing a suitable aeration quantity was very important to the operating costs of aerobic digestion. Effects of different microbial combinations on aerobic digestion efficiency of sludge were studied by adding 2% (volume ratio, bacteria/excess sludge) strains to the reactor, and the pH was controlled to pH 6.5, DO was controlled to 3–5 mg/L, respectively. Results showed that TOC digestion rates of the different microbial combinations including: (1) strain Bacillus sp. L19, (2) strains Bacillus sp. L19 and Pseudomonas sp. L3, (3) strains Bacillus sp. L19 and Acinetobacter sp. L16, (4) strains Bacillus sp. L19 and Acinetobacter sp. L16 and Pseudomonas sp. L3 were 57.8%, 59.0%, 49.8%, 53.2%, respectively (Fig. 6G), and the MLSS elimination rates of the different microbial combinations were 46.2%, 49.2%, 39.6%, 44.1%, respectively (Fig. 6H) on the 13th day. Strain Pseudomonas sp. L3 and strain Bacillus sp. L19 were combined dosing, sludge stabilization effect was the best, on the 6th day, TOC digestion rate reached to 39%, and the sludge stabilization process was almost completed at this time, which could be applied for the production practice of excess sludge stabilization. When the strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 were combined dosing, sludge stabilization effect was poor, the possible reason was that the existence of mutual competition for the living environment between strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19. The enhanced effects of different dominant microorganisms on sludge aerobic digestion were different. By combining different kinds of microorganisms, microbes might cooperate with each other to obtain a better microbial enhancement effect, however, it was possible to compete with each other to reduce the microbial enhancement effect (Tyagi and Lo, 2012). At present, many researchers had used different combinations of microorganisms to obtain better effect, and enhancement of sludge aerobic digestion by combination of different microorganisms was a feasible method for the residual sludge stabilization (Jin et al., 2016a, b; Liu et al., 2010).
4. Conclusions 3 dominant strains named as Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 were obtained. Compared with the control, addition of the strains mentioned above in the reactors of sludge aerobic digestion respectively, sludge stabilization time were advanced by 4 days, 3 days and 4 days, respectively. When the strains Pseudomonas sp. L3 and Bacillus sp. L19 were combined dosing with the dosage of 2%, under the condition of pH 6.5, DO 3– 5 mg/L, sludge stabilization effect was the best, and the digestion rates of TOC and MLSS were 59% and 49% respectively, in the 13th days.
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Acknowledgements This work was supported by Science and Technology Industrialization Project (Grant No. 15JF021) and the Project of Science and Technology Benefit Plan (Grant No. 2012GS610203). References Abe, N., Tang, Y.Q., Iwamura, M., Ohta, H., Morimura, S., 2011. Development of an efficient process for the treatment of residual sludge discharged from an anaerobic digester in a sewage treatment plant. Bioresour. Technol. 102, 7641– 7644. Anjum, M., Al-Makishah, N., Barakat, M.A., 2016. Wastewater sludge stabilization using pre-treatment methods. Process Saf. Environ. 102, 615–632. Bahar, S., Ciggin, A.S., 2016. A simple kinetic modeling approach for aerobic stabilization of real waste activated sludge. Chem. Eng. J. 303, 194–201. Barbusin´ski, K., Filipek, K., 2003. Aerobic sludge digestion in the presence of hydrogen peroxide and Fenton’s reagent. Pol. J. Environ. Stud. 12 (1), 35–40. Bi, D., Guo, X., Chen, D., 2013. Phosphorus release mechanisms during digestion of EBPR sludge under anaerobic, anoxic and aerobic conditions. Water Sci. Technol. 67 (9), 1953–1959. Bomio, M., Sonnleitner, B., Fiechter, A., 1989. Growth and biocatalytic activities of aerobic thermophilic populations in sewage sludge. Appl. Microbiol. Biotechnol. 32, 356–362. Cai, B., Han, Y., Liu, B., 2003. Isolation and characterization of an atrazine-degrading bacterium from industrialwastewater in China. Lett. Appl. Microbiol. 36, 272– 276. Du, H.X., Li, F.S., 2017. Characteristics of dissolved organic matter formed in aerobic and anaerobic digestion of excess activated sludge. Chemosphere 168, 1022– 1031. Fall, C., Rogel-Dorantes, J.A., Millán-Lagunas, E.L., Martínez-García, C.G., SilvaHernández, B.C., Silva-Trejo, F.S., 2014. Modeling and parameter estimation of two-phase endogenous respirograms and COD measurements during aerobic digestion of biological sludge. Bioresour. Technol. 173, 291–300. Ichinari, T., Ohtsubo, A., Ozawa, T., Hasegawa, K., Teduka, K., 2008. Wastewater treatment performance and sludge reduction properties of a household wastewater treatment system combined with an aerobic sludge digestion unit. Process Biochem. 43 (7), 722–728. Jang, H.M., Cho, H.U., Park, S.K., Ha, J.H., Park, J.M., 2014. Influence of thermophilic aerobic digestion as a sludge pre-treatment and solids retention time of mesophilic anaerobic digestion on the methane production, sludge digestion and microbial communities in a sequential digestion process. Water Res. 48, 1– 14. Jin, N.B., Jin, B., Zhu, N.W., Yuan, H.P., Ruan, J.B., 2015. Disinhibition of excessive volatile fatty acids to improve the efficiency of autothermal thermophilic aerobic sludge digestion by chemical approach. Bioresour. Technol. 175, 120– 127. Jin, N.B., Shou, Z.Q., Yuan, H.P., Lou, Z.Y., Zhu, N.W., 2016a. Selective simplification and reinforcement of microbial community in autothermal thermophilic aerobic digestion to enhancing stabilization process of sewage sludge by conditioning with ferric nitrate. Bioresour. Technol. 204, 106–113. Jin, N.B., Li, W.H., Shou, Z.Q., Yuan, H.P., Lou, Z.Y., Zhu, N.W., Cai, C., 2016b. Comparison of effects of ferric nitrate additions in thermophilic, mesophilic and psychrophilic aerobic digestion for sewage sludge. J. Taiwan Inst. Chem. E 67, 346–354. Kim, Y.K., Eom, Y.S., Oh, B.K., Lee, W.H., Choi, J.W., 2001. Application of thermophilic aerobic digestion process to industrial waste activated sludge treatment. J. Microbiol. Biotechnol. 4, 570–576.
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Strengthen effects of dominant strains on aerobic digestion and stabilization of the residual sludge Yongjun Liu ⇑, Min Gao, Aining Zhang, Zhe Liu Key Lab of Northwest Water Resource, Ecology and Environment, Ministry of Education, Xi’an University of Architecture and Technology, Xi’an 710055, PR China
h i g h l i g h t s Strains which have obvious tolerance to the dissolved oxygen were isolated. Addition of dominant strains has a significant role on the sludge reduction. The stabilization time of sludge was shortened considerably. Accumulation of nitrogen and phosphorus in sludge supernatant was increased.
a r t i c l e
i n f o
Article history: Received 4 January 2017 Received in revised form 4 March 2017 Accepted 8 March 2017 Available online 11 March 2017 Keywords: Aerobic digestion of sludge Stabilization Dominant strains Strengthen
a b s t r a c t In order to strengthen the aerobic digestion of residual sludge, shorten the time of sludge stabilization and further reduce operating costs, 3 dominant strains identified as Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 were isolated from long-term aerobic digestion sludge. Results showed that the sludge stabilization time were reduced by 3–4 days compared with the control when the dominant strains were added to the process of sludge aerobic digestion. The addition of dominant strains accelerated the accumulation of TOC, nitrate nitrogen and ammonia nitrogen in the digestive solution at different levels, and it was beneficial to the dissolution of phosphorus. Controlling DO 3–5 mg/L, pH 6.5, the strains of Pseudomonas sp. L3 and Bacillus sp. L19 were combined dosing with the dosage of 2% in the process of sludge aerobic digestion, compared with the control, digestion rates of TOC and MLSS were increased about 19% and 16%, respectively. Ó 2017 Elsevier Ltd. All rights reserved.
1. Introduction As the activated sludge processes were widely used in various industries of sewage treatment, a large number of the residual sludges were also produced (Du and Li, 2017). The residual sludge contains lots of refractory organic matters, heavy metals and salts, as well as pathogenic microorganism, parasite eggs and other toxic and harmful substances, if the residual sludges were not stabilized properly, it would cause new pollution to the environment, and the problem of sludge pollution has posed serious challenge to human living environment (Zhu et al., 2017; Jin et al., 2016a,b). At present, the main disposal methods for residual sludge were sanitary landfill, sludge incineration, sludge compost and sludge for agricultural use, etc (Martínez-García et al., 2016). However, due to the site restrictions, high infrastructure and operating man-
⇑ Corresponding author. E-mail address: [email protected] (Y.J. Liu). http://dx.doi.org/10.1016/j.biortech.2017.03.060 0960-8524/Ó 2017 Elsevier Ltd. All rights reserved.
agement costs and secondary pollution, sludge disposal problem has not been fundamentally resolved (Wang et al., 2016). Reduction and stabilization were the main targets for sludge treatment. Sludge digestion including aerobic digestion and anaerobic digestion, its essence was that the microbial cells and organic matters were degraded into small molecular matters by the processes of endogenous respiration and microbial biodegradation, and then the small molecular matters were used again as substrates by the microorganisms to achieve the purpose of sludge stabilization and reduction (Zhang et al., 2016; Rafieenia et al., 2017). The sludge volume after digestion could be reduced by 30%-50%, and the dewatering performance of sludge was greatly improved, it was one of the feasible methods for sludge reduction and stabilization (Fall et al., 2014; Bahar and Ciggin, 2016). But the existed problems were that aerobic digestion required long time aeration, the operation costs were large, and infrastructure costs for anaerobic digestion were high, at the same time, operation management was complex (Anjum et al., 2016). Therefore, it was
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imperative to find an efficient and cost-effective sludge treatment technology. The knowledge of microbial population told us that in a given system, only the most suitable microorganisms could adapt to the environment and survive (Jin et al., 2016a, b). Based on this point, researchers in the world screened many efficient strains from the environment and added it to the specific system, the efficiency of metabolic decomposition of pollutants were enhanced by increasing the dominant microbial concentration and optimizing the microbial population structures (Liu et al., 2011; Manaia et al., 2003). The use of microbial enhanced technology could improve the reduction and stabilization of sludge, which provided a new way for sludge treatment (Li et al., 2009; Kim et al., 2002). The aerobic digestion processes of residual sludge were simulated by long time aeration in this study, the dominant microorganisms were screened and used to study on the effects of sludge stabilization in order to shorten the aerobic digestion time, and further reducing the operation costs, which has very important theoretical and practical significance for the efficient treatment of residual sludge.
2. Materials and methods 2.1. Experiment equipment and the residual sludge Four sets of experimental devices were used in this study, each device was 45 cm high, and the inner diameter was 10 cm (Fig. 1). Aeration header was arranged at the bottom of the device. The sampling ports were set at 5 cm, 20 cm and 35 cm distance from the bottom of the device, respectively, sampling from 3 sampling
1
ports simultaneously, and then the samples were mixed for detection. The residual sludge used in this study were taken from the second sedimentation tank of the Fourth sewage treatment plant of Xi’an city, China, the properties of sludge were shown in Table 1. 2.2. Domestication and isolation of superior strains for sludge aerobic digestion 1 mL Excess sludge taken from aerobic digestion reactor were diluted to 101, 102, 103, 104 and 105 five concentration gradients in the bacteria-free environment, and the dilution plate method was used for separation of the microorganisms from different concentration gradients, cultural conditions were as follows: Lucia-Bertani solid culture medium (peptone 10 g, yeast extract 10 g, NaCl 5 g, agar 18 g, distilled water 1000 mL), pH 6.8–7.2, 30 °C, for 24 h. The single strain which has different colony morphology and microscopic structure was picked out respectively and saved. Method of acclimation was divided into 5 steps. The first step, the isolated strains mentioned above were cultured with LB liquid culture medium (no agar) at 30 °C, for 24 h, respectively. 30 mL culture liquid of different strains were taken, respectively, then centrifuged under 4000 rpm for 20 min. Remove supernatant, and the sediments were rinsed two times with pure water. The obtained different single strain cells were mixed, then added to 100 mL residual sludge, and acclimated under aerobic condition for 10 days. The second step, 30 mL residual sludge after acclimation was taken, then added to 70 mL fresh excess sludge, and acclimated under aerobic condition for 10 days. The third to the fifth step, repeat the second step. 1 mL sludge taken from the last acclimation step were diluted to 101, 102, 103, 104 and 105 five concentration gradients in the bacteria-free environment, then cultured in LB solid medium at 30 °C, for 24 h, the single strain which has different colony morphology and microscopic structure was picked out and named as L1-L25, respectively. 2.3. Screening of dominant strains for aerobic digestion The isolated strains named as L1-L25 mentioned above were cultured in the liquid LB medium at 30 °C, for 24 h, respectively. 30 mL culture liquid of different strains were taken, then centrifuged under 4000 rpm for 20 min. Remove supernatant, and the sediments were rinsed two times with pure water. The obtained different single strain cells were added to 100 mL new residual sludge, respectively, then cultured under aerobic condition. TOC degradation rate of each test group was measured 6 days later, the dominant strains for aerobic digestion were screened out.
2
6 3
2.4. Identification of the dominant strains
4
5
Fig. 1. Diagram of the aerobic digestion reactor. 1: Main device of the reactor, 2: Sampling holes; 3: Aeration plate, 4: Gas flow meter, 5: Aeration pump, 6: Temperature controller.
16s rDNA fragments of the screened dominant strains of aerobic digestion were amplified with the primers 27F (50 -AGAGTTT GATCMTGGCTCAG-30 ) and 1492R (50 -CGGYTACCTTGTTACGACTT30 ). PCR reaction was performed according to Cai et al. (2003). The PCR products were cloned into the pMD 18-T vector (TaKaRa Biotechnology Company) following the manufacturer’s instruction.
Table 1 Properties of the residual sludge in this study. The residual sludge flocs
The sludge supernatant
TOC (mg/L)
MLSS (mg/L)
SVI (mL/g)
pH
TOC (mg/L)
NO 3 (mg/L)
NH+4 (mg/L)
TP (mg/L)
3535.3 ± 270.1
9770.4 ± 328.6
95.8 ± 6.6
6.8 ± 0.2
19.3 ± 3.2
24.6 ± 2.2
0.5 ± 0.1
26.5 ± 2.3
Notes: TOC: Total organic carbon; MLSS: Mixed liquid suspended solid; SVI: Sludge volume index; TP: Total phosphorus.
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16s rDNA fragments were sequenced by Shanghai Biological Engineering Technology and Services Co., Ltd. BLAST program was used to search the similar sequences of 16S rRNA database, and the phylogenetic trees were constructed with the software ClustalX Version1.81 by using the neighbor-joining method. The 16S rRNA gene fragments of strains named as L3, L16 and L19 were deposited in the GenBank data library with the ID no. 1999166, 1999172 and 1999175, respectively. 2.5. Test of enhanced effect of dominant strains on residual sludge stabilization Dominant strains named as L3, L16, L19 were cultured in the LB medium at 30 °C for 24 h, respectively. 200 mL culture liquid of different strains were taken, then centrifuged under 4000 rpm for 20 min. Remove supernatant, and the sediments were rinsed two times with pure water. The obtained different microbial cells were suspended with pure water to about 10 101°Cfu/mL, the cell suspension was added to the reactor with 2 L fresh residual sludge at the volume ratio (bacteria/residual sludge) of 1%, with no addition of bacteria cells as control, DO in reactors were controlled to 3– 5 mg/L. Sampling every 24 h, TOC, MLSS, SVI of the residual sludge flocs and TOC, nitrate nitrogen, ammonia nitrogen, TP in the sludge supernatant were measured. The standard for the sludge stabilization of aerobic digestion was that the TOC degradation rate attained to 40%. 2.6. Detection methods In this study, TOC, MLSS and SVI of the residual sludge flocs, the total phosphorus, nitrate nitrogen and ammonia nitrogen in the sludge supernatant were measured by using the standard determination methods from ‘‘water and wastewater monitoring analysis method” (SEPAC, 2002a). 3. Results and discussion 3.1. Screening and identification of dominant strains for sludge aerobic digestion In the course of microbial growth, some microorganisms could utilize the substrates of the second time which formed by the death cell to grow, this growth pattern was called invisible growth (Yan et al., 2008), and the dissolution of dead cells was the key step in this course. Therefore, aerobic digestion of excess sludge might exist some flora which conducive to the dissolution of microbial cells, and the sludge stabilization by aerobic digestion was a kind of excess sludge treatment technology based on invisible growth (Bomio et al., 1989; Jang et al., 2014). 25 dominant strains with good invisible growth were isolated by domestication for many times in this study, the strains were named as L1-L25, respectively. After further sludge digestion test, TOC aerobic degradation rates (average value) of residual sludge by single strain L3, L16 and L19 were 31.2%, 28.8% and 32.7% respectively, within 6 days were used for further study. The TOC degradation rates for other strains were relatively lower from 16.1% to 23.6% within 6 days and not involved in this study. 16s rRNA gene fragments of strains L3, L16 and L19 were obtained respectively, by PCR amplification, and then submitted for sequencing. The blast search of the sequences indicated that the strain L3 was closely related to the species in genus of Pseudomonas, it also exhibited the highest similarity (99.35%) to Pseudomonas sp. and thus tentatively classified strain L3 as Pseudomonas sp. L3 (ID1999166), Fig. 2(A). Strain L16 was matched at 97.68% to Acinetobacter sp. Thus, identified as Acinetobacter sp.
L16 (ID1999172), Fig. 2(B). Strain L19 was matched at 98.39% to Bacillus sp., which was identified as Bacillus sp. L19 (ID1999175), Fig. 2(C). 3.2. Variations of TOC, MLSS and pH during the process of long term residual sludge aerobic digestion The variations of TOC and MLSS in the process of aerobic digestion were related to the length of aeration time, and the changes of pH directly affected the activity of microorganisms, further affected the aerobic digestion efficiency. Determination of digestion time was very important for saving energy and reducing operating costs. As shown in Fig. 3, after 80 days aerobic digestion, MLSS were decreased from 8900 mg/L to 3200 mg/L, and reduced by 64.0%. TOC were decreased from 3800 mg/L to 1250 mg/L, and reduced by 67.1%. During the aerobic digestion, MLSS and TOC were decreased faster in the initial 20 days, and began to enter the plateau after 20 days. When the digestion time was estimated by empirical method, usually activated sludge were 15–20 days, mixed sludge were 20–25 days (Martins et al., 2003), which were basically consistent with the experimental results of this study. During aerobic digestion of sludge, pH gradually decreased from pH 6.67 to pH 4.46. On the 70th day, the pH reached a minimum pH 3.08, and then increased gradually from pH 3.08 to pH 4.46 on the 80th day (Fig. 3). In the process of aerobic digestion, there was a significant negative correlation between pH and nitrate content, the decrease of pH was mainly due to nitrification (Jang et al., 2014). 70 days later, pH began to rise in this study, indicating a gradual decrease of nitrate content. Studies have shown that in aerobic digestion process, with the digestion of microbial cells, the intracellular substances were continuously dissolved, and the content of total phosphorus was also accumulated in the solution (Bi et al., 2013). Under the condition of large amount of nitrate and total phosphorus in sludge, the denitrifying phosphorus removal reaction with nitrate as electron acceptor could be induced, denitrifying phosphorus removal decreased nitrate content and pH increased gradually (Jin et al., 2015). 3.3. Enhanced effects of dominant strains on aerobic digestion and stabilization of the residual sludge 3.3.1. The time of residual sludge stabilization completion and the enhancement effects of the dominant strains According to the pollutant discharge standard of the municipal wastewater treatment plant, the degradation rate of organic matter in sludge aerobic digestion reached 40%, which was regarded as the stabilization of excess sludge had completed (SEPAC, 2002b). Results of the enhancement effects by the dominant strains on aerobic digestion of the residual sludge were shown as Table 2 and Fig. 4. Variations of TOC during sludge aerobic digestion were shown as Table 2, the control group had reached the sludge TOC 40% degradation rate on the 13th day, the groups of addition strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 had completed the sludge stabilization on the 9th day, 10th day and 9th day, respectively. Compared with the control group, completion time of sludge stabilization were advanced for 4d, 3d and 4d, respectively, indicated that added of the dominant strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 could obviously improve the efficiency of sludge stabilization (Fig. 4A). Variations of MLSS values during sludge aerobic digestion by adding dominant strains were shown as Table 2. Because of the completion time of sludge stabilization for the control group was on the 13th day in this study, the sludge aerobic digestion time was determined for 2 weeks. The degradation rate of MLSS after
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Fig. 2. Phylogenetic trees based on 16S rDNA sequences of strains named as L3, L16, L19 and their related species. A: Strain L3 (ID1999166) closely related to the species in genus of Pseudomonas and thus identified as Pseudomonas sp. L3; B: Strain L16 (ID1999172) closely related to the species in genus of Acinetobacter and thus identified as Acinetobacter sp. L16; C: Strain L19 (ID1999175) closely related to the species in genus of Bacillus and thus identified as Bacillus sp. L19.
bacterial addition was increased with the passage of time, and higher evidently than that the control at the end of aerobic digestion. When the digestion reactions were carried out to 13th day, the MLSS degradation rates of adding strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 were 40.1%, 36.8% and 41.3%, respectively, compared with the control (33.1%) increased by 7.0%, 3.7% and 8.2%, respectively, indicated that the dominant microorganism could promote the residual sludge reduction (Fig. 4B). The results of SVI variations during sludge stabilization by adding strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19, respectively were showed as Table 2. When all test groups reached the sludge stabilization state (on the 13th day), compared with the control group, SVI of Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 strengthen groups decreased
by 14.1%, 9.0% and 11.1%, respectively, indicated that the addition of the dominant strains were beneficial to sludge dewatering treatment. The operation cost of residual sludge treatment could often account for 50%–60% of sewage treatment plant operation costs, and the energy expenditure of aeration in aerobic digestion accounted for most of the aerobic digestion costs (Abe et al., 2011). In the process of sludge aerobic digestion, adding dominant microorganisms could significantly improve sludge digestion efficiency and shorten sludge stabilization time in this study. Strengthen effects of strains Pseudomonas sp. L3 and Bacillus sp. L19 could shorten the aeration time by 31% in theory, which could reduce the operating costs of sewage treatment plant effectively. Adding dominant strains could shorten the time required for
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9000
decomposition and utilization by other microorganisms (Du and Li, 2017; Yan et al., 2008).
7
MLSS
7000
TOC
6
pH
6000
5
3.3.2. Variations of dissolved substances during the process of sludge stabilization Variations of dissolved substances in the supernatant of the different test groups during the sludge stabilization process were shown in Fig. 5. When the digestive reaction proceeded to 13th day, TOC in the supernatant of different reactors which were added the dominant strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 respectively, were increased to 60 mg/L, 42 mg/L and 53.4 mg/L, respectively. Compared with the control (40 mg/L), TOC cumulative volume in strengthen reactors increased by 50%, 5% and 33.5%, respectively (Fig. 5A). Nitrate nitrogen, ammonia nitrogen and dissolved phosphorus all had different degrees of accumulation in the digestive juice. On the 13th day, the amount of nitrate nitrogen in the supernatants of strengthen groups by adding strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 increased to 105.6 mg/L, 81.8 mg/L and 93.0 mg/L, respectively (Fig. 5B), compared with the control (69.2 mg/L), increased by 52.6%, 18.2% and 34.4%, respectively. The amount of ammonia nitrogen increased to 112.7 mg/L, 99.64 mg/L and
5000 4 4000 3000
pH
MLSS and TOC (mg/L)
8000
3
2000 2
1000 0
1 0
5 10 15 20 25 30 35 40 45 50 55 60 65 70 75 80 Aerobic digestion time (d)
Fig. 3. Variations of the mixed liquid suspended solid (MLSS), total organic carbon (TOC) and pH during long term aerobic digestion of the residual sludge at 25 °C.
sludge stabilization, the main reason might be that the dominant microorganisms could secrete more extracellular enzymes, these hydrolases could accelerate the hydrolysis of the dead cells, improve the efficiency of sludge hydrolysis, promote macromolecular organic matter degradation into small molecules, further
Table 2 Variations of the TOC, MLSS and SVI during the process of the residual sludge aerobic digestion by adding dominant strains.
TOC (mg/L)
Time (day)
MLSS (mg/L)
SVI (mL/g)
control
L3
L16
L19
control
L3
L16
L19
control
L3
L16
L19
0
3794.3
3796.8
3798.5
3794.2
8900.2
8896.5
8892.8
8898.0
95.3
94.8
95.1
94.6
1
3705.1
3699.3
3701.2
3696.6
8677.5
8659.7
8658.9
8657.9
96.2
98.2
100.6
97.2
2
3610.3
3572.1
3496.2
3534.7
8321.5
8205.8
8232.5
8232.5
100.6
104.1
102.5
100.1
3
3477.6
3344.3
3306.4
3325.0
8099.6
7786.5
7832.8
7743.3
103.4
106.3
106.3
105.4
4
3268.5
3154.3
3116.3
3135.4
7787.5
7342.5
7565.5
7253.5
105.1
108.2
110.7
109.7
5
3161.6
2850.0
2888.3
2812.7
7431.5
6897.5
7342.5
7075.5
106.3
105.7
110.5
104.8
6
3040.0
2625.6
2698.0
2565.0
7120.0
6452.5
7031.3
6897.5
110.0
103.6
110.3
100.3
7
2850.0
2470.0
2621.8
2432.4
6897.5
6230.0
6675.7
6452.5
110.3
95.0
100.2
90.5
8
2622.6
2321.8
2470.7
2394.5
6586.2
6007.5
6230.0
6006.5
110.6
90.8
90.0
85.2
9
2546.3
2280.5
2394.6
2276.8
6408.2
5785.6
6007.5
5783.4
100.7
83.7
85.6
81.7
10
2470.0
2223.8
2279.3
2192.6
6230.0
5696.7
5874.6
5740.5
90.5
79.4
80.5
76.3
11
2432.2
2166.7
2204.3
2090.7
6096.5
5607.4
5829.5
5607.8
86.4
71.7
74.4
72.4
12
2394.5
2090.0
2158.4
2054.3
6052.5
5518.2
5785.4
5473.5
80.0
67.6
70.6
69.2
13
2295.2
2052.0
2128.2
2014.6
5951.6
5329.3
5613.6
5224.0
75.7
65.0
68.8
67.2
14
2268.6
2013.3
2105.2
1968.4
5918.5
5286.6
5624.8
5188.7
73.7
63.4
66.7
65.8
Notes: (1) TOC: Total organic carbon; MLSS: Mixed liquid suspended solid; SVI: Sludge volume index. (2) Conditions of the aerobic digestion: inoculation amount of the dominant strains was 1% (volume ratio); dissolved oxygen 3–5 mg/L, 25 °C and natural pH. (3) Control: indicates no dominant strain addition; L3: means inoculation strain Pseudomonas sp. L3; L16: stands for strain Acinetobacter sp. L16; L19: stands for strain Bacillus sp. L19. (4) The values shown in the table were the average values of the 3 parallel samples.
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50
60
TOC digestion rate (%)
MLSS elimination rate (%)
45
50
A
40 30
control L3 L16 L19
20 10
40
B
35 30 25
control L3 L16 L19
20
15 10 5
0
0 0
1
2
3
4
5
6
7
8
9
10 11 12 13 14
0
1
2
3
Aerobic digestion time (d)
4
5
6
7
8
9
10 11 12 13 14
Aerobic digestion time (d)
Fig. 4. Effects of the different dominant strains on the mixed liquid suspended solid (MLSS) and total organic carbon (TOC) in the process of the residual sludge aerobic digestion under the condition of dissolved oxygen 3–5 mg/L, 25 °C and natural pH. A: Effects of the different dominant strains on the TOC. Control: indicates no dominant strain addition; L3: means inoculation strain Pseudomonas sp. L3 at the rate of 1% (volume ratio), the same below; L16: stands for strain Acinetobacter sp. L16; L19: stands for strain Bacillus sp. L19. B: Effects of the different dominant strains on the MLSS.
120
A
60
NO3 --N in supernatant (mg/L)
TOC in supernatant (mg/L)
70
50 40 30
20 control
10
L3
L16
L19
0
B
100 80 60 40
control
20
L16
L19
0 0
2
4
6
8
10
12
14
0
2
4
6
8
10
12
14
Aerobic digestion time (d)
Aerobic digestion time (d)
80
140 120
70
C
TP in supernatant (mg/L)
NH 3 -N in supernatant (mg/L)
L3
100
80 60 40 control
20
L3
L16
L19
D
60
50 40
30 20 control
10
L3
L16
L19
8
10
12
0
0
0
2
4
6
8
10
12
14
Aerobic digestion time (d)
0
2
4
6
14
Aerobic digestion time (d)
Fig. 5. Variations of dissolved substances in the sludge supernatant during the process of enhanced sludge stabilization under the condition of natural pH and 25 °C. A: Total organic carbon in the sludge supernatant. Control indicates no dominant strain addition, L3 means inoculation strain Pseudomonas sp. L3 at the rate of 1% (volume ratio), the same below, L16 stands for strain Acinetobacter sp. L16, L19 stands for strain Bacillus sp. L19. B: Nitrate nitrogen in the sludge supernatant. C: Ammonia nitrogen in the sludge supernatant. D: Total phosphorus in the sludge supernatant.
118.4 mg/L, respectively (Fig. 5C), compared with the control (88.5 mg/L), increased by 27.3%, 12.5% and 33.8%, respectively. The amount of dissolved phosphorus increased to 74.8 mg/L, 69.4 mg/L and 70.6 mg/L, respectively (Fig. 5D), compared with the control (58.5 mg/L), increased by 27.9%, 18.6% and 20.7%, respectively. The accumulation of the TOC, nitrate nitrogen, ammonia nitrogen and dissolved phosphorus in the bacterial strengthen
group were all higher than that the control. The contents of nitrate and total phosphorus in digestive juice were positively correlated with the degradation rate of MLSS, indicated that the accumulation of nitrogen and phosphorus in digestive juice were mainly due to the degradation of MLSS. The source of nitrogenous substances during aerobic digestion of sludge was mainly the ammonia nitrogen which was produced
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60
60
under natural pH
pH 6.5
50 40 30
A
20
50
MLSS elimination rate (%)
TOC digestion rate (%)
70
10
under natural pH
40 30 20
B
10
0
0 0
2
4
6
8
10
12
0
14
2
4
Aerobic digestion time (d)
70 1%
2%
5%
MLSS elimination rate (%)
TOC digestion rate (%)
0.50%
50 40 30
C
20 10 0
8
10
12
14
0.50%
50
1%
2%
5%
40
30 20
D
10 0
0
2
4
6
8
10
12
14
0
2
4
Aerobic digestion time (d)
6
8
10
12
14
Aerobic digestion time (d)
60
60
DO 4mg/L DO 7mg/L
MLSS elimination rate (%)
DO 2mg/L DO 6mg/L
50 TOC digestion rate (%)
6
Aerobic digestion time (d)
60
60
40 30 20
E
10
0
DO 2mg/L DO 6mg/L
50
DO 4mg/L DO 7mg/L
40 30 20
F
10
0 0
2
4
6
8
10
12
14
0
Aerobic digestion time (d)
50
MLSS elimination rate (%)
L19 L19+L3 L19+L16 L19+L16+L3
60
40 30
G
20
2
4
10
6
8
10
12
14
Aerobic digestion time (d)
60
70
TOC digestion rate (%)
pH 6.5
L19 L19+L3 L19+L16 L19+L16+L3
50 40 30 20
H
10 0
0
0
2
4
6
8
10
Aerobic digestion time (d)
12
14
0
2
4
6
8
10
12
14
Aerobic digestion time (d)
Fig. 6. Analysis of influencing factors of sludge aerobic digestion. A: Effect of pH on the total organic carbon (TOC); B: Effect of pH on the mixed liquid suspended solid (MLSS); C: Effect of inoculation amount (volume ratio) on the TOC; D: Effect of inoculation amount on the MLSS; E: Effect of dissolved oxygen (DO) on the TOC; F: Effect of DO on the MLSS; G: Effect of different dominant strains combination on the TOC, L3, L16, L19 stand for strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19, respectively; H: Effect of different strains combination on the MLSS.
Y.J. Liu et al. / Bioresource Technology 235 (2017) 202–210
by the sludge hydrolysis and ammonization continuously, and then converted to nitrate nitrogen by nitrification. Nitrate nitrogen and ammonia nitrogen were accumulated from beginning to end in this study, the results were differ from those obtained by other researchers (Li et al., 2014; Lee et al., 2004) whose results showed that ammonia accumulated to a certain amount would no longer accumulate. The main reason was that the sludge digestion time was shorter in this study, and the number of nitrifying bacteria was less, which led to a slower conversion rate from ammonia nitrogen to nitrate nitrogen, and the ammonia nitrogen had been always accumulated in the digestive juice. The dissolution rate of phosphorus in sludge depended on the degree of hydrolysis of sludge microbial cells (Bi et al., 2013). The dissolution rate of phosphorus in the bacterial addition group was higher than that in the blank control in this study, indicated that the added microbes were favorable for sludge hydrolysis. However, the trend of phosphorus change curves were similar, indicated that the added microorganisms did not change the dissolution mechanism of phosphorus in the process of sludge aerobic digestion. 3.4. Analysis of influencing factors in the process of enhanced sludge aerobic digestion During short-term aerobic digestion (two weeks in this study), pH decreased gradually. 1 mol/L NaOH was added to the digestion reactor every 12 h to regulate pH 6.5, dissolved oxygen (DO) was controlled to 3-5 mg/L, and strain Bacillus sp. L19 was added as the test strain. On the 13th day, degradation rates of TOC (Fig. 6A) and MLSS (Fig. 6B) were 56.3% and 48.4% respectively, when the pH value was adjusted to pH6.5. Degradation rates of TOC and MLSS were 48.7% and 40.8%, respectively, in the reactor without adjusting pH, indicated that controlling pH value during sludge aerobic digestion had a significant effect on sludge stabilization. In fact, sludge aerobic digestion process was the role of microbial endogenous respiration, and the appropriate pH was favorable for microorganism to maintain its activity. When pH value was too low, it would make the changes of microbial cell permeability, thus changing the microbial enzyme catalytic capacity, and ultimately affect the microbial mass transport and metabolism process (Ichinari et al., 2008; Yan et al., 2008). DO was controlled to 3-5 mg/L, pH value was controlled to pH 6.5, and strain Bacillus sp. L19 was added as the test strain. On the 13th day, when the dosing ratios of test strain Bacillus sp. L19 were 0.5%, 1%, 2% and 5% (volume ratio), respectively, TOC degradation rates were 45.3%, 55.6%, 58.2%, 59.1% (Fig. 6C), and MLSS removal rates were 33.9%, 42.3%, 46.4%, 49.8% (Fig. 6D), respectively. The results indicated that the higher dosing ratio could improve the efficiency of sludge stabilization. Dosing ratio increased from 0.5% to 1%, the digestion rates of TOC and MLSS were increased by 10.3% and 8.4%, respectively. When the dosing ratio increased from 1% to 5%, On the 13th day, the improvement of TOC and MLSS removal rates were not significant. But for the reactor with higher initial dosing ratio, removal rates and the removal speeds of TOC and MLSS were higher, the main reason was that with a higher initial dosing ratio, the dominant microorganisms would be able to reach a larger number soon. However, because of the type and structure of dominant microorganisms were unchanged, the final removal effects of TOC and MLSS were not very different with the passage of time (Park et al., 2014). The dosing ratio of the dominant strain in the actual project could not be too low, otherwise, it was difficult to form the dominant population, and could not enhance the stabilization effect of residual sludge. Thinking about the fungus cultivation costs, suitable dosing ratio was 1% in the practical engineering. The digestion rates of TOC and MLSS in aerobic digestion reactors under different DO concentrations were shown in Fig. 6.
209
When DO concentrations were 2 mg/L, 4 mg/L, 6 mg/L and 7 mg/ L, respectively, On the 13th day, TOC digestion rates were 43.3%, 50.6%, 51.6%, 52.5% (Fig. 6E), and MLSS removal rates were 33.9%, 41.8%, 43.78%, 45.0% (Fig. 6F), respectively. With the increase of DO concentration in the reactor, TOC and MLSS removal rates were improved to some extent. When DO was 4 mg/L, TOC and MLSS removal rate reached a relatively ideal state, DO increased further, the sludge digestion effect was not improved obviously in this study. DO was one of the key factors affecting sludge aerobic digestion efficiency, meanwhile, the aeration cost was the main part of aerobic digestion costs (Kim et al., 2001). In practical engineering, increasing the amount of aeration could improve aerobic digestion efficiency and shorten aerobic digestion time, but it would increase operating cost at the same time. While reducing the amount of aeration, aerobic digestion efficiency would be reduced, the required aerobic digestion time would become longer, there might not reach the corresponding treatment effects (Barbusin´ski and Filipek, 2003; Zhou, 2003). Choosing a suitable aeration quantity was very important to the operating costs of aerobic digestion. Effects of different microbial combinations on aerobic digestion efficiency of sludge were studied by adding 2% (volume ratio, bacteria/excess sludge) strains to the reactor, and the pH was controlled to pH 6.5, DO was controlled to 3–5 mg/L, respectively. Results showed that TOC digestion rates of the different microbial combinations including: (1) strain Bacillus sp. L19, (2) strains Bacillus sp. L19 and Pseudomonas sp. L3, (3) strains Bacillus sp. L19 and Acinetobacter sp. L16, (4) strains Bacillus sp. L19 and Acinetobacter sp. L16 and Pseudomonas sp. L3 were 57.8%, 59.0%, 49.8%, 53.2%, respectively (Fig. 6G), and the MLSS elimination rates of the different microbial combinations were 46.2%, 49.2%, 39.6%, 44.1%, respectively (Fig. 6H) on the 13th day. Strain Pseudomonas sp. L3 and strain Bacillus sp. L19 were combined dosing, sludge stabilization effect was the best, on the 6th day, TOC digestion rate reached to 39%, and the sludge stabilization process was almost completed at this time, which could be applied for the production practice of excess sludge stabilization. When the strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 were combined dosing, sludge stabilization effect was poor, the possible reason was that the existence of mutual competition for the living environment between strains Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19. The enhanced effects of different dominant microorganisms on sludge aerobic digestion were different. By combining different kinds of microorganisms, microbes might cooperate with each other to obtain a better microbial enhancement effect, however, it was possible to compete with each other to reduce the microbial enhancement effect (Tyagi and Lo, 2012). At present, many researchers had used different combinations of microorganisms to obtain better effect, and enhancement of sludge aerobic digestion by combination of different microorganisms was a feasible method for the residual sludge stabilization (Jin et al., 2016a, b; Liu et al., 2010).
4. Conclusions 3 dominant strains named as Pseudomonas sp. L3, Acinetobacter sp. L16 and Bacillus sp. L19 were obtained. Compared with the control, addition of the strains mentioned above in the reactors of sludge aerobic digestion respectively, sludge stabilization time were advanced by 4 days, 3 days and 4 days, respectively. When the strains Pseudomonas sp. L3 and Bacillus sp. L19 were combined dosing with the dosage of 2%, under the condition of pH 6.5, DO 3– 5 mg/L, sludge stabilization effect was the best, and the digestion rates of TOC and MLSS were 59% and 49% respectively, in the 13th days.
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