Physical and Chemical Treatment Processes for Leachate

Chapter 2

Physical and Chemical Treatment Processes for Leachate 2.1 TYPICAL PHYSICAL AND CHEMICAL TREATMENT PROCESSES FOR LEACHATE Struvite, so-called magnesium ammonium phosphate (MAP), can effectively remove NH3-N in leachate. However, the cost of pure chemical precipitation is extremely high and it is impossible to be applied in engineering. Therefore, it is possible to recycle the raw material by recycling used struvite with high temperature thermal decomposition. Flocculating settling process, which regards ferric salts or aluminium salts as the flocculants, can remove 50% COD from leachate after biological pretreatment. During the flocculating settling process, optimum ranges of pH for ferric salts and aluminium salts are 4.5
4.8 and 5.0
5.5, respectively. The minimum dosage of flocculants is 250
500 g (Fe or Al)/m3, which is approximate to the theory value of these salt flocculants. The amounts of sludge generated, the adjustment of pH, the high salt concentration in effluents, and the lower removal efficiency of NH3-N are the main problems, and should be considered together. Activated carbon adsorption could be one of the advanced leachate treatment process, which has been used in Europe in the 1970s, and around 50%
60% COD could be removed, and around 80% of COD could be further removed by limestone. The linear relation between amount of activated carbon and COD removal is consistent with 3.0
3.2 mg COD/g activated carbon, when activated carbon dosage is at the range of 800
1200 g/m3. The high cost is a main barrier for the widespread application of activated carbon adsorption process. However, lower effluent COD and AOX are obtained when the combination of activated carbon adsorption with flocculating settling process are used after biological treatment. A chemical oxidizing process could reduce the contaminant directly, through the oxidization and mineralization of the pollutants in leachate. The chemical oxidants used include chlorine, calcium hypochlorite, potassium permanganate, and O3. COD removal efficiency is less than 50% when

Pollution Control Technology for Leachate From Municipal Solid Waste. DOI: https://doi.org/10.1016/B978-0-12-815813-5.00002-4 © 2018 Elsevier Inc. All rights reserved.

31

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Pollution Control Technology for Leachate From Municipal Solid Waste

calcium hypochlorite is chosen as an oxidant. For O3, there is no excess sludge produced despite the removal efficiency of COD is below 50%. Moreover, the effectiveness is poor using O3 to treat the acidity leachate containing massive organic acids that have the ozone resistance property. Hydrogen peroxide is often used to deal with bad smell due to its great performance on eliminating hydrogen sulfide with the ratio 1.5
3 between soluble sulphur and hydrogen peroxide. Mechanical Vapor Compression (MVC) process used in leachate treatment is a physicochemical separation process. The raw leachate is fed into MVC equipment after filtration, and the resulted evaporated leachate is condensed and further treated through a DI ion exchange system to remove ammonia. The noncondensable gases created from evaporation will reach the standard after acid and alkali treatment. MVC has been used for treatment of high inorganic salt and organic wastewater, as well as the production of pure water and the concentration of chemical wastewater, since lower energy consumption, better effluent quality, and more convenient on operation management are found in this process. However, high energy consumption and heavy scaling has refrained MVC from wide application in leachate treatment. Most leachate samples and on-site engineering research are conducted at Shanghai Refuse Landfill, which was constructed in 1985 along the shore of the East China Sea, formed by the sedimentation of silt carried by the Yangtze River. Shanghai Refuse Landfill is constructed as a 5-phase projects. The first 3 phases were conducted in the 1980s and 1990s, with an area of 4 km2, in which 56 landfill cells were formed, as shown in Fig. 2.1. The Phase IV landfill was constructed in late 2000, just for a landfill with 4900 t/d. Phase V was used in 2010, including a refuse landfill, an incineration plant with 3000 and 6000 t/d together, a hazardous wastes landfill, and a dewatered sewage sludge landfill. The landfill leachate was collected by a leachate collecting system under the bottom of the cell and discharged through a separate pipeline.

FIGURE 2.1 Landfill cells distribution at Shanghai Refuse Landfill.

Physical and Chemical Treatment Processes for Leachate Chapter | 2

33

2.2 AMMONIA NITROGEN REMOVAL USING STRUVITE PRECIPITATION COD and NH3-N or TN in leachate are the two predominant pollutants that are most difficult to be removed, using physical chemical or biological means. Addition of lime to leachate can increase pH and effectively strip NH3-N upon injection of air, but a serious secondary pollution may be caused and has been abandoned. Struvite, so-called MAP, can effectively remove NH3-N in leachate, which greatly reduces the microbial toxicity caused by NH3-N to the biochemical treatment section. However, the cost of pure chemical precipitation is extremely high and it is impossible to be applied in engineering. Therefore, it is possible to recycle the raw material by recycling used struvite with high temperature thermal decomposition. Thus, cost of dosing chemicals can be greatly reduced and NH3-N concentration in leachate can be adjusted and controlled directly, swiftly and effectively. The traditional struvite treatment method means dosing a proper amount of magnesium salt and phosphate to remove NH3-N by ammonium magnesium hexahydrate phosphate precipitation when treating high NH3-N concentration wastewater. In thermal cycle struvite treatment process, the formed ammonium magnesium hexahydrate phosphate precipitation is pyrolyzed under certain temperature on basis of conventional method. Amorphous magnesium phosphate, as a product, will be reused to remove NH3-N in wastewater so that dosage cost can be saved. In order to determine impact factors and optimum process parameters, NH3-N removal and struvite recovery rates are tested when recycling the materials many times. The results are fitted and the fitting equation can predict the change trend of dosage cost with recycle times. Struvite was prepared by magnesium chloride hexahydrate, ammonium chloride, and anhydrous disodium hydrogen phosphate at a molar ratio of 1:1:1. After stirring ingredients for 20 minutes in sodium hydroxide solution at pH 10, the product was then obtained by suction filtration, washing, 12 hours drying and smashing, and ammonium chloride solution (1000 mg NH3-N/L); mature leachate from an aged closed landfill and fresh leachate from Shanghai Refuse Landfill were used as leachate influents. NH3-N and COD concentration of mature leachate was 700
800 mg/L and 1000
2400 mg/L, respectively. NH3-N and COD concentrations of fresh leachate were 2100
2300 mg/L, 54,000
56,000 mg/L, respectively. Structure of the struvite experiences variation of type (2.1) to (2.5) with the increase of temperature. Struvites gradually lose crystal water and become anhydride with collapsed crystal structure. First, the struvites lost five crystal waters from the hydrate between 80 C and 90 C and then became anhydrous MgNH4PO4 after pyrolysis under the condition of 100 C. The crystal structure of MgNH4PO4 was constant at the temperature of 65 C.

34

Pollution Control Technology for Leachate From Municipal Solid Waste

When the pyrolysis temperature was up to 100 C, MgHPO4 and MgNaPO4 were generated due to the dehydration and ammonia release of struvites under the alkaline environment. When the pyrolysis temperature increased further, the amorphous pyrophosphate appeared; and when the pyrolysis temperature was up to 600 C, Mg2P2O7 crystals generated. So pyrolysis products of struvite at 100 C can be recycled and reused to remove ammonia nitrogen from leachate in alkaline environment. 80 C

MgNH4 PO4 U6H2 O






! MgNH4 PO4 UH2 O 1 5H2 O 100 C1NH3

MgNH4 PO4 UH2 O






! MgNH4 PO4 100 C

MgNH4 PO4 1 NaOH






! MgNaPO4 1 NH3 m 1 H2 O 100 C

MgNH4 PO4






! MgHPO4 1 NH3 m 600 C

2MgHPO4






! MgP2 O7 1 H2 O

ð2:1Þ ð2:2Þ ð2:3Þ ð2:4Þ ð2:5Þ

2.2.1 Crystal Variation and Microstructure of Struvite Thermal analysis of struvites produced from various sources leachate is given in Fig. 2.2. It can be seen that the peak value of the weight loss rate TGA curves reached between 95 C and 100 C in fresh and mature leachate, which was also the most obvious endothermic peak of the DSC. It indicated that the crystal structure of struvite hexahydrate MAP was destroyed under this temperature, unceasingly absorbing the heat of the environment and gradually losing its water of crystallization and ammonium ions. Then phosphate further decomposed and the impurities like organic matter gasified through pyrolysis at high temperature. Because the pyrolysis products of the struvite were noncrystalline in this temperature, no significant weightlessness and endothermic peaks appeared. Compared with the struvite in mature leachate, the weightlessness rate of the struvite in fresh leachate was higher by nearly 10% under high temperature of 800 C because of the existence of organic impurity in the struvite. Weight loss rate and heat flow of struvite in fresh leachate was significantly lower than struvite in mature leachate, suggesting that morphology and structure of struvite in fresh leachate was damaged incompletely under high temperature, especially incomplete nitrogen release under 100 C, which led to ineffective regeneration under high temperature.

2.2.1.1 Effect of Pyrolysis Temperature on Crystals Variation of Struvite As shown in Fig. 2.3, the crystal structure of struvite produced from mature leachate experiences destruction and remodeling with the increase of

Physical and Chemical Treatment Processes for Leachate Chapter | 2

35

FIGURE 2.2 Thermal analysis of struvites produced from various sources leachate (Samples were dried at 65 C for about 10 hours and stored in dryer; left: Thermo gravimetric Analysis, TGA, right: Differential Scanning Calorimetry, DSC).

12,500

12,500

(A)

(B) MgNH4PO46H2O

10,000

7500

Intensity

Intensity

10,000

5000

7500

5000

2500

2500

0

0 10

20

30

40 2 Theta

50

10

60

12,500

(C)

40 2 Theta

50

60

Mg2P2O7

20,000

7500

Intensity

Intensity

30

(D)

10,000

5000

15,000

10,000

2500

5000

0 10

20

25,000

20

30

40 2 Theta

50

60

0 10

20

30

40

50

60

2 Theta

FIGURE 2.3 X-ray diffraction pattern of the struvite at various pyrolysis temperatures (The struvite produced from mature leachate was dried at 65 C for about 10 h and stored in a dryer. (A) raw struvite; (B) the struvite after pyrolysis at 100 C for 4 h; (C) the struvite after pyrolysis at 200
400 C for 4 h; (D) the struvite after pyrolysis at 600 C for 4 h).

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Pollution Control Technology for Leachate From Municipal Solid Waste

temperature. From Fig. 2.3A, X-ray diffraction analysis of the struvite without pyrolysis are exactly in accordance with the standard map of MgNH4PO4  6H2O. Then all the diffraction peaks of struvite disappeared after pyrolysis at 100 C for four hours (Fig. 2.3B), which indicated that all crystalline structures of MgNH4PO4  6H2O were destroyed. From Fig. 2.3C, it can be seen that pyrolysis products of struvite under the condition of 200
400 C has no obvious peak shape, proving that pyrolysis products still do not form crystal. However, the amorphous phosphate was gradually transformed into amorphous pyrophosphate and the diffraction peak of Mg3(PO4)2 crystal disappeared in the X-ray diffraction pattern, which suggested that pyrolysis products of the struvitev was a mixture of amorphous MgHPO4 and MgNaPO4. As shown in Fig. 2.3D, the struvite is decomposed into Mg2P2O7 under 600 C.

2.2.1.2 Effect of Recycle Times on the Crystals Structure of Struvite The fresh struvite obtained from leachate was heated and products were used for another leachate for ammonium precicpitation, namely one recycle, and so on. As shown in Fig. 2.4, crystal structure of struvite is significantly damaged at pyrolysis temperature of 100 C and diffraction peak basically disappeared. It’s not hard to find that all disappeared except the diffraction peaks at 15.8 and 31.8 degrees after recycling once. All the diffraction peaks disappeared after recycling five times, indicating that MgNH4PO4  6H2O crystal was already transformed into amorphous phosphate. The increase of recycle times was bound to increase the impurity content in the pyrolysis products of struvite. Simultaneously, the repeated high temperature pyrolysis and annealing treatment would also affect crystal structure of pyrolysis products. This phenomenon was more obvious at 600 C. Impurity peaks of pyrolysis products were little after recycling once and the main component of impurity was pyrophosphate magnesium. However, the peaks shape became complicated after recycling five times. The peaks appeared in 14.1 and 15.6 degrees may be the crystal diffraction peaks of the decomposed MgNH4PO4  6H2O and CaNH4PO4  7H2O, respectively. The peaks of crystal components doped with carbon appeared in the 27.9, 37.9, and 57.7 degrees. In addition, the peaks of Na3PO4, K2HPO4, and Mg2P2O7 crystals doped with other metal compositions appeared between 30 and 40 degrees. 2.2.1.3 Effect of Leachate From Various Sources on Crystals Structure of Struvite The crystals structure of struvite based on various sources of leachate were different. MgNH4PO4  6H2O crystal was pure white, while the struvites produced from mature leachate and fresh leachate were gray and tawny, showing that the impurities of the leachate had some influence on struvites and

Physical and Chemical Treatment Processes for Leachate Chapter | 2

10,000

10,000 (B)

8000

8000

6000

6000

Intensity

Intensity

(A)

4000 2000

4000 2000

0 10

20

30

40 2 Theta

50

0 10

60

25,000

20

30

40 2 Theta

50

60

10,000 (C) Mg2P2O7

20,000

8000

15,000

Intensity

Intensity

37

10,000

6000

Mg2P2O7

CaH+N2O7 SiO2 CaH3N5O2 CO(NH2)2 Cyclic

MgNH4PO4-6H2O Na7Mg9(P2O7)4 Mg3(PO4)2 Na3PO4 K4(HPO4)2

aromatic hydrocarbons

4000 (D) 2000

5000 0 10

20

30

40 2 Theta

50

60

0 10

20

30

40 2 Theta

50

60

FIGURE 2.4 X-ray diffraction pattern of the struvite after various pyrolysis recycling times ((A) pyrolysis temperatures of 100 C and recycling once; (B) pyrolysis temperatures of 100 C and recycling five times; (C) pyrolysis temperatures of 600 C and recycling once; (D) pyrolysis temperatures of 600 C and recycling five times).

pyrolysis products. Fig. 2.5 shows X-ray diffraction spectrum of struvite and its pyrolysis products produced at 100 C. As shown in Fig. 2.5A, impurity contents are less in the struvite produced from mature leachate and the spectral phase is more close to that of a pure sample. The struvite produced from fresh leachate contained a lot of CaCO3, which was proved by the dramatic rise in the diffraction peak of 29.5 degrees in Fig. 2.5B. The crystal structure of struvite produced from mature leachate disappeared after pyrolysis (Fig. 2.5C) and impurities like CaCO3 and NaCl still existed in the struvite produced from fresh leachate (Fig. 2.5D). Combined with the results of ammonia nitrogen removal efficiency, Ca and high concentrations of volatile organic acids in fresh leachate may be the main influence factors for crystallization and pyrolysis of the struvite.

2.2.1.4 Particle Size of Struvite Diffraction peaks within 10
25 degrees are calculated by diffraction line broadening and average crystal sizes are calculated by Scherrer’s equation after deducting background, picking up the peak and fitting, etc., by JADE 6.5 software, as shown in Table 2.1 and Figs. 2.6 and 2.7, respectively. Because the pyrolysis regeneration and drying of struvite crystals cooled

2500

12,500 (A)

(B)

MgNH4PO4-6H2O

2000 Intensity

10,000 Intensity

MgNH4PO4-6H2O CaCO3

7500 5000

1500 1000 500

2500 0 10

20

30

40

50

0 10

60

2 Theta

10,000

20

40

60

CaCO3 NaCl

(D) 1000

8000

50

2 Theta

1200

(C)

30

Intensity

Intensity

800 6000 4000

600 400

2000 0 10

200

20

30 2 Theta

40

50

60

0 10

20

30

40

50

60

2 Theta

FIGURE 2.5 X-ray diffraction pattern of the struvite produced from fresh and mature leachate and heating products ((A) the struvite produced from mature leachate; (B) the struvite produced from fresh leachate; (C) the struvite produced from mature leachate after pyrolysis at 100 C; (D) the struvite produced from fresh leachate after pyrolysis at 100 C).

TABLE 2.1 Average Particle Size of Struvite Crystals and Its Pyrolysis Products Reference value

Hexahydrate magnesium ammonium phosphate

The struvite in mature leachate

Alkalinity precipitation in fresh leachate

The struvite in fresh leachate

Magnesium pyrophosphate

Pyrolysis products of the struvite in mature leachate at 400 C

Pyrolysis products of the struvite in fresh leachate at 600 C

Particle size (nm)

72.3

82

89.3

150.1

30.6

53.2

62.4

The standard deviation of particle size (nm)

1.5

4.7

17.8

0.9

0.9

1.6

2.4

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Pollution Control Technology for Leachate From Municipal Solid Waste MgNH4PO4·6H2O

(A)

(B)

FW (S)*Cos (Theta) 0.474

Mature leachate map FW (S)*Cos (Theta) 0.524

0.000

0.000

0.746

0.000

0.74

0.000

Sin (Theta)

Sin (Theta) *Fit Size Only: XS (?) = 723 (15), Strain (%) = 0.0, ESD of Fit = 0.00065, LC = 0.648

(C)

*Fit Size Only: XS (?) = 820 (24), Strain (%) = 0.0, ESD of Fit = 0.00078, LC = 0.742

(D)

Fresh leachate NaOH precipitation

Fresh leachate map FW (S)*Cos (Theta) 0.347

FW (S)*Cos (Theta) 0.155

0.000

0.000 0.000

0.000

0.71

0.642 Sin (Theta)

Sin (Theta) *Fit Size Only: XS (?) = 893 (47), Strain (%) = 0.0, ESD of Fit = 0.0006, LC = 0.739

*Fit Size Only: XS (?) = 1501 (178), Strain (%) = 0.0, ESD of Fit = 0.00048, LC = 0.66

FIGURE 2.6 X-ray diffraction analysis of the struvites in reference to average particle size. (A)

MgNH4PO4·6H2O after 600ºC

FW (S)*Cos (Theta) 0.857

0.000 0.000

0.732 Sin (Theta)

*Fit Size Only: XS (?) = 306 (9), Strain (%) = 0.0, ESD of Fit = 0.00158, LC = 0.009

(B)

Mature leachate map after 400ºC

(C)

FW (S)*Cos (Theta) 1.264

Mature leachate map after 600ºC

FW (S)*Cos (Theta) 1.076

0.000 0.000

0.73 Sin (Theta)

*Fit Size Only: XS (?) = 532 (9), Strain (%) = 0.0, ESD of Fit = 0.00066, LC = 0.698

0.000 0.000

0.731 Sin (Theta)

*Fit Size Only: XS (?) = 624 (178), Strain (%) = 0.0, ESD of Fit = 0.00081, LC = 0.549

FIGURE 2.7 X-ray diffraction analysis of pyrolysis products of the struvites in reference to average particle size.

Physical and Chemical Treatment Processes for Leachate Chapter | 2

41

slowly in the dryer, which had been in the oven and muffle furnace for several hours after solid-liquid separation, there were no additional micro stress during the preparation process. The average particle size of hexahydrate MAP produced from ammonium chloride preparation was the minimum, only 72.3 nm. Particle size of the struvite in mature leachate was slightly larger than that of the former (82.0 nm), indicating that there was a certain amount of impurity composition in its crystal structure. Particle size of the struvite in fresh leachate was 150.1 nm, showing that a large amount of impurities, especially the Ca ions, led to lattice doping as large changes had taken place in the morphology of crystals. Fig. 2.8 reflects microstructure of pyrolysis products produced from the struvite in mature leachate under different temperatures. Fig. 2.8 (1) is hexahydrate MAP crystals with complete crystal structure, which was the grain growing from MAP crystals with regular geometry and good symmetry. The crystals was equiaxed, sphenoid, stumpy, or thick plated and their surface was very smooth due to a good crystallization, and the grain sizes were 5 B 20 μm in scanning electron microscopy (SEM), which was 50 B 200 times of raw crystal sizes using X-ray diffraction pattern. Fig. 2.8 (2) shows the pyrolysis products at 100 C with irregular geometry and no obvious

FIGURE 2.8 Micromorphology variation of pyrolysis products from the struvite at different pyrolysis temperatures (1 hexahydrate MAP; 2 pyrolysis products from the struvite at 100 C after recycling many times; 3 pyrolysis products from the struvite at 400 C after recycling many times; 4 pyrolysis products from the struvite at 600 C after recycling many times).

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Pollution Control Technology for Leachate From Municipal Solid Waste

symmetry, as clintheriform and equiaxed crystals rupture in the middle to gradually form caryokinesis, where a large number of free broken granulums and organic matter was attached. Fig. 2.8 (3) shows pyrolysis products at 400 C with preliminary geometry and a large amount of impurity particles due to the filiform crystals under high temperature pyrolysis and flocculent particles were transformed into columnar crystals. Fig. 2.8 (4) is pyrolysis products at 600 C with clear crystal structure, rough grain surface and no obvious symmetry, which should be long-plate polyhedral structure but for distortion of part crystal face and a large amount of impurities combined at the same time. Its crystalline condition was relatively complex due to inactivation effect, indicating that impurities accumulated in the leachate and combined with effective components of the struvites such as phosphorus, magnesium and ammonia under the condition of high temperature after recycling many times.

2.2.2 Leachate Quality Characterization via Struvite Addition and Formation The NH3-N in the leachate can be precipitated by the form of struvite. Meanwhile, the organic matters may be removed by adding struvite. Analysis of three-dimensional fluorescence spectrum and gel filtration chromatography of fresh and mature leachate before and after struvite precipitation is tested. As shown in Fig. 2.9, distribution form of organic matters in fresh and mature leachate is different. The main fluorescent peaks in the fluorescence spectrum of mature leachate were in the Ex/Em 250/460 and Ex/Em 310/440, which represented the fulvic acid in UV region and fulvic acid in visible region, respectively. The certain fluorescence effect appeared in the Ex/Em 360/480 indicated the presence of a certain amount of humic acid. On the contrary, the main fluorescent peaks in the fluorescence spectrum of fresh leachate appeared in the Ex/Em 360/480 and Ex/Em 270/460 represented the presence of humic acid and its degradation products, respectively. The characteristic peak of fluorescence spectrum of mature leachate appeared no obvious change, while the peaks intensity appeared slightly change after adding NaOH solution and the struvite. The peak intensity of Ex/Em decreased gradually while the peak intensity of Ex/Em was almost the same, indicating that some fulvic acid in UV region was removed while the fulvic acid in visible region was basically unchanged. The characteristic peak of fluorescence spectrum of fresh leachate had an obviously change after adding NaOH solution and struvite. The small molecules degradation substances of humic acid was removed with the increase of pH and the precipitation of struvite, but the fulvic acid peaks showed up in the process. The results indicated that the struvite had a performance for

Physical and Chemical Treatment Processes for Leachate Chapter | 2

43

FIGURE 2.9 Fluorescence spectrum analysis of leachate from various sources and effluents before and after struvite precipitations.

removing fulvic acid under the alkaline environment and could also remove fulvic acid by sweep coagulation. Analysis of gel filtration chromatography for leachate was conducted. The results of molecular weight distribution of different leachate are shown in Fig. 2.10. It can be seen that the molecular weight distribution of mature leachate was single, while that of fresh leachate was comparatively broader. In addition, the retention time of fresh and mature leachate shortened slightly after adding NaOH solution, indicating the slight increase of molecular weight. The reason may be the compression of double charge caused by alkaline environment and the molecular weight of some organic molecules were enhanced by the reinforcement. A strong peak appeared at 21.7 C after

44

Pollution Control Technology for Leachate From Municipal Solid Waste

Mw 1M

100 k

Mw

10 k

1k

100

1M

1500

100 k

10 k

1k

100

25,000 Mature leachate

Fresh leachate

1250 20,000

Intensity

Intensity

1000

750

15,000

10,000 500

5000

250

0 10

15

20

0 10

25

15

Retention time (min) 1M

100 k

20

25

Retention time (min)

Mw 10 k

Mw 1k

100

1M

2500

100 k

10 k

1k

100

25,000 pH adjusted mature leachate

2000

pH adjusted fresh leachate

1500

Intensity

Intensity

20,000

15,000

1000

10,000

500

5000

0 10

0 10

15

20

25

15

1M

100 k

1k

100

1M

100 k

1k

10 k

100

60,000 Struvite treated mature leachate

50,000

50,000

Struvite treated fresh leachate

40,000 Intensity

Intensity

25

Mw

Mw 10 k

70,000

60,000

20 Retention time (min)

Retention time (min)

40,000

30,000

30,000 20,000 20,000 10,000

10,000

0 10

0 15

20 Retention time (min)

25

10

15

20

25

Retention time (min)

FIGURE 2.10 Gel filtration chromatography of leachate from various sources and effluents after struvite precipitation.

adding magnesium salt and phosphate, illustrating that the remaining free chlorine ion and sodium ion had an influence on the molecular weight distribution when the struvite was added. From Fig. 2.11 it can be seen that the molecular weight of influent and effluent from various sources may be classified into four ranks, i.e., , 250, 250
2.5k, 2.5k
5k, and . 5k Da, which represent the easily biodegradable, biodegradable, degradation-resistant, and typical degradation-resistant macromolecular organics, respectively.

Physical and Chemical Treatment Processes for Leachate Chapter | 2

45

FIGURE 2.11 Molecular weight analysis of leachate from various sources and effluents after struvite precipitation.

The molecular weight distribution of leachate in 250
2500 Da increased due to the presence of salt ions in the leachate (0 # sample). The organic matter contents with medium and small molecular weight of mature leachate reduced gradually, while those with larger molecular weight increased. Compared with mature leachate, the molecular weight distribution of fresh leachate remained basically unchanged with the increase of recycling times. The total organic matter contents of fresh leachate was nearly 25
50 times than mature leachate based on COD at a dosage 1.1 for the struvite and the concentration of ammonia nitrogen. The struvite was quickly saturated for

46

Pollution Control Technology for Leachate From Municipal Solid Waste

treating fresh leachate because of its high concentration organic matters, with little effect on molecular weight distribution of fresh leachate. Addition of struvite in mature leachate helped remove part of small molecule organic matter with molecular weight , 250 Da, which tended to be fulvic acid in the UV region, while the small molecular organic matter removed in fresh leachate tended to be degradation products of humic substances (HS) under microbial action. The removal capacity of struvite was saturated quickly and couldn’t be regenerated by pyrolysis, indicating that the organic matters with small molecule of leachate can lead to the inactivation of the struvite. The main soluble organics in leachate were fulvic acid and humic acid with large molecular weight in the visible region, which led to a poor removal efficiency of soluble organic constituents.

2.2.3 Ammonia Nitrogen Removal Process and Reagent Cost Analysis The dosage of MAP after pyrolysis was 1.1 times of ammonia nitrogen content. pH was maintained 9.5
10.0 using 4 mol/L sodium hydroxides, and the picture was shown in Fig. 2.12 after stirring for 20 minutes. The suspension was filtrated by filter membrane of 0.45 μm. Then the supernatant was diluted for detection and the solid was weighed after being washed by deionized water and then dried at 65 C in an oven for 12 hours, and finally was pyrolysed for four hours under different temperatures. Fig. 2.12A and B are fresh leachate and mature leachate with the addition of NaOH, and indicate that part of pollutants of fresh leachate precipitated in alkaline environment. Fig. 2.12C is fresh leachate after struvite precipitation, the chromaticity and turbidity of which decreases obviously, and the struvite is gray black. Fig. 2.12D is mature leachate after struvite precipitation, with chromaticity and turbidity reduced slightly, and the struvite is pale. The difference was mainly contributed to the composition of organics present.

FIGURE 2.12 Ammonia nitrogen removal in fresh and mature leachate using struvite precipitation (denoted as A, B, C, and D from left to right).

Physical and Chemical Treatment Processes for Leachate Chapter | 2

47

The leachate had some inhibition effect on the pyrolysis recycle of struvite. Impurities such as organics, salt, and silica in leachate were rolled into the solid phase in the process of crystallization and precipitation, which had a competition in the formation of MgNH4PO4  6H2O crystal. At the same time, the impurities would have the inhibition effect on the re-generation of struvite via coercion, adulteration, and replacement. The inhibition effect of leachate on pyrolysis re-generation of the struvite mainly included the following four points. 1. Inhibition effect of organics and metal ions on the crystallization of MgNH4PO4  6H2O High concentration of organic matters in leachate, especially a large number of volatile organic acids of fresh leachate and CaCO3 in alkaline environments made the Mg21 and PO32 4 hard to combine into crystals by collision. As a result, the removal efficiencies of ammonia nitrogen were different at the same addition of MgCl2  6H2O and anhydrous Na2HPO4 in mature leachate and fresh leachate. 2. Competition of crystallization between CaNH4PO4  7H2O and MgNH4PO4  6H2O CaNH4PO4  7H2O and MgNH4PO4  6H2O generated simultaneously with higher Ca21 content in leachate, and both of them could effectively remove ammonia nitrogen. But the lattic stability and pyrolysis temperature of CaNH4PO4  7H2O were higher because of the bigger ion radius of Ca21. However, exceedingly high temperature would turn the pyrolysis products of MgNH4PO4  6H2O into Mg2P2O7, leading to ammonia nitrogen removal capacity of struvites from fresh leachate slump at the same pyrolysis temperature. 3. Inhibition effect on pyrolysis products of MgNH4PO4  6H2O The insoluble matters, including crystalline impurities and precipitation in alkaline environments, were collected with struvites, which were hard to be removed by simple washing. In addition, the contents of insoluble matters gradually increased in the process of repeated pyrolysis recycles, causing inhibition effect on pyrolysis of MgNH4PO4  6H2O and yield reduction of ideal pyrolysis products like MgHPO4 and MgNaPO4. 4. Inactivation of pyrolysis products caused by the lattice doping The impurities in the leachate accumulated, replaced, and mingled effective components, such as phosphorus, magnesium, ammonia, etc., of the struvite under the condition of high temperature, leading to the crystal lattice defects, which made the pyrolysis products difficult to adsorb ammonia effectively to generate MgNH4PO4  6H2O crystal again in leachate. There was a good linear relationship between ammonia nitrogen removal rate and the recycling times in ammonium chloride solution and mature leachate using struvite precipitation. Linear fitting formula can be obtained after recycling once as follows.

48

Pollution Control Technology for Leachate From Municipal Solid Waste

Ammonium chloride solution: Ra1 5 100 2 3:04n;

ðR2 5 0:988Þ

Mature leachate: Ram 5 100 2 3:29n;

ð2:6Þ

ðR2 5 0:999Þ

ð2:7Þ

Ra1: ammonia nitrogen removal rate in ammonium chloride solution using struvite precipitation, %; Ram: ammonia nitrogen removal rate in mature leachate using struvite precipitation, %; n: recycling times of the struvite. It can also be inferred that the ammonia nitrogen removal rate declines by percentage with the increase of recycling times. Exponential fit of the ammonium chloride solution, mature leachate, and fresh leachate can get good results, with a fitting equation as follows. Ammonium chloride solution: Rb1 5 100 0:968n ; Mature leachate: Rbm 5 100 0:965n ;

ðR2 5 0:965Þ

ðR2 5 0:955Þ

Fresh leachate: Rby 5 77:0 0:717n ;

ðR2 5 0:965Þ

ð2:8Þ ð2:9Þ ð2:10Þ

Rb1: ammonia nitrogen removal rate in ammonium chloride solution using struvite precipitation, %; Rbm: ammonia nitrogen removal rate in mature leachate using struvite precipitation, %; Rby: ammonia nitrogen removal rate in fresh leachate using struvite precipitation, %; n: recycling times of the struvite Reagent dosing cost of the struvite for single time is extremely high and treatment efficiency is low, as a pure chemical method is used to remove ammonia nitrogen from leachate. But removal capacity of ammonia nitrogen rises sharply after recycling. Equivalent weight of the struvite is 4.39 after recycling five times in mature leachate, equivalent to 4.39 times of raw struvite and reagent cost reduces to 22.78% of the raw material for single use. Equivalent weight of the struvite is 1.63 after recycling five times in fresh leachate, equivalent to 1.63 times of raw struvite and reagent cost reduces to 61.35% of the raw material for single use. Xn An 5 a ð2:11Þ 0 n n

a0 5 1:1URe0 ; an 5 Ren UL1 Rcn

ð2:12Þ

Physical and Chemical Treatment Processes for Leachate Chapter | 2

49

Where, An: total equivalent weight of the struvite after recycling n times, dimensionless; an: equivalent weight of the struvite as recycling n times, dimensionless; Reo: ammonia nitrogen removal rate as recycling n times, %; Rcn: the struvite removal rate as recycling n times, %; As shown in Table 2.2, the cost of ammonia nitrogen removal in mature leachate using struvite pyrolysis under the best conditions is speculated to compare with ammonia nitrogen removal in ammonia chloride solution. Cost budget and ammonia nitrogen removal effect of ammonium chloride solution and mature leachate were consistent and they can get good results whether using linear fitting or exponential fitting methods. Equivalent weight of the struvite fitted was 4.24 after recycling five times, which was slightly lower than the actual experimental results of 4.70. Ammonia nitrogen removal efficiency of fresh leachate was lower and cost budget was higher. Besides, equivalent weight of the struvite fitted was 2.05 after recycling five times, which was slightly higher than the actual experimental results of 1.63. The two main limiting factors of the recycling process included the loss of struvite, namely the solid powder cannot be completely recycled leading to loss of raw material, and the struvite combined with impurities like organic matters of leachate, leading to inactivation of MAP that did not react with ammonia nitrogen during repeated precipitation and pyrolysis. The required amount of struvite for ammonia nitrogen removal can be calculated, according to the recovery rate and ammonia nitrogen removal rate using pyrolysis recycle as the calculating unit, where the molar ratio of struvite and ammonia nitrogen is 1.1:1. Removal cost of ammonia nitrogen per unit weight can be calculated according to the fact that MAP with equivalent weight of 19.64 kg can remove 1 kg ammonia nitrogen. Reagent cost of ammonia nitrogen removal up to discharge standards can be calculated as for 1000 mg NH3-N/L ammonium chloride solution, 750 mg NH3-N/L mature leachate, and 2000 mg NH3-N/L fresh leachate. As shown in Fig. 2.13, the total equivalent weight of the struvite rises sharply within 10 times with the increase of recycling times, including the recovery loss and inactivation of struvite as for mature leachate, and then rises slowly. Chemicals delivery cost of ammonia nitrogen removal using struvite precipitation greatly reduced to 15 RMB/kg NH3-N within the top 10 times, then it reduced slowly and ground to a halt after 15 times. Equivalent loss of the struvite in the experiment was mainly due to the loss in solidliquid separation and pyrolysis recycle according to the two-fitting formula of ideal conditions and impurities in mature leachate. So, it is important further improve working efficiency of the struvite, to reduce the chemicals delivery cost, to design reasonable solid-liquid separation as well as process

TABLE 2.2 Reagent Cost Evaluation of Ammonia Nitrogen Removal From Leachateusing Struvite Precipitation via Precipitation-pyrolysis-precipitation Process Leachate

Recovery rate of struvites (%)

Fit Method

Fitted equation

Ammonium chloride solution

 88.7

Linear fitting

R 5 100
3.04n R2 5 0.988

Exponential fit

Mature leachate

 88.9

Linear fitting

Exponential fit

Fresh leachate

 91.3

Exponential fit

R 5 100 0.968n R2 5 0.928

R 5 100
3.29n R2 5 0.999



R 5 100 0.965n R2 5 0.955



R 5 77 0.717n R2 5 0.965

Recycling times

Theoretical equivalent weight of the struvite

Equivalent reagent cost, Yuan (ton)

Ammonia nitrogen removal costs, Yuan (ton)

Ammonia nitrogen removal cost in leachate, Yuan (ton)

0

1.00

4000

78.56

78.56

5

4.24

943

18.52

18.52

10

5.73

698

13.72

13.72

20

6.65

601

11.81

11.81

0

1.00

4000

78.56

78.56

5

4.24

944

18.53

18.53

10

5.75

696

13.66

13.66

20

6.78

590

11.58

11.58

0

1.00

4000

78.56

58.92

5

4.24

944

18.54

13.90

10

5.71

701

13.76

10.32

20

6.60

606

11.90

8.93

0

1.00

4000

78.56

58.92

5

4.23

945

18.57

13.92

10

5.73

698

13.70

10.28

20

6.76

592

11.63

8.72

0

0.74

5405

106.16

212.32

5

2.05

1951

38.32

76.64

10

2.21

1810

35.55

71.10

20

2.23

1794

35.23

70.46

Physical and Chemical Treatment Processes for Leachate Chapter | 2

51

FIGURE 2.13 Fitting prediction of recycling times impact on removal cost of ammonia nitrogen in mature leachate using struvite precipitation.

flow and equipment of pyrolysis, and to reduce weight loss of the struvite and final products of pyrolysis recycle process. As shown in Fig. 2.14, the total equivalent weight of the struvite is basically saturated within 5 times with the increase of recycling times, including the recovery loss and inactivation of struvite as for fresh leachate. Chemical delivery cost of ammonia nitrogen removal using struvite precipitation greatly reduced to 36 Yuan/kg NH3-N, then it reduced slowly and ground to a halt after 10 times. Equivalent loss of the struvite in the experiment was mainly due to impurities in fresh leachate, according to the two-fitting formula of ideal conditions, which had far less influence on inactivation effect of MAP than the former. Therefore, it is important to improve working efficiency of the struvite furthermore and to reduce the chemicals delivery cost. Corresponding impurity removal process or purification steps is needed before adding the struvite to remove soluble organic matter and unnecessary metal ions that will cause inactivation of struvite.

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Pollution Control Technology for Leachate From Municipal Solid Waste

FIGURE 2.14 Fitting prediction of recycling times impact on removal cost of ammonia nitrogen in fresh leachate using struvite precipitation.

2.2.4 Optimum Process Conditions for Ammonia Nitrogen Removal 2.2.4.1 pH MAP is insoluble in leachate, but is hydrolyzed in acid environment, so the acid environment is not conducive to the formation of MAP precipitation. Appropriate increase of pH makes the system beneficial to sediment and ammonia nitrogen removal in alkaline environment, as shown in Fig. 2.15. Ammonia nitrogen removal rate is lower than 40% when pH value is below 8. The removal rate rises quickly in the weak alkaline environment and the removal rate can be nearly 90% at pH 10 or so. Ammonia nitrogen removal rate rises slowly with the further increase of pH value. Because the form of ammonia nitrogen varies with pH, it tends to be NH1 4 (ion state) in weak acid environment and NH3 (molecular state) in alkaline environment. When the pH value is more than 10, the ammonia

Physical and Chemical Treatment Processes for Leachate Chapter | 2

53

FIGURE 2.15 NH3-N removals in leachate at different pH using magnesium chloride hexahydrate and anhydrous disodium hydrogen phosphate methods (750 6 50 mg NH3-N/L, mature leachate, dosage of magnesium chloride hexahydrate and anhydrous disodium hydrogen phosphate at a molar ratio of 1:1, pH is adjusted with hydrochloric acid and sodium hydroxide solution).

nitrogen releases in the form of ammonia gas, which is not easy to collect. A large amount of ammonia releases quickly with rapid agitation, according to mass transfer and magnesium phosphate precipitate in the solution, at the same time at pH value over 11, which will have great influence on the formation of MAP. Therefore, the formation of ammonia and magnesium phosphate precipitation should be avoided in the process of ammonia nitrogen removal using MAP and cost of chemical delivery for adjusting pH should be considered at the same time. It is advisable to control pH in the range of 9.5
10.0.

2.2.4.2 Temperature Fig. 2.16 illustrates the effect of pyrolysis temperature and recycling times on removal effect of ammonia nitrogen. First of all, the pyrolysis temperature has a huge impact on ammonia nitrogen removal. Ammonia nitrogen removal rate was lower than 60% at 65 C and 200 C after recycling once and it was 40% at 400 C and 600 C, which was far lower than that of 80 C and 100 C, nearly 95%. With the increase of recycling times, ammonia nitrogen removal rate using pyrolysis products of the struvite always remained above 85% at 100 C, and it decreased to 70% at 80 C, and was generally lower than 10% at 65 C, 200 C, 400 C, and 600 C. Because pyrolysis products of the struvite formed new crystal under high temperature while ammonia release was not complete under too low temperature, which

54

Pollution Control Technology for Leachate From Municipal Solid Waste

FIGURE 2.16 Ammonia nitrogen removal rate at different pyrolysis temperature and recycling times using magnesium chloride hexahydrate and anhydrous disodium hydrogen phosphate method (750 6 50 mg NH3-N/L mature leachate, adding pyrolysis products of the struvite as 1.1 times of the ammonia nitrogen concentration, using hydrochloric acid and sodium hydroxide solution to adjust pH).

cannot generate precipitation combined with ammonia nitrogen of leachate and did not have removal ability of ammonia nitrogen. Therefore, the optimum pyrolysis temperature was 100 C.

2.2.4.3 Recycling Times Ammonia nitrogen removal capacity of pyrolysis products of the struvite gradually declines as recycling times increase. Ammonia nitrogen removal capacity decreased slowly with the increase of the recycling times at 80 C and 100 C, which decreased by 25% and nearly 10%, respectively, after recycling 5 times. The effect was obvious when pyrolysis temperature was above 200 C or below 65 C and ammonia nitrogen removal rate was lower than 20% after recycling twice. Ammonia nitrogen removal efficiency of struvite in mature leachate declined slower than that in fresh leachate as recycling times increased, because the impurities in fresh leachate affected the formation and decomposition of the struvite. Recovery rate of the struvite recycled from fresh leachate was slightly higher than that in ammonia chloride solution, which proved that the struvite was mixed with elemental composition except MAP. The hardness of leachate and small molecule organic acids make crystal structure of struvite damaged to some extent, whose nitrogen releasing and regeneration were affected, which cannot be removed by pyrolysis and washing.

Physical and Chemical Treatment Processes for Leachate Chapter | 2

55

2.2.4.4 Recovery Ammonia nitrogen removal in ammonium chloride solution and two kinds of leachate are experimented at pH of 9.5
10 and pyrolytic temperature of 100 C, using struvite precipitation, whose indicators are ammonia nitrogen removal rate and the recovery rate of pyrolysis products, as shown in Figs. 2.17
2.19, respectively. Results showed that the ammonia nitrogen removal efficiency and recovery rate of the struvite in ammonium chloride solution and leachate had small differences in many experiments. Ammonia nitrogen removal efficiency of the struvite in fresh leachate was significantly worse than two kinds of leachate, while the recovery rate was slightly higher the others.

2.3 AIR STRIPPING FOR REVERSE OSMOSIS SYSTEM CONCENTRATED LIQUID FROM RAW LEACHATE It has long been considered that leachate may be deeply treated in one step using reverse osmosis (RO) facilities without any pretreatment of biological processes. It is really a short-cut means. However, around 20%
40% concentrated liquid will be generated, with an extremely high COD, NH3-N, and salts (conductivity), which is a painful toxic liquid that must be managed properly. One-step RO system treatment flowsheet had ever been practiced widely in China and other countries, and the corresponding concentrated liquid had also been studied deeply.

FIGURE 2.17 Ammonia nitrogen removal in ammonium chloride solution using struvite precipitation (1000 mg NH3-N/L ammonium chloride solution, adding MgCl2  6H2O and Na2HPO4 as molar ratio 5 1:1 and adding pyrolysis products as 1.1 times of the ammonia nitrogen concentration, using hydrochloric acid and sodium hydroxide solution to adjust pH to 9.5
10, pyrolysis temperature 100 C).

56

Pollution Control Technology for Leachate From Municipal Solid Waste

FIGURE 2.18 Ammonia nitrogen removal in mature leachate using struvite precipitation (750 6 50 mg NH3-N/L mature leachate, adding MgCl2  6H2O and Na2HPO4 as molar ratio 5 1:1 and adding pyrolysis products as 1.1 times of the ammonia nitrogen concentration, using hydrochloric acid and sodium hydroxide solutions to adjust pH to 9.5
10, pyrolysis temperature 100 C).

FIGURE 2.19 Ammonia nitrogen removal in fresh leachate using struvite precipitation (2200 6 100 mg NH3-N/L fresh leachate, adding MgCl2  6H2O and Na2HPO4 as molar ratio 5 1:1 and adding pyrolysis products as 1.1 times of the ammonia nitrogen concentration, using hydrochloric acid and sodium hydroxide solutions to adjust pH to 9.5
10, pyrolysis temperature 100 C).

Physical and Chemical Treatment Processes for Leachate Chapter | 2

57

Colloidal particles, formed by reaction between leachate and dispersed materials such as small crystallite, antisludging agents, and so on, exist in RO concentrated liquid, with a typical pollutants of COD 15,000
30,000 mg/L, NH3-N 2000
3000 mg/L, conductivity 22,000
30,000 μs/cm. Addition of flocculating agent is aimed to remove some ions in concentrated liquid by aggregating particles to form large particles, which are easy to be separated by aggregation. It is considered that different flocculants can destabilize colloids in different ways, like compression of double-charged layer, charge neutralization, adsorption bridging, and furling of precipitation. Therefore, adding few electrolytes can cause particles aggregation and precipitation by compressing the double-charged layer and decreasing Zeta potential (ζ potential). Besides, with a dosage enough to precipitate hydroxides swiftly, colloid can be precipitated simultaneously due to furling caused by formation of hydroxides precipitation. Conductivity removal by MgO is mainly caused by coprecipitation. RO concentrated liquid, with high concentration NH3-N, may cause great hazard to natural environment and human beings. Common ways to remove NH3-N include biological methods and physico-chemical methods. For biological methods, NH3-N is ultimately transformed into nitrogen through nitrification and denitrification caused by nitrifying bacteria and denitrifying bacteria. C/N ratio of concentrated liquid is too low for biological treatment. For physico-chemical methods, ammonia may be removed via evaporation, ion exchange, breakpoint chlorination, wet air oxidation, chemical precipitation, membrane filtration, air striping, and so on. Air stripping can decrease NH3-N concentrations in wastewater efficiently. It refers to that, after adjusting pH of wastewater to alkalinity, increasing gas-liquid interface and decreasing partial pressure of nitrogen swiftly by blast aeration can strip free ammonia from wastewater to air with a removal up to 99%. Factors that can influence this process are various, like hydraulic load, NH3-N concentration in influent, leachate temperature, air temperature, barometric pressure, contact time, pH, and/or contact area. Results of experiments with different initial parameters may differ greatly. Increasing pH can shorten stripping time. When pH ranges from 9.5 to 10.5, NH3-N stripping efficiency is directly relevant to temperatures of air and wastewater. Important factors that influence air stripping process follows the sequence of pH value. temperature. air stripping time. gas liquid ratio. One-step RO concentrated liquid without biological pretreatment was collected from a Shanghai municipal solid waste landfill, and its quality is shown in Table 2.3, using supernatant of coagulative precipitation mixture for tests. COD concentration was 9000 mg/L while NH3-N concentration was 2900 mg/L approximately in this supernatant.

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Pollution Control Technology for Leachate From Municipal Solid Waste

TABLE 2.3 RO-concentrated Liquid Quality From a Municipal Solid Waste Landfill Without Biological Pretreatment Index

COD

TN

TP

NH3-N

pH

Concentration (mg/L)

22,000

10,235

32

9170

7.8

2.3.1 Effect of Mass Air Flow on Air Stripping Efficiency Air stripping tests on concentrated liquid with a NH3-N concentration 2120 mg/L at three mass air flow of 3, 6, 12 L/(min  L) are carried out as shown in Table 2.4. It can be seen that NH3-N removal increased with the increase of mass air flow. When mass air flow was 3 L/(min  L), NH3-N concentration was still over 400 mg/L after 5 hours with a slow change of NH3-N removal. When mass air flow was 6 L/(min  L), air stripping process can reach equilibrium in 4 hours with a NH3-N removal of 81.9%, and the NH3-N removal of over 90% can be obtained at a mass air flow of 12 L/(min  L) with an equilibrium time of 3.5 hours. In Fig. 2.20, beaker 1#, 2# and 3# represents initial air stripping condition under three different mass air flows, of 3, 6, and 12 L/(min  L), respectively. Beaker 4# represented later period air stripping condition under a mass air flow condition of 12 L/(min  L). When mass air flow was relatively high, it was hard to control the process as many bubbles were produced. Therefore, mass air flow should be less than 6 L/(min  L). It can be seen from Fig. 2.21 that when mass air flow is 3 L/(min  L), NH3-N concentration in effluent is unstable and relatively high. When mass air flow was 6 L/(min  L), air stripping efficiency was enhanced stably. Thus, the optimal mass air flow in air stripping process should be 6 L/(min  L).

2.3.2 Effect of Temperature on Air Stripping Efficiency Air stripping tests on RO concentrated liquid under four different temperatures conditions of 10 C, 20 C, 30 C, and 0 C were conducted. The temperature had a great influence on NH3-N stripping efficiency, which was proved by the fact that NH3-N stripping efficiency increased swiftly with the increase of temperature (Table 2.5). Free NH3 can combine with H2O to form NH3  H2O by hydrogen bond. This micro force between molecules can increase solubility of NH3 in leachate and interfere with NH3-N stripping process. Increase of temperature can not only change the equilibrium partial pressure of NH3 in gas phase to decrease solubility of NH3, but also increase molecular energy to break hydrogen bonds. From Figs. 2.22 and 2.23, it can be seen that NH3-N concentration in air stripping process can reach equilibrium after 5 hours. Air stripping

TABLE 2.4 Effect of Mass Air Flow on NH3-N Stripping Efficiency in RO Concentrated Liquid Treatment at pH 11 Adjusted by CaO No.

Mass air flow (L/min  L)

3

6

12

Time

NH3-N (mg/L)

Removal (%)

NH3-N (mg/L)

Removal (%)

NH3-N (mg/L)

Removal (%)

1

10 min

1710

19.3

1469

30.7

1203

43.2

2

30 min

1553

26.7

1239

41.5

766

63.9

3

1h

1179

44.4

805

62.0

463

78.2

4

2h

913

56.9

636

70.0

335

84.2

5

3h

684

67.7

503

76.3

178

91.6

6

3.5 h

720

66.0

464

78.1

137

93.5

7

4h

599

71.7

397

81.3

138

93.5

8

4.5 h

527

75.1

388

81.7

137

93.5

9

5h

454

78.6

384

81.9

137

93.5

60

Pollution Control Technology for Leachate From Municipal Solid Waste

FIGURE 2.20 RO concentrated liquid treatment by air stripping under different mass air flow conditions, with NH3-N concentration 2120 mg/L, average temperature 30 C, pH 11.

FIGURE 2.21 Effect of mass air flow on NH3-N removal in RO concentrated liquid treatment by air stripping.

performance was poor and NH3-N concentration in effluent was relatively high at 10 C. The performance became better at higher temperatures, as NH3-N stripping efficiency increased from 70% at 20 C to 80% at 30 to 40 C.

2.3.3 Effect of pH on Air Stripping Efficiency Adjust pH of concentrated liquid to 9, 10, 11, 11.5, and 12 by adding CaO. After adjusting pH, initial NH3-N concentrations changed with different dosage of CaO. According to the actual NH3-N concentration, NH3-N stripping

TABLE 2.5 Effect of Temperature on NH3-N Stripping Efficiency From RO Concentrated Liquid Temperature ( C)

10

20

30

40

Air stripping time

NH3-N (mg/L)

Removal (%)

NH3-N (mg/L)

Removal (%)

NH3-N (mg/L)

Removal (%)

NH3-N (mg/L)

Removal (%)

10 min

1881

11.30

1579

25.50

1469

30.70

1202

43.30

30 min

1816

14.30

1333

37.10

1239

41.50

956

54.90

1h

1492

29.60

1197

43.50

805

62.00

888

58.10

2h

1447

31.70

879

58.60

636

70.00

612

71.10

3h

1236

41.70

776

63.40

503

76.30

493

76.70

4h

1119

47.20

697

67.10

397

81.30

376

82.30

5h

991

53.30

704

66.80

384

81.90

342

83.90

7h

897

57.70

636

70.00

383

81.90

334

84.20

62

Pollution Control Technology for Leachate From Municipal Solid Waste

FIGURE 2.22 Effect of temperature on NH3-N removal in RO concentrated liquid treatment (NH3-N concentration 2120 mg/L, pH 11, mass air flow 6 L/(min  L)).

FIGURE 2.23 Effect of different temperatures on NH3-N removal in RO concentrated liquid by air stripping process (NH3-N concentration 2120 mg/L, pH 11, mass air flow 6 L/(min  L), reaction time 5 h).

efficiency can be calculated and the optimal pH can be obtained. Results are shown in Fig. 2.24 (operating conditions: mass air flow 6 L/(min  L), average temperature 30 C). It can be seen in Table 2.6 that NH3-N removal increases when increasing pH, which is due to the fact that the NH3-N stripping process can only remove free NH3 in leachate rather than NH1 4 , the ionic state. Ammonia exists in leachate in both forms of ammonium salt and free ammonia. pH is one of the main factors that influences the proportion of free ammonia in leachate. When increasing pH, the proportion of free

Physical and Chemical Treatment Processes for Leachate Chapter | 2

63

2600 pH = 9 pH = 10 pH = 11 pH = 11.5 pH = 12

NH3-N concentration in effluent (mg/L)

2400 2200 2000 1800 1600 1400 1200 1000 800 600 400 200 0

1

2

3 Reaction time (h)

4

5

FIGURE 2.24 Effect of pH on NH3-N removal in RO concentrated liquid by air stripping process. (NH3-N concentration 2120 mg/L, average temperature 30 C, mass air flow 6 L/(min  L)).

ammonia is bigger, thus facilitate the NH3-N removal. Therefore, pH should be increased in air stripping process to increase proportion of free ammonia. From Fig. 2.24, it can be seen that curves with pH of 9, 10 are distant from curves with pH of 11, 11.5, and 12. When pH is higher than 11, NH3-N concentration is relatively stable. After reaction for 4 hours, NH3-N concentration in effluent was less than 400 mg/L and residual NH3-N concentration decreased. After 5 hours, air stripping process nearly reached equilibrium. Concentrated solution, as a buffering system with a wide buffering range, has a complex composition of organic and inorganic matters. The increase of pH will cost massive alkali and higher running expenses. Besides, superabundant alkali may cause scaling in pipes easily. Thus, pH increase should have a limit. After a deep comparison, the most economical pH range was 11
11.5 while pH range in effluent of NH3-N removal process was 9.0
9.5.

2.3.4 Effect of Time on Air Stripping Efficiency Air stripping for concentrated liquid under different air stripping time of 10, 30 minutes, 1, 2, 3, 4, and 5 hours was conducted to determine the optimal air stripping time. Effect of time on NH3-N stripping efficiency is shown in Table 2.7 and Fig. 2.25 (pH 11, mass air flow 6 L/(min  L), temperature 30 C, initial NH3-N concentration 2121 mg/L). The time is one of most important factors that influences NH3-N stripping efficiency. The NH3-N removal increased as the air stripping time extended. Therefore, enough air stripping time is one of essential conditions to ensure NH3-N removal. From Fig. 2.25, it can be seen that NH3-N removal increases with time. In initial

TABLE 2.6 Effect of pH on NH3-N Removal From RO Concentrated Liquid by Air Stripping Process pH

9

10 Removal (%)

NH3-N (mg/L)

Removal (%)

11 Removal (%)

NH3-N (mg/L)

Initial NH3-N concentration (mg/L)

2711

10 min

2394

11.7

1866

17.4

1469

30.7

1670

20.2

1791

20.7

30 min

2224

18.0

1816

19.6

1239

41.5

1386

33.8

1560

30.9

1h

2103

22.4

1484

34.3

805

62.0

820

60.8

956

57.7

2h

2080

23.3

1499

33.6

636

70.0

609

70.9

722

68.0

3h

1830

32.5

1301

42.4

503

76.3

446

78.7

582

74.2

4h

1140

57.9

763

66.2

397

81.3

386

81.6

381

83.1

5h

1040

61.6

701

69.0

384

81.9

282

86.5

270

88.0

2121

NH3-N (mg/L)

Removal (%)

12

Air stripping time

2259

NH3-N (mg/L)

11.5

2093

NH3-N (mg/L)

Removal (%)

2259

Physical and Chemical Treatment Processes for Leachate Chapter | 2

65

TABLE 2.7 Effect of Time on NH3-N Stripping Efficiency From RO Concentrated Liquid Index

Stripping time 10 min

30 min

1h

2h

3h

4h

5h

NH3-N concentration after stripping (mg/L)

1469

1239

805

636

503

397

384

NH3-N removal (%)

30.7

41.6

62.0

70.0

76.3

81.3

81.9

FIGURE 2.25 Effect of stripping time on NH3-N removal from RO concentrated liquid (NH3N concentration 2120 mg/L, pH 11, average temperature 30 C, air volume 6 L/(min  L)).

2 hours, there was a linear increase in NH3-N removal. After 2 hours, increasing tendency became slower. After running for 4 hours, air stripping process reached equilibrium approximately, with little change in NH3-N stripping efficiency.

2.4 RO CONCENTRATED LIQUID TREATMENT BY COAGULATIVE PRECIPITATION COD concentration in the influent RO concentrated liquid was 22,000 mg/L and NH3-N concentration was 9170 mg/L. Removal of COD and NH3-N under 10 different coagulation conditions are shown in Fig. 2.26. Way 2 and way 7 had relatively better COD removals while way 7, 8, and 9 had relatively better NH3-N removals. When using MgO for coagulation only, the optimal dosage was about 7.5 g/L. At dosage of 10 g/L, settling ratio and COD removal decreased. Addition of Ca(OH)2 can help settlement of leachate and removal of NH3-N. Adopting way 7 can get relatively good COD and NH3-N removals. By adding 6 g/L MgO and 2 g/L Ca(OH)2 and then

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NH3-N COD

80

Removal (%)

70 60 50 40 30

1

2

3

4 5 6 7 Coagulation way

8

9

10

FIGURE 2.26 Removal of COD and NH3-N in RO concentrated liquid under different coagulation conditions (COD concentration 22,000 mg/L and NH3-N concentration 9170 mg/L) in Ways 1, 2, 3, 4, 5, 6, 7, 8, 9, 10, the addition of Ca(OH)2 1 MgO (g/L) were 0 1 5, 0 1 7.5, 0 1 10, 1 1 4, 2 1 3, 2 1 5, 2 1 6, 2 1 7, 2 1 8, 4 1 6, respectively.

TABLE 2.8 Treatment of Two Different Curing Agents on RO Concentrated Liquid Additive

Concentration (g/L)

COD (mg/L)

NH3-N (mg/L)

Conductivity (μs/cm)

pH

Curing agent (MgO 40% 1 CaO 60%)

10

7440

1028

19,400

7.5

stirring the mixture for 8 hours and standing for 1 hours, the best COD and NH3-N removals can be obtained, with COD removal of 55%
60% and NH3-N removal of 70%
75%. As a result, COD concentration in effluent was 8700
9000 mg/L while NH3-N concentration was about 2400
2900 mg/L, with the final pH of 9.6.

2.4.1 RO Concentrated Liquid With a Conductivity of 20,000 µs/cm Colloidal particles, formed by reaction between leachate and dispersed materials such as small crystallite, antisludging agents, and so on, exist in concentrated liquid. Addition of a flocculating agent is aimed to remove some ions in the solution by making particles form large particles easy to be separated by aggregation. The tested leachate COD concentration was 15,000 mg/L, NH3-N concentration was 2000 mg/L, conductivity was 22,000 μs/cm; 10 g/L curing agent (namely M1, MgO 40% 1 CaO 60%) is added to the concentrated liquid and the results are shown in Table 2.8.

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67

It can be seen that COD concentration decreased by 60% while NH3-N concentration by 50%, approximately. However, conductivity had little change, which meant that organic absorption of curing agents was dominant and inorganic salt absorption was faint. Inorganic salts have a big contribution to high conductivity of concentrated liquid and ions like Cl2, Na1, K1, and NO2 3 are hard to remove. Decrease in conductivity was due to the fact that addition of curing agents can remove charged colloidal particles, suspended solids, and particulate matters in concentrated liquid by adsorbing mechanism. Test was carried out to remove pollutants by coprecipitation and absorption of potassium alum and kaolin with different concentrations. Results are shown in Table 2.9. When dosage of potassium alum increased, conductivity and concentrations of K1 and SO22 increased subsequently. Increase of 4 NH3-N concentration of concentrated liquid may be a consequence of interference to spectrophotometry, which meant that addition of potassium alum and kaolin increased turbidity and influenced absorbance. Addition of potassium alum and kaolin had little effect on pollutant removal. In another test, lightweight MgO powder is used and the results are shown in Table 2.10. After adding MgO and stirring for 24 hours, the leachate was fluffy and color changed from black to light yellow. After standing for 1 hours, it was found that leachate with 5 g/L MgO dosed had a worse layering than it with 10 g/L dosed. From Table 2.11, it can be seen that MgO has a better conductivity removal effect than other materials. Different flocculants can destabilize colloid in different ways, like compression of double charged layer, charge neutralization, adsorption bridging, and furling of precipitation. Therefore, adding few electrolytes can cause particles aggregation and precipitation by compressing double charged layer and decreasing Zeta potential

TABLE 2.9 Treatment of RO Concentrated Liquid With a Conductivity of 20,000 μs/cm Using Potassium Alum and Kaolin Additive

Concentration (g/L)

NH3-N (mg/L)

Conductivity (μs/cm)

Potassium alum

2

2283.7

234,000

6

2250.8

243,000

10

2146.6

251,000

1.255

2119.2

219,000

2.4994

2228.9

225,000

Kaolin

Kaolin 2 g/L 1 Potassium alum 10 g/L

2689

24,900

Kaolin 4 g/L 1 Potassium alum 10 g/L

2530

27,200

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TABLE 2.10 Treatment of RO Concentrated Liquid With a Conductivity of 20,000 μs/cm Using Lightweight MgO Additives

Concentration (g/L)

COD (mg/L)

NH3-N (mg/L)

Conductivity (μs/cm)

pH

1319

16,500

9

MgO

5

8872

MgO

10

8856

464.1

15,500

9

MgO 5 g/L 1 Kaolin 10 g/L

9496

1154.6

16,300

8.5

MgO 5 g/L 1 Cement 10 g/L

9320

1039.5

16,300

9

MgO 10 g/L 1 Kaolin 10 g/L

9608

491.5

15,600

8.5

MgO 10 g/L 1 Cement 10 g/L

9648

606.6

15,900

8

(ζ potential). With a dosage enough to precipitate hydroxides swiftly, colloid can be precipitated simultaneously due to furling caused by formation of hydroxides precipitation. Conductivity removal by MgO was mainly caused by coprecipitation. Dosage of 5 g/L MgO can decrease conductivity by 10,000 μs/cm while conductivity did not change much with a higher dosage. Organic removal by MgO was inferior to the curing agents. After adding kaolin and cement to the MgO 1 concentrated liquid system, COD, NH3-N concentration, and conductivity all increased, showing that the addition of kaolin and cement interfered with coagulative precipitation of MgO mixture. NH3-N concentration decreased to 464.1 g/L with a removal of 75% when dosage of MgO reached 10 g/L, as MgNH4PO4 precipitated by reaction 32 among Mg21, NH1 4 and PO4 , making a contribution to NH3-N removal.

2.4.2 RO Concentrated Leachate With a Conductivity of 34,000 µs/cm Various materials are added to the leachate with a conductivity of 34,000 μs/cm and the results are shown in Table 2.11. MgO had a highest conductivity removal of 32%. However, addition of 5 g/L MgO had a relatively slow settling velocity and a unclear layering in 1 hours while addition of 10 g/L MgO will coagulate directly without layering for gel hardening caused by mixing MgO and MgCl2. Alumina and refuse incineration bottom slag had second best treatment effects. Main composition of bottom slag is a mixture of alumina and silicon oxide with a pH of 8.5
9.5. pH range for aluminium hydroxide precipitation is from 6 to 10. Therefore, adding alumina can help form flocculent precipitation of aluminium hydroxide and decrease conductivity by coagulative precipitation. Adding too much bottom slag can increase conductivity, in turn, for its limited coagulative ability.

TABLE 2.11 Treatment of RO Concentrated Liquid With a Conductivity of 34,000 μs/cm by Various Materials Additives

MgO

Agar

Cement

Alumina

MgO

Clay

Dried sludge

Alumina

MSW incineration bottom slag

Concentration (g/L)

5

10

10

10

10

20

20

20

40

Conductivity (μs/cm)

24,100

33,600

30,900

28,700

23,400

34,200

35,800

25,600

25,900

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Pollution Control Technology for Leachate From Municipal Solid Waste

32,000

Conductivity (µs/cm)

30,000 28,000 26,000 24,000 22,000 20,000

20

30 40 50 60 70 Incineration bottom slag dosage (g/L)

80

FIGURE 2.27 Effect of refuse incineration bottom slag dosage on conductivity removal from RO concentrated liquid with a conductivity of 34,000 μs/cm.

Leaching conductivity of bottom slag with a content of 40 g/L was about 24,000 μs/cm, as shown in Fig. 2.27 (Table 2.12).

2.4.3 RO Concentrated Liquid With a Conductivity Higher than 50,000 µs/cm Concentrated leachate with a conductivity of 51,800 μs/cm was treated using MgO, cement, CaO. Results are shown in Table 2.13; 10 g/L MgO can decrease conductivity by 10,000 μs/cm, while other additives have little influence. High polymer has ability of adsorption bridging. Therefore, a minimal conductivity of 47,600 μs/cm can be obtained with addition of 0.1 mg/L PAM to the concentrated liquid, as shown in Fig. 2.28. Adding 5 g/L MgO subsequently can decrease conductivity to about 40,000 μs/cm.

2.5 COD REMOVAL BY FLOCCULANTS FROM CONCENTRATED LIQUID OF REVERSE OSMOSIS SYSTEM FOR BIOLOGICAL PRETREATMENT EFFLUENTS There are more than 300 kinds of flocculants now, and, according to the chemical composition, they can be divided into inorganic flocculants and organic flocculants. Inorganic flocculants rely mainly on a cohesion function to neutralize charges on the particles, also known as flocculant (coagulation). Organic flocculants rely mainly on the bridging function and then make the particle settle down, also known as flocculant (flocculation). There are lesser kinds of inorganic product. The main products are aluminum and iron salts and their hydrolytic polymerization products, which play an important role in leachate treatment. There are many kinds of organic products, mainly high

TABLE 2.12 Treatment of RO Concentrated Liquid With a Conductivity of 51,800 μs/cm by Various Materials Additives

MgO

Agar

Cement

Alumina

MgO

Clay

Dried sludge

Alumina

MSW incineration bottom slag

Concentration (g/L)

5

10

10

10

10

20

20

20

40

Conductivity (μs/cm)

42,100

49,600

48,900

47,700

40,400

52,200

54,100

46,200

45,300

TABLE 2.13 Effect of Different Kinds of Flocculants on COD Removal From Concentrated Liquid With COD 444 mg/L of Reverse Osmosis System for Biological Pretreatment Effluents Type of flocculants

PAC

PFS

PAFC

Aluminum sulfate

Ferric sulfate

Alchlor

Ferric chloride

Quantity of reagent (mg/L)

1000

1000

1000

1000

1000

1000

1000

COD (mg/L)

375.0

361.2

363.3

425.8

425.8

378.9

359.4

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Pollution Control Technology for Leachate From Municipal Solid Waste

Conductivity (us/cm)

48,600

Conductivity

48,400 48,200 48,000 47,800 47,600 0.0

0.1

0.2 0.3 0.4 PAM concentration (mg/L)

0.5

FIGURE 2.28 Effect of PAM concentration on conductivity removal from RO concentrated liquid with a conductivity of 50,000 μs/cm.

polymer compounds, which can be divided into two parts, natural and synthetic, but the amount is smaller than inorganic class. According to its charged characteristics, organic flocculant can be divided into four classes: cationic polymer, anionic polymers, nonionic polymer, and amphoteric polymer.

2.5.1 Selection of Inorganic and Organic Polymer Flocculants Considering the purpose and the actual situation of removal test, seven kinds of common iron and aluminum salt flocculant were tested, including aluminum sulfate, chloride, ferrous sulfate, ferric chloride, polymeric ferric sulfate (PFS), polyaluminium chloride (PAC), and polymeric aluminum ferric chloride. The initial state is solid. According to the normal flocculant test, the minimum coagulation dose of appearing alum shavings should be determined first. However, according to the test, the chemical coagulation method of dealing with the biological leachate effluent needs a large amount of the coagulation, which is far higher than the minimum amount of alum flowers. The optimum pH of coagulation was 5, and dosed 1 L leachate with 6 mL poly iron (equivalent to 960 mg/L) in the coagulation process after the biological pretreatment, in order to save manpower and material resources. The results are shown in Table 2.13. According to the data from Table 2.13, ferric sulfate is chosen as the flocculant. Ferric salts are low molecular inorganic flocculants, including anhydride and crystal leachate substance and liquid, among which ferric sulfate with six crystals is commonly used. Ferric sulfate is a dark brown crystal and easily soluble in leachate. Its solubility increases with temperature rising. The precipitability of alum shavings is better than that of aluminum salts.

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As shown in Table 2.13, the removal ability of inorganic polymer flocculant is the same as inorganic low molecular flocculant, but the price of high polymer flocculant is far higher than that of low molecular flocculants. In addition, the efficiency of organic matter removal was not high, although the dosage of coagulation was great. Therefore, inorganic polymer was not needed for low concentration leachate after biochemical treatment. The main composition of low concentration refractory leachate from biological pretreatment is humus, which is negatively charged material, so cationic high polymer flocculants are chosen as the treatment agent, which has function of charge neutralization and adsorption. Cationic polymer can remove turbidity and color. It has the advantages of low dosage, producing less quantity of sludge and good decoloring and better COD removal compared with inorganic flocculant. The main varieties of cationic polyelectrolyte are the copolymer or homopolymer of dimethyl diallyl ammonium chloride and acrylamide. High molecular weight cationic polyelectrolyte is produced by free radical polymerization, such as the cationic polymer produced by vinyl polymerization type or macromolecular reaction type, with high molecular weight, low charge density, or powdery or liquid product forms. Low molecular weight cationic polyelectrolyte is produced by free radicals polycondensation reaction, with low molecular weight, high charge density, and liquid product forms. Liquid cationic polyelectrolyte was provided by Aristotle University Greece. Its commodity name was dzsyeflock—EF, and chemical composition was aggregated amine, the pH value of its 10% solution was 5, with a specific weight of 1.2 g/mL. It was soluble, and had strong viscosity (400 CPS), with a wide application pH range (pH 2
10) and ability to collect cationic organic substances in the leachate, such as humic acid, claybank acid, and so on. In addition, the polymer amine product had good decoloring performance.

2.5.2 Quantity of Reagent Required Beaker stirring test, zeta potential using microscopic electrophoresis apparatus, colloid titration, and filtering can be used for determining the optimum quantity of flocculant. Aluminum salt should be added in the form of dilute solution, while iron salt should be added in the form of dry cast or concentrated solution. It can be seen from Fig. 2.29 that ferric trichloride was useful for COD removal from leachate to some extent. y 5 2 4E 2 06x2 1 0:0156x 1 7:986 y: COD removal rate (%); x: concentration of ferric chloride (mg/L); R2: square error.

R2 5 0:9527

ð2:13Þ

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FIGURE 2.29 Relationship between concentration of ferric trichloride and COD removal from concentrated liquid of reverse osmosis system for biological pretreatment effluent.

From Eq. (2.13), the highest COD removal rate may be 21.8% by ferric trichloride with a flocculant dosing of about 1780 mg/L, 21% COD removal rate can be obtained at 1200 mg/L ferric chloride, and 18.34% COD removal rate at 1000 mg/L ferric chloride. From 1000 to 1200 mg/L, flocculant dosage increased by 200 mg/L, COD removal efficiency increased by about 2.7%, and from 1200 to 1780 mg/L, dosage increased by about 600 mg/L, but the removal rate only increased by 0.84%. Hence, the best dosage quantity of reagent was 1200 mg/L.

2.5.3 pH pH is one of the important factors that affect the coagulation performance. As for ferric chloride, the optimum pH range was from 5 to 8.4. The best dosage quantity of ferric chloride (1200 mg/L) was added into leachate, and then pH was adjusted to 4, 5, 6, 7, and the original, respectively. Relationship between pH and COD removal rate in the presence of 1200 mg/L ferric chloride is given in Fig. 2.30. It can be seen that COD removal rate increased as pH increased and then decreased when pH continued to rise, and the highest COD removal was at pH 5
6.

2.5.4 Flocculant Aid on Coagulants Flocculant didn’t work well alone when the concentration of leachate was low. Besides, COD removal efficiency was only 22% by ferric chloride of 1200 mg/L. So, flocculant aid was needed to improve ferric chloride flocculant coagulation performance. Polymer flocculant aid can be used to increase skeleton and weight of alum flowers and improve the structure of the alum flowers. Polyacrylamide is used as a flocculant aid and the results are shown in Table 2.14.

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75

FIGURE 2.30 Relationship between pH and COD removal rate in the presence of 1200 mg/L ferric chloride from concentrated liquid of reverse osmosis system for biological pretreatment effluent.

TABLE 2.14 Relationship Between Quantity of PAM and COD in the Presence of 1200 mg/L Ferric Chloride From Concentrated Liquid of Reverse Osmosis System for Biological Pretreatment Effluent Quantity of PAM (mg/L)

0

1

2

3

4

Concentration of ferric chloride (mg/L)

1200

1200

1200

1200

1200

COD (mg/L)

347.0

314.7

324.3

324.7

322.4

As shown in Table 2.14, the dose variation of PAM (average molecular weight was 7 million) had less effect on COD removal. Comparing with no PAM, COD removal can improve 7.3% (from 21.9% to 29.2%) with 1 mg/L PAM. Higher PAM dosing quantity may increase COD removal. As a result, under the condition of 1 mg/L PAM, 1200 mg/L ferric chloride and 1 mg/L anionic polyacrylamide (average molecular weight of 9 million), and stirred quickly for 1 minutes with stirring speed of 250 r/min, and then stirred slowly for 15 minutes with stirring speed of 80 r/min, and then precipitate for 10 minutes, COD can be removed to some extent.

2.5.5 Organic Polymer Flocculants Used for the flocculation precipitation of leachate are 120, 240, 360, 480, and 600 mg/L flocculants Dyeflock-EF, a cationic polymer amine, and the results are shown in Fig. 2.31.

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FIGURE 2.31 Relationship between cationic polymer amine concentration and COD removal rate from concentrated liquid of reverse osmosis system for biological pretreatment effluent.

1400 Ferric trichloride Cationic polymerized amine

Dosage (mg/L)

1200 1000 800 600 400 200 0 5

10

15 20 Removal rate of COD (%)

25

30

FIGURE 2.32 Comparison of ferric trichloride and cationic polymer amine on COD removal from concentrated liquid of reverse osmosis system for biological pretreatment effluent.

As shown in Fig. 2.31, cationic polymer amine can remove some of COD, and the maximum removal rate is 22% when flocculant dose is about 360 mg/L. Fig. 2.32 shows the removal performance of flocculant ferric trichloride and cationic flocculant polymer amine. It can be seen that COD removal rate of cationic flocculant polymer amine was better than that of ferric trichloride, with around 25% COD removal rate. However, under the same COD removal rate, the additive amount of cationic polymer amine was lower than that of ferric trichloride. Hence, whatever the flocculant used, the removal rate of COD from leachate was quite low due to the presence of humus.

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TABLE 2.15 Relationship Between the Quantity of Reagent and the Color From Concentrated Liquid of Reverse Osmosis System for Biological Pretreatment Effluent Quantity of Dyeflock
EF (mg/L)

120

240

360

480

600

Color (multiple)

110

90

75

55

40

Table 2.15 shows the color removal from leachate using cationic polymer amine, Dyeflock—EF. The color removal increased with the Dyeflock—EF dosing quantity increase, and the leachate chroma removal rate can reach 50% at Dyeflock—EF dosing quantity of 180 mg/L, and reach 80% at Dyeflock—EF dosing quantity of 600 mg/L. A Dyeflock—EF dosing quantity of 360 mg/L may be used in the practical applications.

2.6 LEACHATE COD REMOVAL BY COMPOSITE FLOCCULANTS COUPLED WITH OXIDANTS FROM CONCENTRATED LIQUID OF REVERSE OSMOSIS SYSTEM FOR BIOLOGICAL PRETREATMENT EFFLUENT 2.6.1 Composite Flocculant Mixed With Potassium Permanganate and Sodium Hypochlorite Since humus can help increase the stability of a colloid, the performance of flocculants on leachate is not good. To obtain better contaminants’ removal efficiency, organic matters present in the leachate should be decomposed to break stability of colloid and reduced toxicity, using preoxidation by potassium permanganate, which has a strong oxidizing property for humus, followed by removing residual organics by coagulative precipitation using a combination of manganese dioxide and flocculants. Ferric trichloride and potassium permanganate are chosen as composite flocculant. Yellow FeCl3treated leachate has chromaticity pollution, while potassium permanganate has a good decoloration effect; thus, under proper dosage proportions, combination use of potassium permanganate and flocculants can handle chroma pollution caused by flocculant. Using sodium hypochlorite for leachate pretreatment can destabilize hydrophilic organic impurity, facilitate coagulation process, reduce dosage of flocculant, and maintain nice sanitary conditions of leachate treatment structures. Chromaticity pollution and stink can also be removed due to the removal of organics. Sodium hypochlorite can enhance coagulation effects of ferrous ions, which may be due to hydrolysis property of ferrous ion. Ferrous ions ionized by ferrous sulfate can only generate simple complex, with a

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Pollution Control Technology for Leachate From Municipal Solid Waste

coagulation effect inferior to ferric salts. Sodium hypochlorite can oxidize ferrous ions, generating ferric ions. The specific reaction equation is: 6Fe21 1 3ClO2 1 3H2 O 5 2FeðOHÞ3 1 3Cl2 1 4Fe31

ð2:14Þ

Using sodium hypochlorite only can contribute to pH increase of leachate easily, which may make effluent unable to meet national wastewater discharge standards Since flocculants will release hydrogen ions during hydrolysis process, combination of flocculant and sodium hypochlorite can bring down pH by neutralizing hydroxyl ions. Meanwhile, only when pH is more than 8.0 can ferrous ions ionized by ferrous sulfate be oxidized to ferric ions by dissolved oxygen in solution. Then polynuclear complex can be generated, producing a good coagulation effect. Residual ferric ions will cause chromaticity pollution of treated leachate. When ferric ions react with colored substance, soluble matters with a darker color will be generated. Sodium hypochlorite has a great decoloration effect in leachate treatment. Therefore, combination use of sodium hypochlorite and ferrous sulfate can also handle the chromaticity pollution in effluent.

2.6.2 Composite Mixing Effects of Different Flocculants and Potassium Permanganate In order to observe composite mixing effects of potassium permanganate and different flocculants, three different common flocculants, ferric trichloride, ferrous sulfate, and aluminum sulfate, are chosen as composite mixtures with potassium permanganate. COD removals of these flocculants in low concentration leachate are shown in Fig. 2.33. It can be seen that coagulation-

FIGURE 2.33 COD removals of ferric trichloride, ferrous sulfate and aluminum sulfate for concentrated liquid from reverse osmosis system with biological pretreatment (concentration of flocculants 800 mg/L).

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79

supporting effect of potassium permanganate differs when using different flocculants for composite mixing. Composite mixing of ferric trichloride and potassium permanganate had a better COD removal effect than that of ferrous sulfate and potassium permanganate. When using aluminum sulfate for composite mixing with potassium permanganate, COD removal increased gradually as dosage of potassium permanganate was increased. When dosage of potassium permanganate was increased to 400 mg/L, the maximal COD removal, about 30%, can be obtained and leachate quality of effluent can reach COD 300 mg/L. When dosage of potassium permanganate was further increased, COD removal decreased, meaning the presence of the optimal dosage for coagulationsupporting effect of potassium permanganate on aluminum salt. The best dosing proportion of aluminum sulfate to potassium permanganate was 2:1, approximately. When composite mixing of ferrous sulfate and potassium permanganate, COD removal increased gradually as dosage of potassium permanganate was increased. Besides, using ferric trichloride can get a COD removal 10% higher than using ferrous sulfate, which meant potassium permanganate had coagulation-supporting effect on ferric salt. When using ferric salt only, ferric trichloride can get a COD removal about 5% higher than using ferrous sulfate. Specific data is shown in Table 2.16. When being mixed with potassium permanganate, ferric trichloride can get a COD removal 10% higher than using ferrous sulfate and 5% higher than using flocculant only, which meant potassium permanganate oxidized ferrous sulfate to ferric salt to enhance coagulation effect with cost of some potassium permanganate. Therefore, for COD removal, composite mixing of potassium permanganate and ferric trichloride had the best treatment performance; for enhancement of flocculant, coagulation supporting effects of potassium permanganate on ferrous salt was better.

TABLE 2.16 COD Removal Comparison of Ferric Trichloride and Ferrous Sulfate for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment Types

COD removal (%)

Ferric trichloride Ferrous sulfate

Dosage (mg/L) 200

400

600

800

1000

10.5

14.5

17.1

17.5

19.2

7.4

10.0

10.0

12.5

13.6

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Pollution Control Technology for Leachate From Municipal Solid Waste

2.6.3 Comparison of Composite Flocculants and Flocculants 2.6.3.1 Composite Mixing of Potassium Permanganate and Ferric Trichloride COD removal effect comparison of flocculant alone and composite coagulation agent made by potassium permanganate and ferric trichloride is shown in Fig. 2.34. Dosage of potassium permanganate was 400 mg/L, COD concentration of leachate used was 427.2 mg/L. As shown before, the optimal dosage of potassium permanganate was 500 mg/L and the corresponding COD removal was 23.8%. According to data fitting equation of potassium permanganate dosage and COD removal, when potassium permanganate dosage was 400 mg/L, COD removal was 21.0%. From these data, it can be seen that increasing 100 mg/L in potassium permanganate dosage only increased COD removal slightly (2.8%). Thus, a potassium permanganate dosage of 400 mg/L was determined. Potassium permanganate had an obvious coagulation-supporting effect for concentrated liquid from reverse osmosis system with biological pretreatment. COD removal was increased by about 20% compared to using ferric trichloride alone at potassium permanganate of 400 mg/L for composite mixing,. When dosing potassium permanganate with a concentration of 400 mg/L alone, COD removal was 20.5%. It meant that composite mixing of potassium permanganate and ferric trichloride produced a mutual superimposition effect with interaction. When using potassium permanganate for coagulation aid, curve of ferric trichloride optimal dosage remained the same and curve of COD removal got a translation.

FIGURE 2.34 Comparison of flocculant alone and composite coagulation agent on COD removal for concentrated liquid from reverse osmosis system with biological pretreatment (COD concentration of influent 427.2 mg/L).

Physical and Chemical Treatment Processes for Leachate Chapter | 2

81

2.6.3.2 Composite Mixing of Sodium Hypochlorite and Ferrous Sulfate COD removal comparison of ferrous sulfate alone and composite coagulation agent made by sodium hypochlorite and ferrous sulfate is shown in Fig. 2.35. Dosage of sodium hypochlorite was 300 mg/L and COD concentration of leachate used was 427.2 mg/L. According to the specific reaction equation of ferrous ion and hypochlorite, mole number ratio of ferrous sulfate and sodium hypochlorite was 2:1. Therefore, it can be calculated that oxidizing 1 mg ferrous sulfate costed 0.245 mg sodium hypochlorite. In the experiments, the maximal dosage of ferrous sulfate was 1000 mg/L and the corresponding dosage of sodium hypochlorite was 245 mg/L. Taken oxidation process, which may consume some sodium hypochlorite into consideration, the final dosage of sodium hypochlorite was determined to be 300 mg/L. In Fig. 2.35, under the same dosage of ferrous sulfate, dosage of sodium hypochlorite with a concentration of 300 mg/L can increase COD removal by 20%B25% from leachate. When dosage of ferrous sulfate was 600 mg/L, combination use of sodium hypochlorite and ferrous sulfate had the optimal effect. When dosage of ferrous sulfate was 600 mg/L, oxidation of ferrous ions needed 147 mg sodium hypochlorite and 300 mg/L sodium hypochlorite in a whole. Thus, residual sodium hypochlorite concentration was 153 mg/L. When using sodium hypochlorite with a concentration of 153 mg/L, COD removal was 11.7%. After composite mixing, COD removal was 34.1%, which meant that ferric salt generated by oxidation of ferrous sulfate can remove COD by 34.1%
11.7% 5 22.4%. When dosing ferric trichloride with a concentration of 600 mg/L only, COD removal was 17.1%. Therefore,

FIGURE 2.35 Comparison of ferrous sulfate alone and composite coagulation agent made by sodium hypochlorite and ferrous sulfate on COD removal for concentrated liquid from reverse osmosis system with biological pretreatment (COD concentration of influent 427.2 mg/L).

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Pollution Control Technology for Leachate From Municipal Solid Waste

ferric salt produced by composite agents has a higher COD removal than ferric trichloride that was directly dosed.

2.6.4 Comparison of Composite Flocculants and Oxidizing Agents 2.6.4.1 Composite Mixing of Potassium Permanganate and Ferric Trichloride COD removal effect comparison of using potassium permanganate alone and using composite coagulation agent made by potassium permanganate and ferric trichloride is shown in Fig. 2.36. In these experiments, dosage of ferric trichloride was 800 mg/L and COD concentration of leachate used was 427.2 mg/L. Seen in Fig. 2.36, under the same dosage of potassium permanganate, dosage of ferric trichloride with a concentration of 800 mg/L can increase COD removal by 15%
20% for concentrated liquid from reverse osmosis system with biological pretreatment. Compared with using potassium permanganate alone, using composite coagulation agent can increase organics removal by 15%
20% (average value was 17.5%) from leachate. Using ferric trichloride only can obtain an organic removal of 17.5%, which meant composite mixing of potassium permanganate and ferric trichloride produced a mutual superimposition effect. 2.6.4.2 Composite Mixing of Sodium Hypochlorite and Ferrous Sulfate COD removal effect comparison of using sodium hypochlorite alone and using composite coagulation agent made by sodium hypochlorite and ferrous

FIGURE 2.36 Comparison of potassium permanganate only and composite coagulation agent made by ferric trichloride and potassium permanganate on COD removal for concentrated liquid from reverse osmosis system with biological pretreatment (COD concentration of influent 427.2 mg/L).

Physical and Chemical Treatment Processes for Leachate Chapter | 2

83

sulfate is shown in Fig. 2.37. In these experiments, dosage of ferrous sulfate was 800 mg/L and COD concentration of leachate used was 470.8 mg/L. Seen in Fig. 2.38, under the same dosage of sodium hypochlorite, dosage of ferrous sulfate with a concentration of 600 mg/L will decrease COD removal by 7%
15% rather than increasing it from leachate. When dosage of sodium hypochlorite was about 1100 mg/L, COD removals of composite coagulation agent and oxidizing agent differed most. When dosage was 1800 mg/L, the maximal COD removal point existed by using sodium hypochlorite, while no optimal removal point existed by using composite coagulation agent. When using acomposite coagulation agent, COD removal increased all throughout as the dosage of sodium hypochlorite increased. After conducting data fitting, following parabolic equation of sodium hypochlorite dosage and COD removal can be obtained. y 5 2 8E 2 06x2 1 0:0361x 1 13:957

R2 5 0:997

ð2:15Þ

where, y: COD removal (%); x: sodium hypochlorite dosage (mg/L); R2: Square error. According to the parabolic equation, at ferrous sulfate with a concentration of 600 mg/L in composite coagulation agent, the optimal COD removal of 55% can be obtained from leachate. At this condition, dosage of sodium hypochlorite was 2275.8 mg/L. Oxidation of 1 mg ferrous sulfate needs 0.245 mg sodium hypochlorite so that oxidation of 600 mg ferrous sulfate needs 147 mg sodium hypochlorite (0.245 3 600 5 147). Using sodium hypochlorite with a concentration of 147 mg/L can obtain a COD removal of

FIGURE 2.37 Comparison of using sodium hypochlorite only and using composite coagulation agent made by sodium hypochlorite and ferrous sulfate on COD removal for concentrated liquid from reverse osmosis system with biological pretreatment.

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Pollution Control Technology for Leachate From Municipal Solid Waste

FIGURE 2.38 Comparison of using sodium hypochlorite only and using composite coagulation agent on COD removal for concentrated liquid from reverse osmosis system with biological pretreatment.

11.2%, making the corresponding curve move up by 11.2%. The resultant curve is shown in Fig. 2.38. It can be seen in Fig. 2.38 that, after the translation, curve of composite coagulation agent coincides with curve of sodium hypochlorite generally. When sodium hypochlorite concentration was relatively low, composite coagulation agent had a better performance. When sodium hypochlorite concentration corresponded with the optimal COD removal, sodium hypochlorite had a better COD removal. The general trends were coincided. Curve after the translation contains two parts: one is for the removal by all sodium hypochlorite and the other is for COD removal by ferric ions generated in oxidation. The coincident trends showed that composite mixing of sodium hypochlorite and ferrous sulfate produced a negative synergism.

2.6.5 Relationship Between Potassium Permanganate Dosage and Residues In composite coagulation agent, if proportion or dosage of potassium permanganate is too high, residual concentration of potassium permanganate may increase. Therefore, it is necessary to discuss the connection between potassium permanganate dosage and its residual concentration, which is shown in Table 2.17. It can be seen in Table 2.17 that, when potassium permanganate dosage ranges from 100 to 500 mg/L, residual concentration can meet the national water comprehensive discharge standard (5 mg/L). When potassium permanganate dosage was 400 mg/L, it was the most completely consumed, the residual was in its lowest amount. Thus, this dosage was the optimal dosage of potassium permanganate in the composite coagulation agent. As the

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85

TABLE 2.17 Connection Between Potassium Permanganate Dosage and Its Residual Concentrations for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment KMnO4 dosage (mg/L)

100

200

300

400

500

Residual KMnO4 concentration (mg/L)

1.86

1.41

1.03

0.8

1.79

TABLE 2.18 Leachate Characteristics of Aged-Refuse-Based Biofilter Effluent pH

6.5
8

COD

670
850 mg/L

BOD

10
40 mg/L

NH3-N

13
20 mg/L

dosage of potassium permanganate in composite coagulation agent became higher, the color of the filtrate turned from bright yellow to a lighter one, indicating that the decolorization effect of potassium permanganate was increasing, while potassium permanganate dosage was 400 mg/L, it was the most completely consumed, the residual was in its lowest amount, and the color of the filtrate was rather light. Thus, in the composite coagulation agent, when the dosage of iron trichloride was 800 mg/L, the dosage of 400 mg/L can be seen as the optimal dosage of potassium permanganate.

2.7 LEACHATE TREATMENT FROM AGED-REFUSE-BASED BIOFILTER EFFLUENT BY BIOFLOCCULANTS OR COMBINED WITH INORGANIC COAGULANTS Microbial flocculants are biodegradable and the degradation products are harmless to the ecosystem. Some of the microbial flocculants have advantages over other types of flocculants. Landfill leachate treated by bioflocculant or combined with other inorganic coagulants was examined and optimized as an appropriate treatment of mature landfill leachate and compared the removal rate of COD with the traditional flocculants. The leachate was taken from the effluent of the aged-refuse-based biofilter being operated at Shanghai Refuse Landfill as shown in Table 2.18. It was more nonbiodegradable than that of mature leachate from the closed landfills.

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2.7.1 pH Effect pH is one of the important factors governing COD removal by coagulation. Initially, a serials jar test was conducted to study the coagulation effectiveness of different flocculants under various pH conditions. The relationship between the removal and the optimized pH value was illustrated in Fig. 2.39. Different flocculants had their own optimized pH value. For the aluminum sulfate and ferric chloride at dosage of 500 mg/L, the removals increased with pH value and reached the maximum at pH 4.5
5 and 5.5
6 respectively. For the bioflocculant, the optimal pH was 7
8 at the dosage of 50 mg/L. The influence of pH value on removal of COD is related to the mechanism of flocculation. More than one mechanism of humic acid removals appear, mostly referred to charge neutralization, entrapment, bridging, and adsorption. For the aluminum sulfate and ferric chloride, the assumed mechanism of coagulation is charge neutralization, which occurs as a result of the interaction between the highly charged cationic aluminum hydrolysis species and the negatively charged colloid particles. The flocculating mechanism of bioflocculant may be entrapment and bridging and, thus, pH does not have great influence on the removals compared with inorganic flocculants.

2.7.2 Optimal Coagulant Dose Several coagulant doses are tested at the optimal pH as given in Figs. 2.40 and 2.41. As for the maximum COD removal, the optimal flocculant dose was similar for the aluminum sulfate and ferric chloride. With increasing coagulation dose, the removals increased gradually, giving a maximum 50

Removal (%)

40 30 20 10 0 4

5

6

7

8

9

10

pH ferric chloride(500 mg/L)

aluminium sulfate(500 mg/L)

bioflocculant(50 mg/L) FIGURE 2.39 Influence of pH on COD removals under various pH conditions for aged-refusebased biofilter effluent by bioflocculants or combined with inorganic coagulants.

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87

FIGURE 2.40 Influence of aluminum sulfate and ferric chloride dosages on COD removals for aged-refuse-based biofilter effluent by bioflocculants or combined with inorganic coagulants.

FIGURE 2.41 Influence of bioflocculant dosage on COD removal for aged-refuse-based biofilter effluent by bioflocculants or combined with inorganic coagulants.

removal at the dose of 1000 mg/L, over which no COD removal increase can be observed. However, ferric chloride was proved more effective than aluminum sulfate in COD removals. For the bioflocculant, the dosage does not have a great influence on removal rate. Supernatant pH decreased with increase of aluminium sulfate and ferric chloride dose, while no change with increase of the bioflocculant dose, due to the formation of Fe(OH)3 or Al(OH)3 for the former ones. With increasing coagulant dose, the dark brown color of raw leachate turned to a clear brown, then to yellow and became clear yellow.

2.7.3 Synergism of Both the Chemical Flocculants and Bioflocculants Adding bioflocculant or organic flocculant to the leachate after aluminum sulfate and ferric chloride can significantly improve the removal (Fig. 2.42).

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Pollution Control Technology for Leachate From Municipal Solid Waste

FIGURE 2.42 Influence of bioflocculant dosage on COD removals in the presence of Fe and Al flocculants for aged-refuse-based biofilter effluent by bioflocculants or combined with inorganic coagulants.

Bioflocculant, owning to it’s biodegradability, do not bring secondary pollution and can be used as an environmentally friendly agent for leachate treatment

2.8 ADVANCED TREATMENT BY OXIDATION FOR CONCENTRATED LIQUID FROM REVERSE OSMOSIS SYSTEM WITH BIOLOGICAL PRETREATMENT 2.8.1 Fenton Oxidation Fenton reagent is especially suitable for the oxidation of organic wastewater that is difficult to be treated by biological processes. Fenton oxidation technology uses hydrogen peroxide as oxidant and ferrous salt as the catalyst in the homogeneous catalytic oxidation method. The complete degradation of organic compound is done in a short time. Fenton oxidation technology is not restricted by the types, components, and concentration of wastewater, and has been widely used in wastewater treatment. Fenton Oxidation treatment is mainly affected by H2O2 consumption, Fe21 consumption, pH, and reaction time. A single factor test can determine the Fenton oxidation range, and then through orthogonal test and optimizing reaction, it can determine the best reaction conditions. Controlling reaction conditions for pH 4 and FeSO4 14 mmol/L, the leachate was added different amounts of H2O2. After completing the reaction and adjusting pH to about 8, the leachate was added, polymeric ferric sulfate flocculant with stirring and standing, until the pH was about 6. The range of H2O2 dosage was determined by the sedimentation ratio and supernatant COD value. When the H2O2 dosage was less than 81 mmol/L, the treatment efficiency was not stable, and the effluent quality was not good. When the dosage was 900 mmol/L, the effluent COD was 2484 mg/L, with the highest removals (Table 2.19, Fig. 2.43).

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TABLE 2.19 Effect of H2O2 Dosage on COD Removal for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment in Fenton Oxidation No.

H2O2 dosage (mmol/L)

Leachate COD (mg/L)

Effluent COD (mg/L)

Sedimentation (%)

COD removal (%)

27

6276

4732

79

24.6

2

45

6276

4668

70

25.6

3

63

6276

3388

66

46.0

4

81

6276

4004

70

36.2

5

90

6276

4408

74

29.8

6

180

6276

4112

84

34.5

7

270

6276

3892

89

38.0

8

360

6276

4860

90

22.6

9

450

6276

2820

87

55.1

10

540

6276

3568

86

43.1

11

630

6276

3364

75

46.4

12

720

6276

3000

78

52.2

13

900

6276

2484

81

60.4

14

1080

6276

3224

71

48.6

COD removal rate/Settlement ratio (%)

1

90 80 70 60 50 40 COD removal rate (%) Settlement ratio (%)

30 20 0

200

400 600 800 H2O2 (mmol/L)

1000

1200

FIGURE 2.43 Effect of H2O2 dosage on COD removal rate and sedimentation rate for concentrated liquid from reverse osmosis system with biological pretreatment (COD 6276 mg/L, pH 4, FeSO4 14 mmol/L, reaction time 1 h).

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Pollution Control Technology for Leachate From Municipal Solid Waste

The dosage of oxidant is one of the important factors that influence the oxidation efficiency of Fenton reagent. The formation mechanism of the free radical oxidation of organic compounds by Fenton is as follows. Fe21 1 H2 O2 -Fe31 1 UOH 1 OH2

ð2:16Þ

Fe31 1 H2 O2 -Fe21 1 HO2 1 H1

ð2:17Þ

H2 O2 1 UOH-H2 O 1 HO2

ð2:18Þ

Through the effect of Fe21 in the reaction of the excitation and transmission, the chain reaction can be sustained until H2O2 is exhausted. The Eq. (2.16) of production  OH is the initial step in the whole process of the reaction. The Eq. (2.18) is the speed control step.  OH production depends on the concentration of Fe21 and H2O2. It is beneficial to improve the degradation efficiency of organic pollutants by increasing the concentration of Fe21 and H2O2 appropriately. When the sedimentation ratio was consistent with COD change and the dosage was not more than 400 mmol/L, sedimentation ratio and COD removal rate were not stable. When H2O2 dosage was more than 200 mmol/L, the sedimentation effect was better and the ratio was more than 80%. When H2O2 dosage achieved 500 mmol/L, the COD removal rate increased with the increase of H2O2 dosage. When it achieved 800 mmol/L, the increase speed became slow, but the sedimentation ratio and COD removal rate decreased with continuing addition of H2O2. When the H2O2 concentration was too low, with the increase of H2O2 concentrations, the more  OH are generated. Excessive H2O2 oxidize rapidly Fe21 into Fe31 in the reaction, which makes the oxidation reaction happens under the Fe31 catalysis, consuming some H2O2 and inhibiting the  OH generation, thus decreasing COD removal rate. Therefore, the H2O2 range was ranged 450
900 mmol/L. Controlling reaction conditions for pH 4 and H2O2 720 mmol/L, the leachate was added different amount of Fe21. The effects of different Fe21 dosage on COD and sedimentation ratio are shown in Table 2.20 and Fig. 2.44, respectively. From Table 2.20, it can be seen that when the FeSO4 dosage is less than 14 mmol/L, the COD removal rate is lower than 50%. When the amount of FeSO4 was 21 mmol/L, the COD removal rate reached the maximum (65.1%) and the sedimentation was the best and the ratio was 89%. From Fig. 2.44, the change rule of sedimentation ratio and COD removal rate is completely consistent. Both increased with the increase of initial FeSO4 concentration. It began to reduce when it reached 21 mmol/L. Fe21 is the necessary condition for the generation of  OH. Without Fe21, H2O2 is difficult to decompose to generate free radicals. When Fe21 is less, the

Physical and Chemical Treatment Processes for Leachate Chapter | 2

91

TABLE 2.20 Effect of Fe21 Dosage on Fenton Oxidation for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment No.

Fe21 dosage (mmol/L)

Influent leachate COD (mg/L)

Effluent COD (mg/L)

Sedimentation ratio (%)

COD removal (%)

3.5

6316

3588

79

43.2

2

7

6316

3940

70

37.6

3

14

6316

2884

82

54.3

4

21

6316

2204

89

65.1

5

28

6316

2408

86

61.9

6

35

6316

2511

87

60.2

COD removal rate/Settlement ratio (%)

1

90 85 80 75 70 65 60 55 50 45 40 35

COD removal rate (%) Settlement ratio (%)

0

5

10

15 Fe2+

20

25

30

35

40

(mmol/L)

FIGURE 2.44 Effect of FeSO4 dosage on COD removal rate and sedimentation rate for concentrated liquid from reverse osmosis system with biological pretreatment (COD 6276 mg/L, pH 4, H2O2 720 mmol/L, reaction time 1 h).

catalytic reaction is slow, and the reaction becomes faster and faster with the increase of Fe21. When Fe21 is too much, the following reactions occur. Fe21 1 UOH-Fe31 1 OH2

ð2:19Þ

Fe31 1 HO2 -Fe21 1 HO1 2

ð2:20Þ

Fe21 1 HO2 1 H1 -H2 O2 1 Fe31

ð2:21Þ

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Pollution Control Technology for Leachate From Municipal Solid Waste

TABLE 2.21 Effect of pH on Oxidation of Fenton for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment No.

pH

Raw leachate COD (mg/L)

Effluent COD (mg/L)

Sedimentation ratio (%)

COD removal (%)

1

2

6233

2682

79

57.0

2

3

6233

2527

70

59.5

3

4

6233

2416

82

61.2

4

5

6233

2528

82

59.4

5

6

6233

3018

84

51.6

6

7

6233

3631

73

41.7

Excessive Fe21 and H2O2 will become the capture agent of  OH. Therefore, the optimal ratio of Fe21 and H2O2 is very important. FeSO4 dosage range was 7
28 mmol/L. Controlling reaction conditions for Fe21 14 mmol/L and H2O2 720 mmol/L, the pH was controlled before the reaction. From Table 2.21, it can be seen that at initial pH 4, the treatment effect is the best. The effluent COD was 2416 mg/L and the removal rate was 61.2% and the settlement ratio was 82%. Increasing or decreasing the pH value, the effluent COD value increased. When the leachate pH value was below 3, it will generate a large amount of acid, making cost high and without direct significance. From Fig. 2.45, it can be seen that at pH 3
4, the COD removal rate is the biggest and the settlement effect is better. The reason is that the Fenton reagent is under the acidic condition. Fe31 1 H2 O2 -Fe21 1 HO2 1 H1

ð2:22Þ

In the alkaline environment, the solution of two ferric ions and ferric ions will precipitate in the form of hydroxide precipitation and loss of catalytic capacity. Hence, it suppresses the generation of hydroxyl radical and the COD removal rate is decreased in the alkaline condition. UOH 1 e2 1 H1 -Fe21 1 H2 O 1

ð2:23Þ

Higher H concentration in leachate at low pH value will play a role as a hydroxyl radical scavenger, which causes the invalid consumption of hydroxyl free radical and makes catalytic blocked, reduce the oxidation ability of Fenton reagent. Therefore, it is appropriate to control the pH value of 3
4, Fe21 14 mmol/L and H2O2 720 mmol/L before the reaction. Take 200 mL leachate at intervals and adjust pH about 8, and add polymeric ferric sulfate flocculant with stirring and standing, until pH about 6. The reaction

COD removal rate/Settlement ratio (%)

Physical and Chemical Treatment Processes for Leachate Chapter | 2

93

85 80 75 70 COD removal rate (%) Settlement ratio (%)

65 60 55 50 45 40

2

3

4

5

6

7

pH FIGURE 2.45 Effect of pH on COD removal rate and settlement ratio for concentrated liquid from reverse osmosis system with biological pretreatment (COD 6276 mg/L, FeSO4 14 mmol/L, H2O2 720 mmol/L, reaction time 1 h).

TABLE 2.22 Effect of Reaction Time on Oxidation of Fenton for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment No.

pH

Raw leachate COD (mg/L)

Effluent COD (mg/L)

Sedimentation ratio (%)

COD removal rate (%)

1

10

6037

3994

89

33.8

2

20

6037

3213

89

46.8

3

30

6037

2602

77

56.9

4

50

6037

2712

80

55.1

5

70

6037

2680

79

55.6

6

90

6037

2559

74

55.6

is determined according to the sedimentation ratio and supernatant COD value. The effect of reaction time on the oxidation of Fenton is shown in Table 2.22 and Fig. 2.46. From Table 2.22, it can be seen that COD removal rate reaches 56.9% at 30 minutes reaction time for Fenton reagent. From Fig. 2.46, it can be seen that Fenton oxidations occur mainly at the beginning of 30
40 minutes and significantly slows down after 40 minutes reaction, where the organic matters in leachate has been degraded very well, and the concentration of organic pollutants and H2O2 all decrease to a great extent. Therefore, the time range

94

Pollution Control Technology for Leachate From Municipal Solid Waste

COD removal rate/Settlement ratio (%)

90 80 70 60 50 COD removal rate (%) Settlement ratio (%)

40 30

0

20

40 60 Time (min)

80

100

FIGURE 2.46 Effect of reaction time on COD removal rate and settlement ratio for concentrated liquid from reverse osmosis system with biological pretreatment (COD 6276 mg/L, FeSO4 14 mmol/L, H2O2 720 mmol/L, pH 3).

TABLE 2.23 Factors and Levels of Orthogonal Experiments for Fenton System t (min)

pH

7

15

3

600

14

30

4

Level 3

750

21

45

5

Level 4

900

28

60

6

Leachate

H2O2 (mmol/L)

Level 1

450

Level 2

Fe21 (mmol/L)

of Fenton oxidation should be 15
60 minutes. In Fenton system, H2O2 dosage, Fe21 dosage, pH, and reaction time t are the 4 factors controlling the oxidation efficiency and their orthogonal experimental design is given in Tables 2.23 and 2.24. COD of leachate was 6500 mg/L and NH3-N was about 390 mg/L. Fenton oxidation process acts as the reaction between H2O2 and Fe21, producing  OH. The increase of H2O2 concentration and Fe21 concentration as well as the decrease of pH are beneficial to the production of  OH. With the increase of H2O2 concentrations, the COD removal rate and the sedimentation rate increase first and then decrease. With the increase of Fe21 concentrations, the COD removal rate increases, but the sedimentation ratio decreases. pH has little effect on sedimentation ratio. COD removal rate

TABLE 2.24 Orthogonal Experiments for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment Using Fenton System T (min)

Effluent COD (mg/L)

COD removal (%)

Settlement ratio (%)

3

15

3308

48.0

90.0

14

4

30

4530

28.8

86.7

450

21

5

45

3364

47.1

84.2

450

28

6

60

3180

50.0

78.1

5

600

7

4

45

3012

52.6

90.4

6

600

14

3

60

2404

62.2

87.0

7

600

21

6

15

3732

41.3

88.1

8

600

28

5

30

2048

67.8

76.3

9

750

7

5

60

3344

47.4

86.9

10

750

14

6

45

2228

65.0

81.8

11

750

21

3

30

2644

58.4

85.4

12

750

28

4

15

1656

74.0

87.9

13

900

7

6

30

4612

27.5

87.0

14

900

14

5

15

3172

50.1

87.8

15

900

21

4

60

2384

62.5

80.2

16

900

28

3

45

1672

73.7

76.2

Fe21 (mmol/L)

No.

H2O2 (mmol/L)

1

450

7

2

450

3 4

pH

(Continued )

TABLE 2.24 (Continued) No.

H2O2 (mmol/L)

Fe21 (mmol/L)

pH

T (min)

K1i

43.5

43.9

60.6

53.3

K2i

56.0

51.5

54.5

45.6

K3i

61.2

52.3

53.1

59.6

K4i

53.5

66.4

45.9

55.5

Ri

17.7

22.5

14.6

14.0

K1j

84.8

88.6

84.7

88.5

K2j

85.5

85.8

86.3

83.9

K3j

85.5

84.5

83.8

83.2

K4j

82.8

79.6

83.8

83.1

Rj

2.6

9.0

2.5

5.4

Effluent COD (mg/L)

COD removal (%)

Settlement ratio (%)

Physical and Chemical Treatment Processes for Leachate Chapter | 2

97

decreases when the pH increases. The reaction tends to be stable after the reaction time is t . 45 minutes. Fe21 1 H2 O2 -Fe31 1 OH2 1 UOH

ð2:24Þ

According to the results of orthogonal experiments, the H2O2 concentration, Fe21 concentration, pH, and the reaction time of the range R were 17.7, 22.5, 14.6, and 14, respectively. In Fenton oxidation, the primary and secondary order of organic matter treatment were Fe21 concentration, H2O2 concentration, pH and reaction time, and the primary and secondary order of sedimentation ratio was Fe21, reaction time, H2O2 and pH, which was consistent with the single factor data analysis. According to Figs. 2.47
2.50, the optimum reaction conditions of Fenton oxidation were H2O2 concentration 750 mmol/L, Fe21 concentration 28 mmol/L, reaction time 45 minutes and pH 3. Under these conditions, COD was reduced to 1340 mg/L which was lower than the theoretical value of 3000 mg/L. ½FeðH2 OÞ6 31 1H2 O-½FeðH2 OÞ5 OH21 1H3 O1

ð2:25Þ

½FeðH2 OÞOH21 1H2 O-½FeðH2 OÞ4 OH1 1H3 O1

ð2:26Þ

When the pH value is 3
5, the complex becomes:  21  41 2 FeðH2 OÞ5 - FeðH2 OÞ8 ðOHÞ2 12H2 O

ð2:27Þ

½FeðH2 OÞ8 ðOHÞ2 41 1H2 O-½Fe2 ðH2 OÞ7 ðOHÞ3 31 1H3 O1

COD removal rate (%) Settlement ratio (%)

64 60 56 52 48 44 40 400

500

600 700 H2O2 (mmol/L)

800

900

90 89 88 87 86 85 84 83 82 81 80 79 78 77 76 75

Settlement ratio (%)

COD removal rate (%)

68

ð2:28Þ

FIGURE 2.47 Effect of H2O2 concentration on COD removal rate and settlement ratio for concentrated liquid from reverse osmosis system with biological pretreatment using Fenton system.

Pollution Control Technology for Leachate From Municipal Solid Waste

COD removal rate (%)

70

90

COD removal rate (%) Settlement ratio (%)

88

65

86 60 84 55

82

50

80

45

78

Settlement ratio (%)

98

76 40 5

10

15

20

25

30

Fe2+ (mmol/L) FIGURE 2.48 Effect of Fe21 concentration on COD removal rate and settlement ratio for concentrated liquid from reverse osmosis system with biological pretreatment using Fenton system.

COD removal rate (%) Settlement ratio (%)

65

90 88 86

60 84 55

82

50

80

45

78

Settlement ratio (%)

COD removal rate (%)

70

76 40

2

3

4

5

6

7

pH FIGURE 2.49 Effect of pH on COD removal rate and settlement ratio for concentrated liquid from reverse osmosis system with biological pretreatment using Fenton system.

½Fe2 ðH2 OÞ7 ðOHÞ3 31 1½FeðH2 OÞ5 HO21 -½Fe3 ðH2 OÞ7 ðOHÞ4 51 12H2 O ð2:29Þ Fe21 concentration is an important factor affecting COD removal in Fenton oxidation. The increase of H2O2 concentration can improve the removal rate of COD, but when it was higher than 750 mmol/L, the increase of removal rate was not obvious. Continuing to increase, it was counterproductive. The removal efficiency was better when the pH value was between 3 and 4. After 30 minutes reaction, the COD removal rate slowed and the

Physical and Chemical Treatment Processes for Leachate Chapter | 2

90 COD removal rate (%) Settlement ratio (%)

65

88 86

60 84 55

82

50

80

45

78

Settlement ratio (%)

70 COD removal rate (%)

99

76 40 0

10

20

30 40 Time (h)

50

60

70

FIGURE 2.50 Effect of reaction time on COD value and sedimentation ratio for concentrated liquid from reverse osmosis system with biological pretreatment using Fenton system.

reaction was basically completed after 45 minutes. The Fenton reaction system involves a series of complex chemical reactions. There is a filling ratio between H2O2 and FeSO4  7H2O. If the proportion maintains, the reaction can be carried out smoothly with a better treatment effect. It can’t improve the process efficiency by increasing the investment amount of one of them, and cause confusion in the reaction process, resulting in paralysis of the entire Fenton system.

2.8.2 Potassium Ferrate Oxidation Potassium ferrate (K2FeO4) is a new type of multifunctional leachate treatment agent, such as oxidation, adsorption, flocculation, coagulation, sterilization, deodorant, and so on. Potassium ferrate is concerned in recent years as a new type of highly efficient oxidant. Compared with other oxidizing agents, it has the advantages of high redox potential, strong flocculation, and nontoxicity. Therefore, it is applied to the oxidative degradation of As31, H2S, cyanide, ammonia, amines, nitrobenzene, surfactants, bisphenol A, and other inorganic and organic pollutants. After coagulation and sedimentation, the cocentrated liquid from reverse osmosis membrane filtration of leachate contains a large number of complex COD substances, and the COD value is about 6300 mg/L. Ferrate has strong oxidation ability that is better than that of chlorine and ozone. After dissolved in the leachate, it not only can effectively kill microorganisms and algae, but also oxidize and decompose various organic and inorganic pollutants. This test used potassium ferrate to oxidize reverse osmosis concentrated liquid, Fenton oxidation effluent and potassium ferrate-TiO2-UV technology for further oxidizing Fenton oxidation effluent. The leachate pH

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Pollution Control Technology for Leachate From Municipal Solid Waste

TABLE 2.25 Effect of Potassium Ferrate on COD Removal for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment No.

Potassium ferrate investment amount (mg/L)

Raw leachate COD (mg/L)

Effluent COD (mg/ L)

COD removal (%)

1

5

6258

5145

17.8

2

10

6258

4236

32.3

3

15

6258

3346

46.5

4

20

6258

3127

50.0

5

25

6258

3032

51.6

6

30

6258

2987

52.3

COD removal rate (%)

50

40

30

20 5

10

15 20 K2FeO4 (mg/L)

25

30

FIGURE 2.51 Effect of potassium ferrate dosage on COD removal for concentrated liquid from reverse osmosis system with biological pretreatment (COD 6258 mg/L, pH 3, reaction time 5 h).

was adjusted to 3 with concentrated sulfuric acid in beaker coagulation method. Potassium ferrate was added to leachate, stirring 5 hours, adjusting pH to 7 by CaO and settling 1 hours. The COD values by determining supernatant are shown in Table 2.25 and Fig. 2.51. With the potassium ferrate dosage concentration 10 mg/L, the amount of COD removed theoretically was 1.21 mg/L, but the removal of COD in the experiment process was 2022 mg/L, which was 1671 times as much as the

Physical and Chemical Treatment Processes for Leachate Chapter | 2

101

theory. This may be that after potassium ferrate was dissolved in leachate, the Fe(VI) did not translate directly to Fe(III) in the decomposition process. But through 15, 14 intermediate oxidation states, the Fe(VI) gradually reduced to Fe (III). Further, in the process of Fe (VI) reducing to Fe (III), it produced a hydrolysis product in positive valence state. These hydrolysis products may have a higher positive charge and a larger network structure than the trivalent aluminum, ferric, and other hydrolysis products. All kinds of intermediate products in Fe (VI) reducing to Fe (III) appearred polymerization process, and Fe (III) generated form Fe (OH)3 colloidal precipitation quickly. These flocculent colloids with high adsorption activity can adsorb and flocculate most of ions, organic matters and suspended matters in a wide range of pH values. Thus, it increases the adsorption of leachate COD. COD is further removed from the leachate in the subsequent flocculation precipitation so that the actual needs of the potassium ferrate investment is significantly lower than the theoretical dosage. With the increase of the potassium ferrate investment amount, the COD removal rate increased gradually. When the investment amount increased to 15 mg/L, the COD removal rate increased slowly. Continuing to add potassium ferrate, COD removal effect was not obvious. This may be that when potassium ferrate reached 20 mg/L, the flocculation effect had reached the best condition. Continuing to increase potassium ferrate, the removal rate of COD was the result of oxidation of organic matters. When potassium ferrate invested 30 mg/L, the COD removal rate reached the maximum value. The effluent COD was 2987 mg/L, and the effluent COD was higher than that after Fenton oxidation (1400 mg/L). On the one hand, because of the weak oxidation of hydrogen peroxide, a lot of organic matters can be oxidized by hydrogen peroxide but cannot be oxidized by potassium ferrate. The removal effect of potassium ferrate on COD mainly depends on the new generation of Fe (OH)3, while the strong oxidation of potassium ferrate only plays a supporting role. On the other hand, potassium ferrate in aqueous solution is of poor stability, breaking down easily, and the storage period is short, which restricts the promotion and application of potassium ferrate. It has been found that FeO22 4 has obvious thermodynamic instability. It is not possible to solve the problem of instability of FeO22 4 from the fundamental point view. The redox potential of potassium ferrate is lower than that of Fenton, but the different oxidants have a certain selectivity to organic compounds. After oxidized by Fenton, the effluent contains many molecules, and aromatic compounds that is difficult to be oxidized. Leachate was adjusted to pH 3 with concentrated sulfuric acid by beaker coagulation method. Potassium ferrate was added to the leachate, stirring 15 minutes, and adjusting pH to 7 by CaO, finally settling 1 hours. The resultant COD value determined by supernatant are shown in Table 2.26. From Table 2.27, it can be seen that COD change is very small, after putting potassium ferrate into the effluent of Fenton oxidation. The effect was

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Pollution Control Technology for Leachate From Municipal Solid Waste

TABLE 2.26 Effect of Potassium Ferrate on COD for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment After Oxidized by Fenton No.

Potassium ferrate investment amount (mg/L)

Leachate COD (mg/L)

Effluent COD (mg/L)

COD removal (%)

1

5

1328

1256

5.4

2

10

1328

1290

2.9

3

50

1328

1310

1.4

4

100

1328

1275

4.0

5

200

1328

1256

5.4

TABLE 2.27 Effect of Fe(V) on COD Removal Rate of Fenton Oxidation for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment No.

Potassium ferrate investment amount (mg/L)

Influent COD (mg/L)

Effluent COD (mg/L)

COD removal (%)

NH3-N removal (%)

1

5

1315

375

0.2

2.8

2

10

1302

365

1.2

5.7

3

50

1228

308

6.8

20.4

4

100

1270

258

3.6

33.3

5

200

1254

235

4.9

39.3

6

400

1234

185

6.4

52.2

7

800

1211

137

8.1

64.6

8

1500

1037

103

21.1

73.4

9

3000

851

98

35.3

74.7

10

5000

750

95

42.9

75.5

not obvious and COD removal rate was below 10%. The reason is that the removal effect of potassium ferrate on COD depends on the new generation of Fe(OH)3. The strong oxidation of potassium ferrate only plays a supporting role. After Fenton oxidation, the solution of macromolecules, colloids, and suspended matter has been significantly reduced, and the oxidation of potassium ferrate is lower than that of Fenton reagent. After Fenton oxidation, further oxidation need to use a stronger oxidizing agent.

Physical and Chemical Treatment Processes for Leachate Chapter | 2

103

Fe(VI)-TiO2-UV combination system has the ability to strengthen the wastewater treatment. The reaction capacity of Fe (V) is about 3 orders of magnitude higher than Fe (VI). Fe (V) has some free radical properties. Fe (VI) -TiO2-UV system was used to treat Fenton oxidation effluent by beaker coagulation, setting UV light bulb with 10 W power above the middle of each beaker. The leachate COD was 1315 mg/L, and NH3-N 376 mg/L. Added TiO2 10 mg/L into leachate, adjusted to pH 3, and then added potassium ferrate, stirred for 4 hours. finally the pH was adjusted to 7 by CaO and settling 1 hours. Using Fe(V) for the enhancement of COD removal rate by Fenton oxidation was better than that of potassium ferrate oxidation alone as Fe (VI) was reduced to Fe (V) in UV radiation of TiO2 suspension: 2 32 FeO22 4 1 ecb -FeO4

ð2:30Þ

The reaction performance and stability of Fe (V) are related to the degree of its proton. Fe (V) is present in at least 3 protonated forms. The attenuation of Fe (V) is the first order when the pH is 7
3.6 and is the second order in the alkaline pH. The rate of Fe (V) oxide increases with the decrease of pH value. Through the reaction of Fe (V) oxidation of thiourea, cyanide, and the reaction kinetics of sulfur, it is found that Fe (V) reaction capacity is 3 orders of magnitude higher than Fe (VI). Fe (V) has some free radical properties. With the increase of the amount of potassium ferrate investment, COD removal rate is on the rise. When the potassium ferrate dosage was 5 g/L, the effluent COD was 750 mg/L and the removal rate of COD was 42.9%. The removal rate of ammonia nitrogen was almost stable (75%). The removal rate of COD and the amount of potassium ferrate’s investment is basically a linear relationship (Fig. 2.52). The change rate of removal rate of COD is larger than that of NH3-N, and the removal rate of NH3-N increased with the increase of potassium sulfate. When the dosage of potassium sulfate was more than 1 g/L, the removal rate of ammonia nitrogen increased. When the dosage of potassium sulfate was more than 2 g/L, the effluent ammonia nitrogen basically reached stable.

2.8.3 Potassium Persulfate Oxidation As a strong oxidant for chemical oxidation treatment of organic substance (E0 5 2.01V), persulfate attracts more and more attention in recent years. In persulfate activating system, H  exists and participates the oxidation reaction of organics. Metal catalyst, such as Ag1, Fe21 et al, is used to activate peroxydisulfate so that organic pollutants in wastewater can be removed, such as 2,4-dinitrotoluene and trichloroethane. It has been studied to use FeO as a catalyst to degrade polyvinyl alcohol by oxidation of potassium persulfate. Persulfate can be activated easily under heating, UV, and metal catalysis conditions. It can produce SO2 4 radical and HO  radicals that have strong

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Pollution Control Technology for Leachate From Municipal Solid Waste

COD/NH3-N removal rate (%)

80 70 60 50 40 30 20

COD

10

NH3-N

0 0

1000

2000 3000 4000 K2FeO4 (mg/L)

5000

FIGURE 2.52 Effect of potassium ferrate investment on COD and NH3-N removals using Fenton oxidation for concentrated liquid from reverse osmosis system with biological pretreatment (COD 1315 mg/L, NH3-N 376 mg/L, 10 w UV bulb, TiO2 10 mg/L, pH 3.0, reaction time 4 h).

oxidation and has strong treatment capacity of hazardous, refractory organic substances. Potassium persulfate is used for advanced treatment of Fenton oxidation effluent. The test method was constant-temperature oscillation. Ground conical flask was fixed on a constant-temperature box. Controlling reaction conditions at temperature 50 C, pH 3, potassium persulfate 150 mmol/L, potassium persulfate/catalyst 10:1B20:1 and reaction time 10 hours. The COD value of testing leachate was 1354 mg/L. The results are shown in Table 2.28 and Fig. 2.53. Without catalysts, the COD removal rate was 4.9%. When adding catalyst, the effluent COD had obvious reduction and the COD removal rate increased. Adding Fe21 and Ag1 as catalysts, the reaction speed was rapid because liquid diffusion velocity was fast. Although Fe21 has catalytic action, the catalytic efficiency is lower than other catalysts because Fe21 is also a free radical trapper and occurs the following reaction: 31 Fe21 1 SO2 1 SO22 4 -Fe 4

ð2:31Þ

When adding different catalysts or different dosages of the same catalyst, the COD removal rate varied. At Fe0 to S2 O22 8 /catalyst of 10:1, treatment efficiency was best and the effluent COD was 250 mg/L. From Fig. 2.53, it can be seen that organic substances removal efficiency is poor without catalysts. When adding MnO2 and Ag2SO4 as catalysts, the dosage of catalysts almost had no impact on COD removal rate. When adding FeSO4 and FeO as catalysts, the COD removal rate improved as the dosage of catalysts increased. Using Ag2SO4, MnO2, and Fe0 to catalyze potassium sulfate, the effects of Fenton effluent oxidation were preferable. FeO is a kind of sustainable free radical initiator. Considering secondary pollution of MnO2 and high

Physical and Chemical Treatment Processes for Leachate Chapter | 2

105

TABLE 2.28 Effect of Catalyst Types and Dosage on the Removal Rate of COD for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment No.

S2 O22 8 /Catalyzer

10:1

20:1

Types of catalysts

Effluent COD (mg/ L)

COD Removal (%)

Effluent COD (mg/L)

COD Removal (%)

1

Blank

1288

4.9

1289

4.8

2

FeSO4

1114

17.7

1162

14.2

3

0

Fe

250

81.5

310

77.1

4

MnO2

339

75.0

349

74.2

5

Ag2SO4

276

79.6

282

79.2

80

2

2S

O

4

70 60 50 40 30 20 10 0 K2S2O8 : Catalyzer 20:1 SO

4

Fe O

nO

g

M

A

COD removal rate (%)

K2S2O8 : Catalyzer 10:1 K2S2O8 : Catalyzer 20:1

K2S2O8 : Catalyzer 10:1 Bl a

nk

Fe

Catalyzer

FIGURE 2.53 Effect of catalyst types and dosage on the removal rate of COD for concentrated liquid from reverse osmosis system with biological pretreatment (COD 1354 mg/L, reaction temperature 50 C, pH 3, K2S2O8 150 mmol/L, potassium potassium sulfate / catalyst 5 10:1 and 20:1, reaction time 10 h).

cost of Ag2SO4, FeO should be the catalyst for advanced oxidation process of Fenton effluent. In order to promote the removal rate of organic substance in leachate, the desirable range of K2S2O8 dosage was determined. Controlling reaction conditions for K2S2O8: FeO 1:1, reaction temperature 50 C and pH 3, the leachate was added different dosage of K2S2O8 and the COD of leachate was 1362 mg/L. Effect of different dosage of K2S2O8 on COD removal rate and sedimentation rate are shown in Table 2.29 and Fig. 2.54.

TABLE 2.29 Effect of Different Dosage of K2S2O8 on COD Removal Rate for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment K2S2O8 concentration (mmol/L)

100

150

200

Oscillating time (h)

COD (mg/L)

Removal (%)

COD (mg/L)

Removal (%)

COD (mg/L)

Removal (%)

1

813

40.3

624

54.2

552

59.5

2

647

52.5

448

67.1

390

71.4

3

568

58.3

360

73.6

297

78.2

4

501

63.2

330

75.8

255

81.3

5

440

67.7

308

77.4

233

82.9

6

392

71.2

282

79.3

207

84.8

7

360

73.6

270

80.2

203

85.1

8

355

73.9

257

81.1

202

85.2

Physical and Chemical Treatment Processes for Leachate Chapter | 2

107

100 90

COD removal rate (%)

80 70 60 50 40 30 K2S2O8 100 mmol/L

20

K2S2O8 150 mmol/L 10

K2S2O8 200 mmol/L

0 0

2

4 Oscillating time (h)

6

8

FIGURE 2.54 Effect of different dosage of K2S2O8 on COD removal rate for concentrated liquid from reverse osmosis system with biological pretreatment (COD 1362 mg/L, K2S2O8: FeO 10:1, reaction temperature 50 C, pH 3).

As shown in Table 2.29, the COD removal rate improves as the time extends and the dosage of K2S2O8 increases. When reaction time was 8 hours, adding 200 mmol/L K2S2O8 into leachate, the COD value decreased to 201 mg/L. From Fig. 2.54, it can be seen that, in the initial stage, potassium persulfate oxidation reaction is rapid and the COD removal rate increases quickly. With the extended reaction time, the COD removal rate increased slowly. The K2S2O8 content was higher, and the reaction went on more quickly. When K2S2O8 content was 200 mmol/L, the process reacted 5 hours to reach the stability status. When K2S2O8 content was 100 mmol/L, the process reacted 7 hours and largely terminated. Higher dosage of K2S2O8 led to the bigger COD removal rate. When the dosage of K2S2O8 increased from 150 mmol/L to 200 mmol/L, COD removal rate had no apparent increase. Therefore, the dosage of K2S2O8 was controlled in 150 mmol/L or so and reaction time was set to 7 hours. Using FeO as a catalyst, K2S2O8 oxidation efficiency can be obviously improved. Controlling reaction conditions for K2S2O8 150 mmol/L, pH 3, reaction time 7 hours, and reaction temperature 50 C, leachate was added different concentration of FeO. The results are shown in Table 2.30 and Fig. 2.55. The value of COD obviously decreased as the proportion of K2S2O8 and FeO increased. When fixing the dosage of K2S2O8, the proportion of K2S2O8 and FeO increased. It meant that FeO content decreased and activated SO2 4 content decreased, so COD removal rate accordingly decreased. The COD removal rate increased as the proportion of K2S2O8 and FeO decreased. When the proportion of K2S2O8 and FeO was below 15, the COD

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Pollution Control Technology for Leachate From Municipal Solid Waste

TABLE 2.30 Effect of K2S2O8: FeO on COD Removal Rate for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment No.

K2S2O8: FeO

Inffluent COD (mg/L)

Effluent COD COD (mg/L)

COD removal (%)

1

5

1358

3588

81.2

2

10

1358

3940

81.4

3

15

1358

2884

79.6

4

20

1358

2204

77.1

5

25

1358

2408

64.7

6

30

1358

2511

56.3

COD removal rate (%)

90

80

70

60

50

5

10

15 20 K2S2O8: FeO

25

30

FIGURE 2.55 Effect of K2S2O8: FeO on COD removal rate for concentrated liquid from reverse osmosis system with biological pretreatment (COD 1358 mg/L, K2S2O8 150 mmol/L, pH 3, reaction time 7 h, reaction temperate 50 C).

removal rate increased as FeO content increased (Fig. 2.55). Therefore, the proportion of K2S2O8 and FeO was controlled in 10 or so. Controlling reaction conditions for K2S2O8: FeO 5 10, K2S2O8 150 mmol/L, reaction temperature 50 C, and reaction time 7 hours, the effects of different pH on COD removal rate are shown in Table 2.31 and Fig. 2.56. The COD removal rate obviously decreased as the pH increased. When pH value was below 3, the reaction can effectively occur. When pH was 2, COD removal rate reached the maximum value of 84.3%.

109

Physical and Chemical Treatment Processes for Leachate Chapter | 2

TABLE 2.31 Effect of pH on COD Removal Rate of Fenton Oxidation for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment No.

pH

Inffluent COD (mg/L)

Effluent COD (mg/L)

COD removal (%)

1

2

1348

212

84.3

2

3

1348

248

81.6

3

4

1348

671

50.2

4

5

1348

1140

15.4

5

6

1348

1151

14.6

COD removal rate (%)

90 80 70 60 50 40 30 20 10 2

3

4 pH

5

6

FIGURE 2.56 Effect of pH on COD removal rate for concentrated liquid from reverse osmosis system with biological pretreatment (COD 1348 mg/L, K2S2O8: FeO 5 10, K2S2O8 dosage 150 mmol/L, reaction temperature 50 C, reaction time 7 h).

pH value has a significant impact on potassium persulfate oxidation of organic substances. The production of peroxydisulfate radical can accelerate by increasing solution acidity. It also proves that potassium sulfate oxidation reactions depend on peroxydisulfate radical to militate and, therefore, optimum pH value should be controlled in 3 or so. In the chemical processes design, reaction temperature is a significant factor for influencing reaction rate, reaction efficiency, economic interests, and technical feasibility. Controlling reaction conditions for K2S2O8: FeO 5 10, K2S2O8 150 mmol/L and pH 3 and reaction time 7 hours. The effects of different reaction temperature on COD removal rate are shown in Table 2.32 and Fig. 2.57. S2 O28- 1H 1 -HS2 O2 8

ð2:32Þ

2 22 1 HS2 O2 8 -SO4 1 SO4 1 H

ð2:33Þ

TABLE 2.32 Effect of Different Reaction Temperature on COD Removal Rate for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment No.

Temperature ( C)

35

45

50

60

Constant temperature oscillation (h)

Inffluent COD (mg/L)

COD removal (%)

Inffluent COD (mg/L)

COD removal (%)

Inffluent COD (mg/L)

COD removal (%)

Inffluent COD (mg/L)

COD removal (%)

1

1

1173

13.2

909

32.7

619

54.2

536

60.3

2

2

1085

19.7

725

46.3

444

67.1

367

72.8

3

3

1063

21.3

544

59.7

357

73.6

296

78.1

4

4

939

30.5

457

66.2

327

75.8

267

80.2

5

5

907

32.9

366

72.9

305

77.4

254

81.2

6

6

886

34.4

332

75.4

280

79.3

235

82.6

7

7

863

36.1

313

76.8

267

80.2

224

83.4

8

8

840

37.8

305

77.4

255

81.1

219

83.8

Physical and Chemical Treatment Processes for Leachate Chapter | 2

111

100 90 COD removal rate (%)

80 70 60 50 40 30

T = 35ºC T = 45ºC T = 50ºC T = 60ºC

20 10 0 0

1

2

3

4 5 6 Oscillation time (h)

7

8

9

FIGURE 2.57 Effect of different reaction temperature on COD removal for concentrated liquid from reverse osmosis system with biological pretreatment (COD 1352 mg/L, K2S2O8: FeO 5 10, K2S2O8 150 mmol/L, pH 3, reaction time 7 h).

Higher COD removal can be obtained when the reaction temperature increased, which was beneficial for persulfate to stimulate free radical formation and to oxidize organic substances. COD removal rate reached the maximum (83.8%) at reaction temperature of 60 C. COD removal rate was 5% by heating the leachate to 50 C, and increased to 38% by adding FeO equals to S2 O22 8 /catalyst 10:1, and further to 81% by adding catalysts (K2S2O8: FeO 5 10) and heating to 50 C. This illustrates that catalysts and reaction temperature are significant factors for the formation of peroxydisulfate free radicals resulting in the synergistic effect of catalysts and reaction temperature on peroxydisulfate. The reaction rate and efficiency increases as the temperature rises (Fig. 2.54). Persulfate is relatively stable in normal temperature. Upon heating, UV and metal oxidation, persulfate can be activated to produce SO2 4  (E0 5 2.6V) that is stronger than peroxydisulfate. - 1 heat=UV-2SO2 S2 O24ðaqÞ 4

ð2:34Þ

When reaction temperature was below 40 C, the oxidation rate was slower within 5 hours, and the reaction rate increased at reaction temperature above 55 C. After 6 hours reaction, COD removals reached the maximum values. Over 45 C may be required for the deep re-oxidation of leachate by potassium persulfaten. GC-MS and infrared spectrum were conducted for the leachate and found that aromatic and aliphatic compounds, some phenols, and esters were present. Molar absorption index (280 nm) and E2/E3 are common parameters of reflecting characteristics of inclusions in aqueous solution. Molar absorption

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Pollution Control Technology for Leachate From Municipal Solid Waste

TABLE 2.33 Spectral Properties for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment Wavelength

Before Fenton oxidation

After Fenton oxidation

After potassium sulfate oxidation

E280

1.5

1.2

0.2

E250/E365

2.5

3.4

8

E465/E665

1.1

1.4

2.7

index can be used to represent the value of DOC and it roughly represents aromatic degree, humification degree, and molecular size. E2/E3 represents the wavelength ratio in 250 nm absorbance and 365 nm absorbance. E4/E6 represents the wavelength ratio in 465 nm absorbance and 665 nm absorbance. They can be used to indicate the sources and structures of natural organic matters. The test data of trinomial index is shown in Table 2.33. 1. E280 nm Although the 280 nm absorbance can’t perfectly represent the structure of all aromatic compounds in humus samples, π-π electron transition of phenolic aromatic hydrocarbon, benzoate and derivatives polyene, and polycyclic aromatic hydrocarbons occurs in the 270
280 nm UV range. These groups are important component units of humus, as it is generally recognized that soluble matter in natural waters are mostly humus, and can roughly represent aromatic degree, humification degree, and molecular size. The bigger value of E280 belongs to higher content of aromatic compounds in soluble matters. For the 280 nm absorbance, with chemical oxidation keeping on, the 280 nm absorbance appeared a decreasing trend. It was the result that all kinds of readily fermentable matters in leachate were decayed, aromatic substance concentration decreased. The absorbance of leachate with potassium persulfate oxidation decreased faster than Fenton oxidation. Fenton oxidation had good effect on degradation of organic macromolecules and potassium persulfate oxidation had good effect on degradation of aromatic substances. 2. E250/E365 E2/E3 can laterally reflect the agglomeration degree and molecular size. For the ratio of E250/E365, it was opposite to the change trend of E280 nm. Through two-stage oxidation, the ratio of E250/E365 reflected an upward trend. The trend of E250/E365 ratio negatively correlated with DOC agglomeration degree. It explained that through Fenton oxidation and potassium persulfate, agglomeration degree in leachate decreased and molecular size reduced.

Physical and Chemical Treatment Processes for Leachate Chapter | 2

113

3. E465/E665 The leachate as treated by coagulation and air stripping was further degraded by two-stage oxidation (Fenton oxidation-Potassium persulfate oxidation), the ratio of E465/E665 in leachate effluent reflected an upward trend. It explained that leachate evolved into small molecular weight, decreasing aromaticity degree, and simplistic structure. It was in accordance with the E280, E250/E365 reaction results.

2.8.4 Potassium Permanganate Permanganate is a strong oxidant, and can degrade many kinds of organic pollutants in leachate, and well react with Fe21, Mn21, S2, CN2, phenol, and other organic compounds with odor. While choosing appropriate dosage, it can kill a lot of algae and microbes, and make the effluent have no smell. The main raw materials of potassium permanganate production are potassium hydroxide and manganese dioxide ore. First make potassium manganate with these materials, and then electrolyze in alkaline solution of potassium manganate to get potassium permanganate. The products of organic pollutants after the oxidation of potassium permanganate are carbon dioxide, alcohols, aldehydes, ketones, hydroxyl compounds, etc. These products are “nonthree—induced substances,” so it can ensure the toxicological safety of the leachate. Table 2.34 shows that potassium permanganate has a good effect on the removal of leachate chroma, and with the increase of potassium permanganate dosage, the chroma removal efficiency increases. Fifty percent leachate chroma can be removed by 100 mg/L potassium permanganate, and increased to 87.5% at 500 mg/L potassium permanganate. However, it must

TABLE 2.34 Relationship Between Oxidants Dosage and Chroma of Leachate (Raw Leachate Chroma 200 Times) Potassium permanganate

Sodium hypochlorite

Dosage (mg/L)

Chroma of effluent

Dosage (mL/L)

Chroma of effluent

100

100

10

4

200

60

15

3

300

40

20

2

400

35

25

1

500

25

30

1

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FIGURE 2.58 Relationship between the removal rate of COD for concentrated liquid from reverse osmosis system with biological pretreatment and potassium permanganate dosage (leachate COD 596.8 mg/L).

be pointed out that when potassium permanganate is over-dosed, the chroma of leachate may increase due to its own color. Potassium permanganate could reduce the chroma of leachate, as it can destroy the organic pollutants in chromophore group. However, it cannot completely destroy the aromatic ring structure of the organic matter. Data show that potassium permanganate has better influence on the decolorization of leachate-soluble organic pollutants in acidic conditions. From Fig. 2.58, it can be seen that potassium permanganate can remove the organics from leachate with a linear relationship between the removal of COD and potassium permanganate dosage. y 5 0:3004x

ðx , 500 mg=LÞ

R2 5 0:9563

ð2:35Þ

where, y: Removal of COD (mg/L); x: Dosage of Potassium permanganate (mg/L); R2: Square error. According to the linear relationship, the reaction between potassium permanganate and organic matter is the first order reaction. 1 mg potassium permanganate can react with 0.3 mg COD. Potassium permanganate has a certain oxidation capability in different media, and can react with strong reducing agent. In acidic medium, the reduction product is Mn21, with pale pink. In the neutral medium, the reduction product is MnO2, with a brownish black precipitation. In the alkaline medium, the reduction product is MnO22 4 , with green color. Potassium permanganate can be used to decompose the organic matters in the leachate. However, it is not able to oxidize all of the organic matters to the final product
carbon dioxide and water
but oxidize some complex organic matters (such as humic acids) into a small amount of molecular weight organic matters. Potassium permanganate is a strong oxidizing agent

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in acid medium, but is a weak oxidant in the neutral and alkaline medium. In strong acid (such as sulfuric acid), potassium permanganate fades soon. In the weak acid (such as acetic acid), potassium permanganate fades slowly. The pH value of leachate is neutral or weak alkaline, and the oxidization product is black manganese dioxide precipitation. Tests showed that, under the neutral condition, the removal rate of organic pollutants by potassium permanganate was much higher than that under the condition of acid and alkali. Potassium permanganate’s biggest characteristic under neutral conditions is to form manganese dioxide. In the nonacid system, MnO2 is more stable. The manganese dioxide solubility in leachate is very low, so the products form as colloidal hydrated manganese dioxide precipitate from the leachate, which results to a very high efficiency in the neutral conditions to remove micro pollutants. The roles of manganese dioxide has two aspects. On one hand, manganese dioxide is a catalyst for oxidation of many organics, and has been proven that it has a significant effect on the catalytic activity of potassium permanganate for the oxidation of organic matters. On the other hand, the new generation of colloidal hydrated manganese dioxide has a large surface area, and could adsorb organic matters in leachate. Therefore, the majority of hydrated manganese dioxide should have both catalytic oxidation and adsorption effects on pollutants. For some easily oxidized contaminants, the effect of catalytic oxidation is bigger in the removal of contaminants; but, to a certain kind of refractory pollutants, the adsorption effect on removal is larger. For example, for alkanes, due to its saturation, is difficult to be oxidized by potassium permanganate, especially for long-chain alkanes. In acidic and alkaline conditions, although potassium permanganate has the strongest oxidation resistance under acidic conditions, potassium permanganate has a bad effect on the removal of long-chain alkanes. Nevertheless, in the neutral conditions, the removal rate of these alkanes by potassium permanganate is very high due to the high content of organic matter in leachate and catalytic reactions of manganese dioxide. Thus, the oxidation rate is improved, and the adsorption capacity is enhanced. The contact time between potassium permanganate and organic matters in leachate is an important parameter, which is a slow process and only in sufficient time can the reaction be complete. Potassium permanganate concentrations are measured using the characteristic absorption wavelength in 520 nm and Fig. 2.59 shows the standard curve between the concentrations of potassium permanganate and its absorbance at 520 nm. The linear formula of the standard curve is obtained by data fitting. y 5 0:0132x 1 0:0014

R2 5 0:9996

where, y: absorbance A; x: Potassium permanganate concentration, mg/L; R2: Square error.

ð2:36Þ

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FIGURE 2.59 Potassium permanganate standard curve in aqueous solutions.

FIGURE 2.60 Relationship between contact time and potassium permanganate’s remaining amount for concentrated liquid from reverse osmosis system with biological pretreatment.

Square error is 0.9996, showing the excellent correlation of the standard curve. As a result, the residual concentration before and after potassium permanganate oxidations can thus be calculated. Fig. 2.60 shows that the longer contact time between potassium permanganate and leachate will lead to the lower residual amount of potassium permanganate in leachate. After reaching a certain contact time, the change of the residual amount of potassium permanganate tended to be gentle, and if it extended the contact time, the amount of residual change became negligible. Potassium permanganate could be effectively used while the contact time was over 1.5 hours. If potassium permanganate dosage is too high, it may cause the excess concentration. Therefore, it is necessary to explore the relationship between the amount of potassium permanganate and its residual concentrations as given in Table 2.35. The second grade of national wastewater discharge standards sets that potassium permanganate’s emissions must be below 5 mg/L. While potassium permanganate dosage was in the range of 100
500 mg/L,

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TABLE 2.35 Relationship Between the Dosage and the Surplus of Potassium Permanganate for Concentrated Liquid From Reverse Osmosis System With Biological Pretreatment KMnO4 dosage (mg/L)

100

200

300

400

500

KMnO4 surplus (mg/L)

1.64

1.33

1.18

1.03

0.73

FIGURE 2.61 Relationship between potassium permanganate’s dosage and consumption for concentrated liquid from reverse osmosis system with biological pretreatment.

the surplus amount of potassium permanganate was all lower than this discharge standard value. When potassium permanganate dosage was more than 600 mg/L, there was a clear color interference, and potassium permanganate surplus was also significantly increased. When the amount of potassium permanganate dosage was 500 mg/L, it was nearly all consumed. Hence, 500 mg/L was the best dosage. Fig. 2.61 shows that when the contact time between potassium permanganate and leachate is enough, the consumption of potassium permanganate is proportional to the amount of dosage, meaning that the reaction between the organic matter and potassium permanganate in leachate is complete and the potassium permanganate dosage is not excessive as to the reaction degree of organic matters. Potassium permanganate’s dosage 500 mg/L was the most appropriate dosage, as it’s remaining amount was the lowest for the tested leachate.

2.8.5 Sodium Hypochlorite It can be seen from Fig. 2.62 that sodium hypochlorite has good effect on removal of organics for concentrated liquid from reverse osmosis system

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FIGURE 2.62 Relationship between concentration of sodium hypochlorite and COD removal rate for concentrated liquid from reverse osmosis system with biological pretreatment (leachate COD 5 596.8 mg/L).

with biological pretreatment. With the increasing dosage of sodium hypochlorite, the removal rate of COD grew gradually. When the sodium hypochlorite dosage increased to 25 mL/L, the removal rate of COD reached the highest (about 63%). When the sodium hypochlorite dosage was more than 25 mL/L, the removal rate of COD was no longer increased. In fact, it is the effective chlorine concentration in sodium hypochlorite that acts in the oxidization role. The available chlorine concentration of sodium hypochlorite solution with the oxidation capacity of each l L solution is equal to 2 times per liter of solution containing chlorine of sodium hypochlorite in quantity. After conversion, 1 g is equal to 0.953 g of the effective chlorine of sodium hypochlorite. According to this conversion relationship, due to the data fitting to the first four points and the origin, the relationship between the COD removal and sodium hypochlorite dosage is thus presented in Fig. 2.62. y 5 0:3082x

ðx , 1300 mg=LÞ R2 5 0:928

ð2:37Þ

where, y: COD removal (mg/L); x: dosage of sodium hypochlorite (mg/L); R2: square error. The linear relationship showed that the sodium hypochlorite and organic reaction was a first-order reaction, each 1 mg sodium hypochlorite can react with 0.3 mg COD. Similar to potassium permanganate reaction, it is inferred that organic compounds that can be oxidized to carbon dioxide by potassium permanganate in RO effluent leachate was homogeneous compared with

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FIGURE 2.63 Relationship between sodium hypochlorite and COD removal rate for concentrated liquid from reverse osmosis system with biological pretreatment.

fresh leachate. The quantity of 1 mol sodium hypochlorite is 74.5 g, and transfer 2 mol electron. One mol potassium permanganate’s quantity is 158 g, and transfer 3 mol electron. Therefore, when the dosage of sodium hypochlorite and potassium permanganate are less than 500 mg/L, under the same dosage condition, the electronic transfer of sodium hypochlorite is about 1.5 times that of potassium permanganate [(158/74.5) 3 (2/3)], but in fact the amount of COD removed is the same. Therefore, it can be referred that in the dosage of sodium hypochlorite, there may be more than 0.5 times the amount of potassium permanganate used to convert high molecular weight organic matter to low molecular one (Fig. 2.63). In the process of hypochlorite ion reduction, it is very easy to get electrons and have very strong oxidation. Combined hypochlorite ion and hydrogen ion in solution, a small neutral molecular state (Hypochlorite HOCl) is present. HOCl has a very strong oxidation. Therefore, the oxidation ability of HOCl is much stronger than that of hypochlorite. Its standard electrode potential is 1.49V, but the standard electrode potential of the hypochlorite is 0.9V, lower than 0.59V by HOCl. The hypochlorite is easily spread to a negatively charged bacterial surface, and penetrate through the cell wall to the inside of the bacteria. The strong oxidizing property of the hypochlorite inside the bacteria can destroy the bacterial enzyme system, which leads to cell death. This is the reason why sodium hypochlorite has a good disinfection effect. In addition, not only in acidic solution, but also in alkaline solution can sodium hypochlorite keep the performance of strong oxidant, and at the same time have a good bactericidal effect.

2.8.6 A Comparison Between Potassium Permanganate and Sodium Hypochlorite Effect on COD Removals When the amount of potassium permanganate dosage is more than 500 mg/L, the effluent is red, which shows the excess of potassium

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permanganate; therefore, the removal rate of COD by potassium permanganate is no more than 30%. The highest removal rate of COD by sodium hypochlorite is less than 70%. It can be shown that when the dosage is more than 500 mg/L, the sodium hypochlorite effect on leachate is obviously higher than that of potassium permanganate. Both the same as oxidant, removal effect is obvious different. The possible reasons for this situation are as follows.

2.8.6.1 Different Electrode Potential Sodium hypochlorite and chlorine have the same disinfection effect, and the mechanism is that sodium hypochlorite in solution after hydrolysis generates hypochlorite ions and contains a positively charged monovalent chloride C11, C11 is then transformed into a negatively charged chloride ions C12 after getting two electrons. In the process of reduction of hypochlorite ion, it is very easy to get electrons and have very strong oxidation. In the solution, hypochlorite ion combined with hydrogen ion, presents a small neutral molecular state. OC12 1 H2 O 1 2e 5 Cl2 1 2OH2

ð2:38Þ

The standard electrode potential of OC2/C12is 0.89V. Potassium permanganate’s outstanding chemical properties are strong oxidizing properties, and under different acid and alkali conditions, the oxidation ability is very different. The leachate shows weak alkaline environment, and potassium permanganate generates manganese dioxide in neutral or weak alkaline environment. According to the above various circumstances, its standard electrode potential is 0.588V. 1 MnO2 4 1 4H 1 3e 5 MnO2 ðsolidÞ 1 2H2 O

ðE 5 1:695VÞ

ð2:39Þ

1 21 MnO2 1 4H2 O 4 1 8H 1 5e 5 Mn

ðE 5 1:510VÞ

ð2:40Þ

2 MnO2 4 1 2H2 O 1 3e 5 MnO2 ðsolidÞ 1 4OH 22 MnO2 4 1 e 5 MnO4

ðE 5 0:588VÞ

ðE 5 0:564VÞ

ð2:41Þ ð2:42Þ

Therefore, potassium permanganate’s standard electrode potential is lower than 0.302 V of sodium hypochlorite. According to the electrochemical theory, the ion concentration of the solution has an effect on the electrode potential, but it is generally similar to the standard electrode potential and the fluctuation is little. The higher the electrode potential of ions, the stronger the oxidation ability, so the sodium hypochlorite oxidation capacity is higher than potassium permanganate. Potassium permanganate can destroy the chromophore group, but it cannot completely destroy the structure of the aromatic ring, which leads to the incomplete degradation of organic matter. According to the characteristics of the oxidation of organic matter by potassium permanganate, it can be considered to add potassium permanganate

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first, destroy the structure of organic matter in leachate, then add the purification agent, and compress leachate colloid double layer, which makes the colloidal particles easily condense into larger ones, easy to deposit. Therefore, sodium hypochlorite has better effect on the organic removal rate of leachate than potassium permanganate in the aspect of electrode potential.

2.8.6.2 Different pH in Solution The pH value of the leachate was about 9. After adding potassium permanganate, there was no change in pH value and the dosage had no effect on the change of pH. However, after adding sodium hypochlorite in the leachate, the pH value of leachate was obviously increased, and with the increase of sodium hypochlorite dosage, pH increased (Table 2.36). Using NaOH, which is alkalinity and nonoxidizable, measure the COD under different pH conditions, it is shown that there is a decreasing relationship between pH and COD of the leachate. The higher pH value led to the lower COD (Table 2.37). A lot of sulfuric acid is still contained in the waste liquid produced by measuring COD, which indicates that high pH does not affect the determination of COD. Sodium hypochlorite dosage caused pH increase of leachate; therefore, in addition to the oxidation on the removal of COD, the increased alkalinity also decreased the COD. The dual function makes the removal rate of COD by sodium hypochlorite higher than potassium permanganate. Generally speaking, the oxidation of sodium hypochlorite is strong acidic medium, and TABLE 2.36 pH of Leachate (Influent pH 5 9) in the Presence of Potassium Permanganate and Sodium Hypochlorite As Oxidants Potassium permanganate dosage mg/L

100

200

300

400

500

pH

8.91

8.89

8.90

8.88

8.90

Sodium hypochlorite dosage mL/L

10

15

20

25

30

pH

9.47

9.69

9.84

9.97

10.04

TABLE 2.37 Corresponding Relationship Between pH Value and COD (NaOH Treatment) of Leachate pH

8.5

10

11

12

COD (mg/L)

508.7

503.2

496.3

465.0

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weak in alkaline medium. However, if the reaction is carried out in a strong acidic medium, NaClO will be lost due to decomposition. Compared with that in weak acidic or weak alkaline medium, the dosage of sodium hypochlorite will be increased. The specific equation for the decomposition reaction of HClO in acid medium is 3HClO 5 2HCl 1 HClO3

ð2:43Þ

In addition, the treated leachate must be adjusted to neutral for discharge. If the processing is in the medium of hydrochloric acid, more alkali must be used to neutralize, which will increase the treatment cost. Therefore, while leachate is in weak alkali conditions, adding sodium hypochlorite treatment is reasonable. The cause of the decrease of COD caused by the increase of pH may be due to the presence of a lot of reducing metal ions, such as Fe21, Mn21, etc., in leachate, with an average concentration of more than 2000 mg/L. In the determination of COD, these reducing substances themselves or the formation of an organic complex is a contribution to the COD, and the pH value increases, makes the reduction of metal ions and hydroxyl binding, and settle down in the form of a hydroxide, consequently to reduce the COD.

2.9 FE-C MICROELECTROLYSIS ADVANCED TREATMENT FOR MBR BIOLOGICALLY DERIVED LEACHATE The sample used was taken from the MBR effluent storage container at a municipal solid waste landfill, a typical valley type landfill site, in Suzhou, China. The specific process for leachate treatment is shown in Fig. 2.64, which includes homogeneous mixing pool to mix the fresh leachate from incineration plant and leachate from landfill for enhancing the biodegradability, and an advanced nitrogen removal A/O system connecting to the A/O/O system to make the whole process an “one-stage A/O/O 1 two-stage A/O 1 external ultra-filtration” bioreactor for enhancing the nitrogen removal ability. This advanced membrane treatment system is actually a parallel combination of a two-section RO system with a treatment capacity of 650 m3/d and a nanofiltration system with a treatment capacity of 300 m3/d. Leachate production rate of RO system and nanofiltration system are 80% and 85%, respectively, and effluents are mixed before discharging to meet the discharge standard regulated in GB16889
2008 (COD , 60 mg/L, BOD5 , 20 mg/L, NH3-N , 8 mg/L, TN , 20 mg/L). During the pilot-scale experiment of bio-treated leachate treatment using Fe-C microelectrolysis, leachate quality of the MBR effluent in landfill shown in Table 2.38 was recorded. In this process, caustic alkali was dosed to control the alkalinity in the biochemical pond. Thus, concentrated sulfuric acid was adopted to control pH value in a range of 6.3
6.7 before the MBR effluent entered nanofiltration system and RO system. For leachate quality of MBR effluent, COD concentration fluctuated from 763 to 1204 mg/L while BOD5 concentration ranged from 40 to 100 mg/L. A low NH3-N

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FIGURE 2.64 Leachate treatment process at a municipal solid waste landfill in Suzhou.

TABLE 2.38 Leachate Quality of MBR Effluent in a Treatment Plant in Suzhou pH

COD

BOD5

NH3-N

TN

SS

6.3
6.7

763
1204

40
100

5.7
23.9

85
238

17
19

concentration ranging from 5.7 to 23.9 mg/L was obtained because nitrification and denitrification were enhanced by connecting the A/O system to the biochemical treatment unit. However, a relatively high TN concentration ranging from 85 to 238 mg/L was obtained in the MBR effluent, which was due to the accumulation of nitrite nitrogen in biological treatment unit caused by a lack of carbon source.

2.9.1 74 µm Magnetized Fe-C Microelectrolysis Process The structure of a magnetized Fe-C microelectrolysis device is shown in Fig. 2.65. It was made up of three 500 L PE leachate buckets (numbered

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FIGURE 2.65 Pilot-scale test for leachate treatment by Fe-C microelectrolysis powderactivated carbon device.

TABLE 2.39 Parameters of Magnetization Device for Leachate Treatment No

Device name

Product model

Specifications

Quantity

1

Corrosion-resistant pump

52FP-11

N 5 0.75 kW, Q 5 4 m3/ h, H 5 11 m

2

2

Three-phrase asynchronous motor

Y90L-4

N 5 1.5 kW; 1390 r/min

1

3

Air compressor

2.5Hp

Q 5 0.08 m3/min, P 5 0.8 MPa, N 5 1.8 kW

1

4

Submersible leachate pump

V250

N 5 0.25 kW, Q 5 9 m3/ h, H 5 7.5 m

1

5

Dosing metering pump

CONC1602

N 5 24 W, Q 5 1.5 L/h, P 5 1.6 MPa

2

bucket1#, 2#, and 3#), two 100 L PE agents buckets, a submersible leachate pump, two acid and alkaline resistant leachate pumps (numbered pump 1#, 2#), two metering pumps (numbered dosing pump 1#, 2#), an electromotor, an air compressor, and five control valves. All connecting pipes with an aperture of DN25, made of PE, were acid and alkaline resistant. Detailed parameters of magnetization device are shown in Table 2.39. In the magnetization device, the main reactor was a 500 L PE bucket with a bar magnet connected to an electromotor inside. The bar magnet was actually multiple ring magnets, with about 8 cm spacing, and fixed to a PE bar. The magnets were coated by an acid-resistant plastic layer, which can facilitate demount of reacted adsorptive iron powders. An electromotor, which provided power for rotary movement of demountable bar magnet, was connected to the top of the magnet. Three stirring impellers were fixed to the

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bottom of bar magnet. An air compressor was connected to corundum aerator, with a diameter of 18 cm for aeration and air diffusion. Super fine iron powders were attached to the bar magnet to be magnetized and the granular activated carbon was dosed directly into the leachate. In the reaction, magnetized super fine iron powders would be oxidized gradually in the surface. The magnetism of organics-contained iron powders decreased, which may cause a fall and final sedimentation of iron powders. Meanwhile, attachment of organics to superfine iron powder would increase volume, causing conglobation and sedimentation. At that time, new superfine materials should be complemented proportionally. Demount filter layer was set in the leachate outlet. Leachate in ultrafiltration outlet was pumped into PE bucket 1# by submersible leachate pump. Under stirring condition, the automatic dosing device was switched on. Industrial sulfuric acid with a concentration of 50% dosed by metering pump was used to adjust acidity of leachate, maintaining pH of sample in a range of 2.5 6 0.2. The sample was then pumped into the main reactor, PE bucket 2#, by acid and alkaline resistant leachate pump. The electromotor and air compressor were turned on, keeping the motor speed at 300 r/min approximately. The air compressor was connected to the corundum aerator with a diameter of 18 cm for aeration and air diffusion in the sample. Meanwhile, superfine iron powders were attracted to the surface of the magnet and granular activated carbon was dosed into the sample directly. After running for a while, the electromotor and air compressor were turned off and the sample let stand for 2 hours. Then, supernatant was pumped into PE bucket 3# by acid and alkaline resistant pump. Under slow stirring condition, automatic dosing device was on and sodium hydroxide with a concentration of 5% was dosed by metering pump, adjusting pH in a range of 9.0 6 0.2. After a standing time of 0.5 hours, physicochemical properties of supernatant in the sample were tested. The control valve subsequently was turned on and the treated sample was discharged to the nanofiltration storage tank through the outlet with a filter cloth. At last, the gauzed was demounted and sludge discharged, as shown in Fig. 2.66. To get 200 mL MBR effluent and adjust pH to about 2.0, activated carbon was dosed 2.0 6 0.02 g and then the sample was stirred for an hour. COD concentration of the supernatant was tested and then removed the supernatant and another pH-adjusted MBR effluent was added again. After stirring the sample for another hour, COD concentration of the supernatant was tested. Redo the cycle steps. COD concentration of leachate of the test was 1052 mg/L. Results are shown in Table 2.40. Adsorptive property of activated carbon is greatly influenced by the activated carbon texture and production process. Activated carbon used was a cheap type named gas-adsorptive industrial activated carbon, of which the adsorptive property to organic pollutants in leachate was relatively poor. Specific adsorbing capacity in the initial one hour was 7.46 mg COD/g while

FIGURE 2.66 Leachate treatment flowsheet by Fe-C microelectrolysis.

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127

TABLE 2.40 Effect of Activated Carbon Adsorption in Fe-C Microelectrolysis on Treatment for Biologically Derived Leachate Index

1

2

3

COD removal (%)

3.6

1.1

0

Adsorbing capacity (mg COD/g)

3.8

1.2

0

FIGURE 2.67 COD removal by Fe-C microelectrolysis for MBR biologically derived leachate.

it was only 2.74 mg COD/g in the second one hour. In the third cycle, activated carbon reached adsorption saturation. Thus, because of the relatively poor adsorptive property, adsorption of activated carbon to COD was less than 5% in Fe-C microelectrolysis with a dosage of 1% (g/L). COD removal and reaction balance time are influenced by particle size of iron powders. The smaller particle size led to a shorter reaction time and a higher COD removal can be obtained with the same dosage and initial pH condition. It was inferred that reaction balance time of Fe-C reaction ranged from 8 to 10 hours when particle size of iron powders was 74 μm on basis of reaction balance time obtained in different particle sizes. COD concentration of leachate to be treated was 1052 mg/L while the one-time treatment capacity was 450 L. Dosage ratio of Fe/C was 1:1 and one-time dosage ratio of solid to liquid was 1%. The results are given in Fig. 2.67. It can be seen that at the iron powders’ particle size of 74 μm, the reaction balance time was 10 hours with the maximal COD removal of 58%.

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In order to test the operational stability of this device, run this device for 14 times continuously with a run cycle of 10 hours. Taking safety and manpower supply into consideration, running time was set only in daytime, which meant a daily treatment capacity of 450 L. It can be seen that the initial pH 2.0 6 0.2 increased to about 5.8 after running for 10 hours. pH change in the treatment process can help to signal whether the reaction reaches equilibrium. When pH reached about 6, it can be judged that reaction reached equilibrium. Therefore, change of pH shown in Fig. 2.68 can prove that 10 hours are the reaction balance time. According to Fig. 2.69, COD concentration ranges from 763 to 1204 mg/L in influent while it ranges from 278 to 532 mg/L in effluent. Since COD removal had a correlation with COD concentration in influent, when COD concentration in influent was relatively low, COD removal was high. The maximal COD removal by 74 μm Fe-C microelectrolysis was 63%, with a corresponding COD concentration in influent of 763 mg/L. However, it was found that COD concentration of MBR effluent ranged mostly from 900 to 1100 mg/L, according to monitoring of nearly 1 year. In this range, the corresponding COD removal was 58.6% in average. From Fig. 2.70, it can be found that TOC concentration ranges from 243 to

FIGURE 2.68 Effect of pH on treatment by Fe-C microelectrolysis for MBR biologically derived leachate.

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FIGURE 2.69 COD removal effects by Fe-C microelectrolysis on treatment for MBR biologically derived leachate.

FIGURE 2.70 TOC removal effects by Fe-C microelectrolysis on treatment for MBR biologically derived leachate.

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FIGURE 2.71 TN removal effects by Fe-C microelectrolysis on treatment for MBR biologically derived leachate.

341 mg/L in influent, and ranges from 88 to 168 mg/L in effluent and TOC removal ranges from 50.7 to 61.9%. TN removal effects of leachate by Fe-C microelectrolysis device are shown in Fig. 2.71. It can be found that TN concentration ranged from 85 to 238 mg/L in influent, and ranged from 42 to 129 mg/L in effluent and TN removal ranged from 41.9% to 51.7%. NH3-N concentration ranging from 5.7 to 23.9 mg/L in MBR effluent was relatively low while TN concentration was still high. It meant, in the biological pretreatment stage, nitration reaction was dominant while denitrification reaction was relatively suppressed because of shortage of carbon source, causing accumulation of much nitrate nitrogen and nitrite nitrogen in bio-treated leachate. Change of NO2-N concentrations before and after treatment is shown in Fig. 2.72. Before treatment, NO2-N concentration ranged from 57 to 187 mg/L and ranged from 7 to 47 mg/L after treatment, with a removal range from 74.8% to 81.0%. In acidic condition, NO2 2 , as electron acceptor, can react with H1 in solution, generating NH3-N. When using alkaline materials to precipitate ferric salt, NH3-N can be transformed to gas, making TN concentration decrease. However, reducing reaction of NO3-N occurs in the microelectrolysis, generating NO2 2 and further forming NH3-N to be removed. As shown in Fig. 2.73, the color of influent leachate is dark brown while the effluent has a color of light yellow with an obvious reduction of

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FIGURE 2.72 NO2-N removal effects by Fe-C microelectrolysis on treatment for MBR biologically derived leachate.

FIGURE 2.73 Influent and effluent color comparison by Fe-C microelectrolysis treatment for MBR biologically derived leachate.

chromaticity. After treatment by Fe-C microelectrolysis, sludge content was approximately 20% and clear effluent can reach 75%
80%. COD concentration in effluent was similar to that in the nanofiltration effluent, but concentrations of TN, nitrate nitrogen, and nitrite nitrogen of Fe-C microelectrolysis effluent were lower than that of nanofiltration effluent. Because of the membrane characteristics, nanofiltration membrane had a good retention effect on bivalent cations and a relatively poor retention effect on monovalent cations.

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Therefore, TN removal was lower for the macromolecular organic nitrogen 2 after nanofiltration, while the concentrations of TN, NO2 3 -N, and NO2 -N were relatively high. Nevertheless, Fe-C microelectrolysis process showed a great advantage in degrading TN, nitrate nitrogen, and nitrite nitrogen. The resultant iron slime and activated carbon can be washed, calcined, and reused as cellular packing materials to increase biodegradability of leachate.

2.9.2 74 µm Fe-C Microelectrolysis Under Venturi Vacuum Negative Pressure In Fig. 2.74, Φ1 is inlet diameter, and Φ2 is branch diameter. R is the ratio of inlet diameter to choke hole diameter. L is the length of Venturi pipe. L1 is the length of inlet section. When gas or liquid flows into the Venturi tube, dynamic pressure (velocity head) reaches the maximum while static pressure (resting pressure) reaches the minimum in the narrowest point of the pipe. Velocity of gas (liquid) increases because the flows cross sectional area in the pipe changes. The whole flow would undergo the narrowing process of pipe in the same time. Therefore, pressure would decrease in the same time, causing a pressure difference. This pressure difference can be used for measurement or to supply an external suction for fluid. For ideal fluid (gas or liquid, incompressible, no frictional property), this pressure difference can be obtained using the Bernoulli equation. The theory of the Venturi effect is that, when wind is blowing through the obstacles, the pressure around the port over the leeward of the obstacles becomes relatively low, consequently generating absorption effect and air flow. In other words, gas flow is narrowed to accelerate flow velocity of gas, forming a “vacuum” zone after the outlet of the Venturi pipe. Workpieces can be absorbed to some degree when getting close to this “vacuum” zone. Compressed air enters the Venturi pipe from the inlet and little of it is

L1

Φ2

Choke

R

Φ1

L FIGURE 2.74 Schematic diagram of Venturi effect.

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TABLE 2.41 Parameters of Venturi Pipe Used Size (L/s)

Ejection flow (L/s)

Output pressure of mixture (MPa)

Mixture flow (L/s)

R

Φ1 (mm)

Φ2 (mm)

L1 (mm)

L (mm)

0.5

0.528

0.05

1.056

0.23

50

32

95

350

discharged through the nozzle with a small cross sectional area. Then, the cross sectional area of pipe gradually decreases. Pressure of compressed air decreases and flow velocity increases, causing a vacuum zone in the inlet of branch and pushing ambient air into the Venturi pipe through the branch. The reaction device included automatic pH adjusting and dosing device, submersible leachate pump, Venturi pipe (detailed size showed in Table 2.41), tray, and U connecting reaction device made up of two PVC buckets. Bucket 1# and bucket 2# were connected by PVC control valve. Bio-treated leachate in ultrafiltration outlet storage pond was pumped into bucket 1# by submersible leachate pump. The automatic pH adjusting and dosing devices were turned on. Under stirring condition, pH of leachate to 2.0 6 0.2 was adjusted. Leachate in bucket 1# was then pumped out by leachate pump to form a continuous leachate flow through the Venturi pipe. In the choke, velocity of liquid increased because of the narrow caliber. In the meantime, a “vacuum negative pressure zone” was generated after the outlet of Venturi pipe and a negative pressure eddy current was generated at the branch port because of the pressure difference. Self-absorption state pushes superfine iron powders and activated carbon powders in dosing hopper at branch port, accompanied by air to enter Venturi pipe without extra power. In the “vacuum negative pressure zone,” solid, liquid, and gas are mixed. After flowing through the choke, the mixture, along with airflow, ejects with a high speed, causing fierce friction with air. A razor-thin contact layer would form in the surface of three phases, increasing the contact area and contact time. Then, the mixture entered bucket 2#. When liquid height in bucket2# exceeded that in bucket1#, control valve 1# was opened to connect bucket 1# and bucket 2#, forming a “U” connecting reaction device. Therefore, the reflux ratio was increased and circular reaction was maintained (velocity of ejecting . 0.5 L/s, distance . 6 m). This device is shown in Figs. 2.75 and 2.76. In order to prevent blockage in the Venturi pipe, granular activated carbon purchased was ground to powders that can fall through 100 mesh sieves. Superfine iron powders and activated carbon powders in dosing hopper can

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FIGURE 2.75 Venturi vacuum testing device for biologically derived leachate treatment.

FIGURE 2.76 Reactors for Venturi vacuum treatment for biologically derived leachate (1bucket 1#, 2-control valve 1#, 3-bucket 2#, 4-leachate pump, 5-control valve 2#, 6-Venturi pipe, 7-solid injection port).

be automatically dosed into the Venturi pipe by vacuum the eddy current effect. Dosing ratio was still 1% and one-time dosage ratio of iron powders to activated carbon powders was 1:1. It is shown in Fig. 2.77 that the Venturi effect enhances the reaction efficiency of leachate treatment by Fe-C microelectrolysis device. Compared to static experiment of magnetization, the Venturi effect can shorten the reaction balance time greatly. After reaction for 2 hours, COD removal can reach 56% and the reaction rate was 4
5 times faster. It was due to the fact that pressure difference can increase velocity of ejection, strengthen collision among particles, and therefore shorten the reaction balance time, when the mixture of leachate, air, superfine iron powders and activated carbon flowed through the choke of Venturi pipe. When reaction time was shortened to 2 hours, daily treating capacity of leachate can be increased to 1800 L/d.

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FIGURE 2.77 Enhancement of reaction efficiency by Venturi vacuum negative pressure for treatment of biologically derived leachate.

COD and TOC removals of biologically derived leachate treatment by Venturi effect enhanced Fe-C microelectrolysis device are shown in Fig. 2.78. It can be found that COD concentration ranged from 890 to 1023 mg/L in influent, and ranged from 359 to 431 mg/L in effluent and COD removal ranged from 55.8% to 60.6%. Besides, TOC concentration ranged from 315 to 397 mg/L in influent, and ranged from 123 to 166 mg/L in effluent and TOC removal ranged from 57.8 to 61.7%. TN and NO2-N removals from the biologically derived leachate by the Venturi effect enhanced Fe-C microelectrolysis device are shown in Fig. 2.79. It can be found that TN concentration ranged from 135 to 198 mg/L in influent, and ranged from 69 to 101 mg/L in effluent, and TN removal ranged from 46.0% to 54.8%. Besides, NO2 2 -N concentration ranged from 89 to 3169 mg/L in influent, and ranged from 13 to 35 mg/L in effluent and NO2 2 -N removal ranged from 74.7% to 89.3%.

2.10 ADVANCED LEACHATE TREATMENT PROCESS USING HYDRATION REACTION FOR BIOLOGICALLY DERIVED LEACHATE Leachate contains a large amount of organic matters with a wide range of molecular sizes, and its oxidation/reduction potential will decrease during storage. Concerning the dissolved matter (DM), the ratios DM/TS in fresh

FIGURE 2.78 COD and TOC removal for treatment of biologically derived leachate using Venturi vacuum negative pressure and Fe-C microelectrolysis device.

FIGURE 2.79 TN and NO2-N removals from biologically derived leachate using Venturi vacuum negative pressure and Fe-C microelectrolysis device.

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leachate, effluents from aging tank, anaerobic, anoxic, and aerobic lagoons were about 0.60, 0.78, 0.83, 0.92, and 0.89, respectively. The percentages of DM in terms of TOC in leachate increased in the biological treatment processes, indicating that the matters removed by the biological treatment should be predominantly in the insoluble matters. Furthermore, it can also be assumed that the biochemical treatment techniques are effective for the removal of the matter with a molecular weight over 1000 Da. In comparison, coagulation, activated carbon absorption, biological treatment processes, and RO technology can remove macro-molecular weight with .3000 Da, 500 Da-3000 Da MW, micromolecular, and dissolved matter, respectively. Therefore, it demonstrates that a RO unit should be used in leachate treatment process in landfills after the biological treatment according to the result of effluents from aerobic tank. Similarly, the ratios of FS/TS in the corresponding leachate or effluents were 0.37, 0.51, 0.67, 0.64, and 0.59, respectively, indicating that FS predominated in the leachate or effluents issued from the treatment stages. Comparing to the effluents from the treatment stages, fresh leachate contained more organic matters and most of the removals were volatile organic matter, occurring during the biochemical processes. With the stricter leachate discharge standard, some physical-chemical treatment processes, such as chemical oxidation, air stripping, ion exchange, membrane process, adsorption on active carbon, and coagulation-flocculation, are coupled with those biological treatment processes, while a high operation cost, the stricter operation condition, the difficult handling of the by-products, or the low removal efficiency prohibit the wide application of these techniques in the practical leachate treatment projects. Hydration reaction has been commonly used in the construction and environmental fields (e.g., for stabilization and solidification) to form a solid mass, and some organic matters, ions (anion and cations) are found to be incorporated into the cement matrix during the hydration process. It has also been observed that the hydration products, e.g., calcium silicate hydrates (CS-H), Ca(OH)2(C-H), AFt (Al2O3-Fe2O3-tri), and AFm (Al2O3-Fe2O3-mono) could gradually bond with a wide range of contaminants, such as heavy metals, inorganic matters, and some organic matters, which are the typical contaminants involved in wastewaters. Up to now, most of the works on hydrating cements focused on the influence of additives, such as metals, anions, polymers, on the cement hydration development processes and the properties of the cementitious products, and the water/cement ratio and the reaction time were usually around 0.5 and 28 days, respectively, to obtain the hardened cement with a high flexural strength and tensile strength.

2.10.1 Hydration Reaction for Biologically Derived Leachate The hydrating cement applied involved several main mineral compositions, i.e., 50% of C3S (i.e., 3CaO  SiO2), 25% of C2S (i.e., 2CaO  SiO2), 12% of

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139

C3A (i.e., 3CaOAl2O3), 8% of C4AF (i.e., 4CaO  Al2O3  Fe2O3), and 3.5% of CSH2. Leachate was collected from the collection tank after being treated by aged refuse bio-filter in Shanghai Refuse Landfill. COD in leachate was 485 mg/L, while BOD5 was below the determination limit. pH and conductivity were around 8.14 and 1029 μs/cm, respectively. TN was 224 mg/L, 2 and the corresponding NH3-N was 91 mg/L. Cl2, SO22 4 , and F concentration were 2303, 905, and 9.9 mg/L, respectively, given that the high percentages of food waste in refuse in China (about 50%
70% in fresh refuse). Ca21 in leachate reached to 90.5 mg/L, possibly due to the mix of the construction and demolition waste in landfill. Fe, Al, and Zn concentration were 3.2, 6.6, and 2.0 mg/L, respectively. Some 50 g hydrating cement and 500 mL leachate were introduced into a quartz glass reactor of 1.0 L equipped with a magnetic stirrer at 100 rpm. The supernatant was centralized for analysis after the end of the defined interval time of 24 hours, and the residual cement slurry stayed in the reactor for the next round, which was kept running until COD in the leachate effluent was above 300 mg/L. Parts of cement slurry samples were taken out from the reactor after 1, 4, 8, and 11 days and prepared for a slice after dried in the glovebox within a dry N2 atmosphere for the analysis.

2.10.2 Treatment Process for Biologically Derived Leachate pH increased dramatically from 8.14 to 12.60 as the hydration reaction proceeded, and maintained at a high level of more than 11.95 after 3rd day (Fig. 2.80A), ascribed that amount of Ca(OH)2 was generated in the initial period, and then CaO  SiO2  nH2O (C-S-H) consumed Ca(OH)2 greatly, which resulted in the decrease of pH from 11.95 to below 9.00 in the 5th day. Finally, it kept at a steady stage of pH 8.10
8.50. Conductivity in leachate increased from 1029 to about 1950 μs/cm in the first 2 days, and then decreased to around 1000 μs/cm after 5 days. COD and TOC in the resultant leachate are shown in Fig. 2.80B. COD was around 150 mg/L in the first 2 days, with a removal rate of 70%, and then presented at a range of 150
200 mg/L in the next 6 days. The removal rate of TOC was about 55% in the first 2 days, and then decreased to about 14% dramatically in the next 4 days. Finally, it reached almost zero after the reaction for 10 days. NH3-N in leachate was below 10 mg/L, with a removal rate of above 90% in the entire operation period of 11 days (Fig. 2.80B). Around 40% TN was removed in leachate in the first 5 days, and then only 22% of TN removed. Variation of cations and anions in the resultant leachate is shown in Table 2.42. Ca21 concentration increased from 90 to 635 mg/L in the 1st day, and then decreased to about 5 mg/L, with a removal rate of 94%, which resulted from the balance between the generation of Ca(OH)2 and the consumption of Ca21 by C-S-H. Al31 was also found to increase from 6.6 to 12.7 mg/L initially, and then C-S-H generated during the hydrating process

FIGURE 2.80 Pollutants removals verse reaction time using hydration reaction for biologically derived leachate ((A)-pH and conductivity; (B)-Removal rate of TOC, COD, NH3-N and TN)).

TABLE 2.42 Ion Concentrations in the Resultant Leachate Collected at the Defined Interval Time (mg/L) During the Hydration Reaction Process for Biologically Derived Leachate Day Metal

1

2

3

4

5

6

7

8

9

10

11

Ca

635

182

115

95.12

54.25

12.34

5.87

6.91

4.92

6.23

5.32

Na

2071

2054

2107

2176

2133

2091

2117

2201

2154

2169

2106

Mg

0.3

2.9

1.2

0.8

0.6

1.5

0.9

0.9

1.0

0.8

0.7

Al

12.7

13.9

7.5

7.1

5.6

1.8

1.5

2.3

1.9

1.6

1.7

Fe

1.1

1.4

0.2

0.8

0.6

0.4

0.6

0.5

0.4

0.5

0.3

Ni

0.13

0.14

0.13

0.16

0.15

0.15

0.14

0.12

0.13

0.13

0.15

Cr

0.08

0.11

0.12

0.11

0.09

0.07

0.10

0.08

0.11

0.09

0.12

Cu

0.04

0.03

0.05

0.02

0.03

0.03

0.05

0.03

0.02

0.04

0.05

Zn


































As


































Pb


































Cl2

1591

1628

1769

1758

1837

2012

2046

2171

2219

2361

2489

SO22 4

192

185

271

264

319

381

506

576

661

645

708

F2

9.0

8.7

7.5

7.2

5.5

6.8

7.2

7.5

8.0

7.4

7.7

a

“
” means “below the detection limit.”

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FIGURE 2.81 Variation of molecular weight distribution (%) using hydration reaction for biologically derived leachate.

was contributed to the reduction of Al31 to about 2 mg/L. Cl2, SO22 4 , and F2 contents decreased significantly in the first 5 days, and then increased as the reaction time extended, compared to the corresponding concentrations of 2303, 905, and 9.9 mg/L in raw leachate. The F2 content decreased slightly in the first 5 days through the generation of CaF2, since CaF2 is a slightly soluble compound.

2.10.3 Molecular Weight (MW) Distribution of Hydration Reaction for Biologically Derived Leachate Molecular weight (MW) distribution of hydration reaction for biologically derived leachate is summarized in Fig. 2.81. The fraction with Mn (numberaverage molecular weight) range of ,100K (100,000) predominated in leachate, with the percentage over 90%. Generally, the percentages of the fractions with Mn range of 1K (1000)
10K (10,000), 10K
100K, and 100K
1000K increased as the reaction time extended, while that with Mn range of ,1K decreased significantly, from 32.9% to 25.4% in the 1st day, and then to 3.2% after 3 days, since the generation of the gel, ettriengitel, and Ca(OH)2 contributed to the contaminants’ removal. Therefore, hydration reaction was effective in the removal of the substances with Mn range , 1K in leachate. It was also observed that MW distribution of leachate effluents varied greatly in the first 3 days, and then kept in a stable level in the latter period, meaning that the hydration reaction process might be more active in the first 3 days (in the acceleratory period), and then reached the deceleratory period as the reaction proceeded.

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FIGURE 2.82 X-ray diffraction of cement slurry after reaction using the hydration reaction for biologically derived leachate; (A) X-ray diffraction of (PC) Ref pastes of Portland cement at 0 day, (B) X-ray diffraction of (PC) Ref pastes of Portland cement at 1 day, (C) X-ray diffraction of (PC) Ref pastes of Portland cement at 4 day, (D) X-ray diffraction of (PC) Ref pastes of Portland cement at 8 day, and (E) X-ray diffraction of (PC) Ref pastes of Portland cement at 11 day.

2.10.4 Characteristics of Hydration Products Using XRD for Biologically Derived Leachate Hydrating cement contains calcite phase, dicalcium silicate (C2S) phase, tricalcium aluminate (C3A) phase, and tricalcium silicate (C3S) phase, as shown in Fig.2.82A. Calcite, dicalcium silicate, and tricalcium silicate were presented in the cement slurry, while C3A disappeared after the reaction for 1 day (Fig.2.82A), meaning that C3A was more active than other clinker minerals in hydrating cement for pollutant removal in leachate. Calcite, dicalcium silicate, and mogonite were found in the cement slurry, and C3S , disappeared after the reaction for 4 days (Fig.2.82C). However, mogonite disappeared in the slurry after 8 days reaction (Fig.2.82D), and quartz , appeared in the slurry after the reaction for 11 days (Fig.2.82E), suggesting that mogonite would convert into quartz in the latter period. Therefore, C2S, C3A, and C3S were the 3 main important hydrating products for pollutant control in the leachate treatment process. C3A and C3S contributed to the reduction of COD, TOC, and NH3-N in leachate in the initial period greatly. C2S would be helpful for the COD removal with a long-term period, and COD in leachate effluents was found to increase greatly after C2S exhausted.

2.10.5 SEM Analysis of Cement Slurry Using the Hydration Reaction for Biologically Derived Leachate The property of the cement slurry was tested in the hydrating cementleachate system (HC-L) and the hydrating cement-pure leachate system

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FIGURE 2.83 SEM images of hydration products using the hydration reaction for biologically derived leachate ((A) Hydration cement-leachate system; (B) Hydration cement-pure leachate system).

(HC-P), using SEM analysis, to identify the potential influence from the pollutants in leachate on the hydration development process (as shown in Fig. 2.83). It was shown that the hydration products in the HC-L system presented sheet and acicular covering structure after the 1st day reaction, and these grain residues were held together by a discontinues but interpenetrating mass of C-S-H forms. Homogeneous micropore structure was also found on the acicular covering surface of C-S-H. For the HC-P system, the rods and acicular covering structure was observed on the appearance and aggregated together. Therefore, more hydration products with sheet structure was presented in the HC-L system, compared to that in HC-P system, and the acicular structure aggregate showed less intensively and lower compact, meaning that the leachate influenced the cement structure greatly. The discontinuities in the cement microstructure were also observed. The aggregation with micro pore-structure might contribute to the removal of pollutants through the surface adsorption, and the generation of some clusters with multiporosities would be helpful for the contaminants adsorption from the leachate.

2.10.6 Characteristics of Hydration Products Using FTIR for Biologically Derived Leachate IR peaks and spectroscopic assignments of the cement slurry are shown in Fig. 2.84. Visible peaks at 884, 1786, and 2510 cm21 in HC-L system indicated that organic pollutants in leachate were adsorbed and complexed in the hydrating cement. The adsorption peak at 3645 and 3412 cm21 might be due to the presence of O-H and some free leachate in the hydrating cement. The IR spectrum of HC-P showed a signal at 2510 cm21 with the bending of CC or CN, as well as the band at 1786 cm21 due to the bending of C-H adjacent to C 5 C. Moreover, the signals at 3440 cm21 might correspond to

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FIGURE 2.84 Infrared spectrum of hydrating cement (1) and hydration products (2,3) using the hydration reaction for biologically derived leachate (1-hydrating cement, 2- hydrating cement 1 leachate system, 3- hydrating cement 1 pure leachate).

the bending of the symmetrical and asymmetrical stretching vibration band of O-H in the leachate. The band of 925 and 1000 cm21 adsorption spectral showed the asymmetrical stretching vibration band of Si-O in silicate tetrahedrons. The signals at 1092 cm21 were due to the S-O stretching vibration 21 in SO22 was the C-O 4 . The strong broad peak at range of 1420
1480 cm stretching vibration band in carbonate, while that in the HC-L system showed a sharp peak, and the reaction between CO2 and C-H in hydrating products might be the predominate contributor.

2.10.7 Pollutant Removal Capacity of Hydrating-Cement and Cost Estimation for Biologically Derived Leachate Treatment COD and TOC in leachate were found to be removed simultaneously in the HC-L system in the first 3 days. After that, COD kept decreasing greatly, while the corresponding TOC almost stopped decreasing. The possible reason was that amounts of hydration products, i.e., C-H, C-S-H, Aft, and AFm, generated in series during the hydrating process, and Aft, AFm, and C-S-H could adsorb, complex, and oxidize some reductive substances, which resulted in the reduction of COD, with a removal rate of over 50% for a

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period of 6 days. TOC removal might be relied on the surface adsorption, which was exhausted after the reaction for 3 days. It could be predicted that most of the TOC contributors in leachate might be in the fraction with the Mn range of 1K
100 K. The removal of some inorganic matters, which influenced the hydration process greatly, was another reason for the reduction of COD in the latter period of the hydration reaction process. It had been proved that some compositions in the hydrating cement, i.e., C3A, C4AF, and C3S, could be complexed with Cl2 to form the nonsoluble compound in leachate as follows: CaCl2 1 C3 A-C3 A  CaCl2  10H2 O

ð2:44Þ

SO22 4 could be also reduced by the slightly soluble sediment CaSO4 generated as follows. Na2 SO4 1 CaðOHÞ2 1 2H2 O-CaSO4  2H2 O 1 2NaOH

ð2:45Þ

CaSO4  2H2 O 1 C3 A 1 12H2 O-3CaO  Al2 O3  CaSO4  12H2 O

ð2:46Þ

With respect to NH3-N removal in leachate, stripping, adsorption, and sedimentation might be the main contributors in the hydration process. Stripping might reduce NH3-N concentration in the first 3 days, ascribed that NH3-N was readily to be stripped by the agitation under pH value of above 11. As pH value decreased to below 10 in the 4th days and then to below 9 in the 5th day, the adsorption or sedimentation might contribute more for NH3-N removal. The maximum removal capacity of the hydrating cement is another critical point to test for the potential practical leachate treatment application, and the accumulated COD, TOC, and NH3 removal were around 23.7 mg (COD)/g, 3.1 mg (TOC)/g, and 9.2 mg (NH3)/g (hydrating cement) in the test period, respectively, meaning that the hydrating cement could be the promising process, to cope with biological treatment for the high strength wastewater disposal. The biological unit responds to the biodegradable substances removal in terms of TOC and BOD, while hydrating reaction contributes to the removal of the residual refractory matters. The operation cost is the critical factor for the choice of advanced leachate treatment process. A preliminary cost comparison is carried out between hydration reaction and Fenton, where Fenton is the common choice for advanced leachate treatment. Only the raw material costs are considered for the comparison here, since both of them has a similar and simple operation process and energy costs, such as dosage pumps and stirrers, are supposed to be the same during the reaction process. The hydrating cement dosage was around 14 kg/t leachate based on the results (the ratio of hydrating cement: leachate was around 1:10, and the hydrating cement was for 7 days), where the material cost was around 0.71 USD/t. For Fenton, the dosages of H2O2 and Fe21 were around 0.4% and 1000 mg/L under the same test conditions, respectively, and the corresponding material cost was around 5.7 USD/t.

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This result showed that hydrating cement was competitive and could be an option for the advanced leachate treatment.

2.11 MEMBRANE TREATMENT PROCESSES FOR LEACHATE Membrane separation processes, such as UF, NF, and RO, have been widely used in leachate treatment. The detailed information of those membrane processes are shown in Table 2.43. Nearly all the membranes can be commercially available.

2.11.1 Ultrafiltration (UF) UF, a kind of membrane separation technique, is able to remove the particle materials in solutions. This separation process is usually used in industry and research areas for purifying and concentrating macromolecular (103
106 Da) solutions, especially protein solutions. Like microfiltration, ultrafiltration is based on size exclusion or particle capture. Ultrafiltration membranes are defined by the molecular weight cut-off (MWCO) of the membrane used, and cross-flow or dead-end mode is applied in different areas. The ultrafiltration membranes are only accessible for small molecules, such as leachate molecules, inorganic salts, and micromolecular organics, rather than some macromolecules like SS, colloid, protein, and bacteria. UF is always used as a way to pretreat leachate with lower organic concentration before nanofiltration (NF). TABLE 2.43 Characteristics of Membrane Separation Technology for Leachate Treatment UF

NF

RO

Membrane flux (L/ h  m2  bar)

10
1000

1.5
30

0.05
1.5

Pressure (bar)

0.1
5

3
20

5
120

Membrane pore size (nm)

2
100

0.5
2

, 0.5

Separation mechanism

Pressure differential / Classification

Pressure differential / Solution-diffusion

Pressure differential / Solution-diffusion

Application

Removal of Macromolecules, bacteria and viruses

Removal of ions and small organisms

Ultrapure leachate, purification

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2.11.2 Nanofiltration (NF) & Colloid Separation Membrane (CSM) NF is a pressure-driven membrane separation technique, and nanofiltration membranes have pore sizes of 1
10 nm, smaller than that used in microfiltration and ultrafiltration, but just larger than that in reverse osmosis. Membranes used are predominantly created from polymer thin films. Compared with RO membranes, NF membranes perform a lower operation pressure demand and energy consumption (10 kWh/m3), and a higher membrane flux under the condition with hyper-saline concentration and low pressure. Therefore, it has been widely applied to leachate treatment, food industry, pharmacy, and many other fields. The aperture of CSM is between the traditional ultrafiltration and the nanofiltration. CSM can effectively intercept the macromolecular colloidal organic matter, such as humus in the concentrated fluid. It is widely used in material separation, clarification, purification, and enrichment of food, medicine, environmental protection, and other fields. CSM has the process advantages of harmless, modularization, and reduction. Organic pollutants can be incinerated, recycled, and cured after being concentrated. The membrane components are highly integrated, modular, and easy to control, and the total water yield is over 97%. Now this technology is applied in engineering projects. The process flow chart of a 140 t/d concentrate leachate treatment project in Jiangsu province is shown in Figs. 2.85
2.87.

2.11.3 Membrane Bioreactor (MBR) MBR is a novel and efficient biological treatment technology that includes biological treatment and membrane separation. It has highly efficient solidliquid separation and can overcome the poor quality of effluent and sludge bulking, compared to the conventional methods. Three typical MBRs have been widely used, based on the different functions of membrane: aerated membrane bioreactor (AMBR), extractive membrane bioreactors (EMBR), Suspended solids, bacteria Microfiltration > 0.1µm Macro-molecular organics, protein, polypeptide, etc.

Ultrafiltration > 0.1-0.01µm or 1000-300,000MW Nanofiltration > Molecular Weight Cut-Off 150-300MW

Micro-molecular organics, dye, heavy metal

CSM

Inorganic salt residual organics

Reverse osmosis > Molecular Weight Cut-Off < 150MW water

FIGURE 2.85 The aperture of Colloid Separation Membrane (CSM).

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149

FIGURE 2.86 Treatment effect of combined process (CSM and NF).

Q = 100 m3/d COD = 1000 mg/L

Q = 85 m3/d COD = 80 mg/L

Q = 98.8 m3/d COD = 80 mg/L

Nanofiltration

Super clear filtrate

Discharge with related standard

3/d

Q = 15 m COD = 6213 mg/L Q = 1 m3/d COD = 80000 mg/L

GSM+NF

Q = 11 m3/d COD = 120 mg/L

Q = 3 m3/d COD = 3960 mg/L High concentration wastewater

Medium concentration wastewater

800-1300 KCal/kg

Incineration/resource/cure

Low concentration wastewater

Q = 2.8 m3/d COD = 100 mg/L Coagulation + oxidation

Q = 13.8 m3/d COD = 80 mg/L Activated carbon

Sludge and biochemical systems Sludge mixed treatment

FIGURE 2.87 Concentrated liquid treatment project by combined process (CSM and NF) in Jiangsu province.

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FIGURE 2.88 Schemes of submerged MBR and side-stream MBR.

and solid-liquid separation membrane bioreactor (SLSMBR). According to the location of membrane module, submerged membrane bioreactors, side-stream membrane bioreactors, and compound membrane bioreactors are typically used in the practical projects, as shown in Fig. 2.88.

2.11.3.1 Side-Stream Membrane Bioreactor Side-stream membrane bioreactor refers to membrane components and bioreactors set separately and connected by pumps and pipelines, thus leading to little interference and easy adjustment. And the membrane modules outside of the bioreactor are easy for being cleaned or replaced, but its power consumption is high for high pressure, which can delay membrane pollution using high-speed cross-flow. Energy consumption of per ton effluent is 2 B 10 kWh, 10 B 20 times of that of traditional activated sludge process; therefore, research on submerged membrane bioreactors with low energy consumption have gradually been highlighted. 2.11.3.2 Submerged Membrane Bioreactor Submerged membrane bioreactors (SMBR) refers to membrane modules positioned directly inside the bioreactor, which relies on gravity or vacuum generated by suction pump or vacuum pump. It reduces the floor space and power consumption cost, about 9/10 lower than that of the side-stream membrane bioreactor. Cost can be completely saved using gravity. However, since the membrane modules are immersed in the bioreactor, they can be contaminated rapidly. Besides, they are difficult to clean, needing to be removed from the reactor. 2.11.3.3 Compound Membrane Bioreactor Compound membrane bioreactor refers to membrane modules placed inside the bioreactor and the effluent is discharged by gravity or negative pressure, while the bioreactor is filled with packing, which can improve the impact

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resistance of the processing system and reduce suspended activated sludge concentration for less membrane fouling and higher membrane flux.

2.11.3.4 Technical Features of MBR Good and stable effluent quality, high impact resistance. Due to efficient membrane separation, the suspended solids and turbidity of the effluent are close to 0, and bacteria and viruses are substantially removed. Even for manure wastewater without ideal treatment effect in conventional buildings, its effluent quality is also better than that is required in Reuse of Recycling Water for Urban Water Quality Standard for Urban Miscellaneous Water Consumption (GB / T 18920
2002). Membrane separation also allows microorganisms completely to remain in the bioreactor, which not only improves the removal efficiency of pollutants, but also has strong adaptability to the influent load (quality and quantity). High water production rate. Conventional treatment technologies are mostly composed of biochemical, physical, and chemical methods, like coagulation and sedimentation, sand filtration, or activated carbon filtration, which can need frequent water backwash, thus consuming a certain amount of clear effluent, the system low water production rate. The membrane bioreactor combines the membrane separation components and biochemical treatment effectively, asking for no coagulation or filtration. So, it saves both labor and effluent backwash operation, leading to high water production rate of close to 100%. Low excess sludge production, no chemical sludge, no secondary pollution. Membrane bioreactor can run in high volume load and low sludge load; moreover, it has low excess sludge production (zero emissions theoretically), leading to low sludge treatment costs and possible pollution diffusion in the process of sludge disposal. In addition, compared with the conventional leachate treatment process, there is no chemical sludge disposal problem caused by large amounts of coagulant for membrane bioreactor, and the secondary pollution can be minimized. Small floor area and easy for construction. Membrane bioreactor can maintain high concentration of microbial biomass for high volumetric load, thus greatly saving the floor area. The whole process with compact structure is simple, suitable for any occasion; can be made as ground, semiunderground, and underground membrane bioreactor. Easy operation and management for automatic control. Membranebioreactor completely separates HRT from SRT, making operation and control more flexible and stable. It is a new easy-equipped technology for leachate treatment, with full automatic operation, thus making operation and management more convenient. Easy to combine with traditional treatment process. Membrane bioreactor and conventional biochemical treatment process can be used together, e.g.,

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TABLE 2.44 Typical Design Parameters for MBR Cooling Device Used in Summer Design parameters

Quantity

MBR daily influent Qd

Qd5800 m3 /d

MBR hourly influent Qh

Qh 5 33.3 m3 /h

Daily total COD load X0 COD

X0 COD 5 19,360 kg COD/d

Daily total BOD load X0 BOD

X0 BOD 5 12 kg BOD/m3 3 800 m3 /d 5 9600 kg BOD/d

Daily total NH3-N load XNH3-N

XNH3-N 5 2000 kg NH3-N/d

Designed denitrification ratio RDi

RDi 5 99.3%

Ambient temperature in summer Tev

Tev 5 40 C

Leachate influent temperature in summer Tin

Tin 5 38 C

Controlled temperature for biochemical process Te

Te 5 35 C

and if the conditions permit, mounting membrane module in the sedimentation tank or other reasonable transformation will improve the effluent quality greatly. However, there are some deficiencies of the membrane bioreactor. The polluted membrane and decreased effluent yield make operation management inconvenient during operation. The manufacturing cost of the membrane and its energy consumption is high. In summer, a lot of heat will be released and a cooling device should be installed as given Tables 2.44 and 2.45.

2.11.4 Reverse Osmosis (RO) RO is a leachate purification technology that uses a semipermeable membrane to remove ions, molecules and larger particles in leachate. An applied pressure is used to overcome osmotic pressure and a colligative property, driven by chemical potential differences of the solvents. Two-stage RO treatment for leachate may be used. The operation pressure should be controlled between 36 and 60 bar, and infiltration capacity should remain at about 15 L/m2/h, as shown in Table 2.46.

2.11.5 Disc Tube Reverse Osmosis (DTRO) Technology The predominant devices for reverse osmosis is DTRO, in which the high pressure reverse osmosis membrane is the core element to separate fresh

TABLE 2.45 Heat Balance Calculation of MBR Cooling Device in Summer Heat calculation

Quantity

(1) Influent heat Qhyd

Qhyd 5 (Qh/3.6) 3 4.2 3 (Tin
Te) 5 (33.3/3.6) 3 4.2 3 (38
35) 5 117 kW

(2) Heat effect by pumps of biochemical section Qpump (Assuming that the ratio of pump shaft power converted to heat is 80%)

Qpump 5 (Pjet pump 1 Pultrafiltration circulation pump 1 Pultrafiltration influent pump 1 Pnitrate reflux pump 1 Pcooling sludge pump) 3 80% 5 360 kW

(3) Heat release by bioreactor Qbio

Qbio 5 Qnitration 1 QCOD degradation 1 Qdenitrification 5 (XNH3-N 3 qnitration 1 (X0 COD-X0 BOD) 3 qCOD degradation 1 XNO3-N 3 qdenitrufication)/24 5 3612 kW

(4) Heat by air blast Qair

Qair 5 (hin
hout) 3 Vair 3 ρair 5 (0.056
0.035) 3 18,360 3 1.09 5 420 kW

(5) Demanded heat for cooling Qsum

Qsum 5 Qhyd 1 Qpump 1 Qbio 1 Qair 5 117 1 360 1 3612 1 420 5 4509 kW

(6) Cooling water for circulation Qrec.

Qrec. 5 Qsum/ΔT/1.16 5 4509/3.5./1.16 5 1111 m3/h (Designed to be 1140 m3/h)

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TABLE 2.46 Treatment Efficiency of Two-stage RO System for Leachate Parameters

Influent leachate

Effluent of the first RO

Effluent of the second RO

pH

7.7

6.8

6.6

Conductivity, μS/cm

17,250

382

20

99.9

COD, mg/L

1797

15

,15

.99.2

NH3-N, mg/L

366

9.8

0.66

99.9

2830

48.4

1.9

99.9

Na , mg/L

4180

55.9

2.5

99.9

Heavy metal ions, mg/L

0.25

,0.005

,0.005

.98

2

Cl , mg/L 1

Removal rates (%)

leachate and impurity, made of polymer materials and aromatic polyamide, which is usually used as the disc diaphragm material for excellent chemical properties. Leachate gets initial pressure by feed pump and then gets pressure by high-pressure pump after the security filter. Circulating pump provides a larger flow in order to meet the required DTRO membrane surface flow velocity. Liquid flow in the disc flow channel in S shape, small molecule particles, and dissolved ions are trapped in the concentrated liquor side, and filtrates are collected. DTRO membrane module structure is different from the conventional spiral-wound module. For instance, concentrate port, disc tube membrane module has unique flow channel design of open flow channel. The liquid enters the pressure vessel through the entrance, flowing to the other side of the module through the channel between the guide disc and shell. At the flange, the liquid, through eight channels, enters into the guide disc; the treated liquid quickly flows through the filter membrane at the shortest distance, and then reverses to another membrane surface, flowing into next guide disc through the central slot of the guide disc, thus formed the double S shape flow path, which turns back between the edge and the center of the guide disc. Finally, the concentrate flow out from the flange in the feed side. The distance is 3 mm of each two guide disc, as the regular bumps distribute on the plate surface. The special hydrodynamic design makes the pressure fluid run into the bumps on the membrane surface to form turbulence, increasing penetration and self-cleaning function, effectively avoiding the phenomenon of membrane plugging and concentration polarization, and successfully extending the service life of the diaphragm. When cleaning, it also helps to clean the fouling easily on the membrane, and ensures disc tube membrane affording the poor influent (Figs. 2.89 and 2.90).

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155

FIGURE 2.89 DTRO flow diagram.

FIGURE 2.90 Diagram of disc/diaphragm position.

Penetrant is separated through the membrane vertically from the outer to inter, and then flows out through the central rod of membrane module. The separation of concentrate and penetrant depends on the seal between disc and diaphragm. DTRO membrane for leachate treatment is a physical process and independent of the leachate biodegradability. Due to the high interception on pollutants of DTRO for leachate with various aging time, the effluent can meet the discharge standards steadily, not affected by the biodegradability and carbon nitrogen ratio. The DTRO module contains 3 mm-wide open flow channel and unique guide disc with bumps, easily forming turbulence in the module to reduce membrane pollution, surface scaling, and concentration polarization. The DTRO membrane module can effectively avoid the fouling, decreasing membrane pollution, then prolonging the service life of RO membrane.

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Pollution Control Technology for Leachate From Municipal Solid Waste

Special structure and hydrodynamic design make modules easy to clean, obtain good flux recovery, so as to extend the life of the membrane bag. DTRO membrane modules are of standardized design, component is easy to remove and maintain, check any piece of filtration membrane bags and other parts. When there not enough component parts, component allows reduction of some membrane bags and guide disc, which doesn’t affect the use of the DTRO membrane module compared to other type membrane modules. Any parts in DTRO modules are allowed to be replaced independently. Filtration part is made of multiple membrane bags and guide disc. The membrane bag can be replaced one by one, while membrane bag of good filtering performance still can continue in use. The replacement cost of filtration membrane is much reduced, which is ahead of other types of membrane, such as spiral-wound membrane and hollow-fiber membrane. When spiralwound membrane is of patch, localized seepage and other quality problems, or need to replace with a new membrane, one has to replace the whole module. The DTRO membrane system is of flexible operation, which can run continuously or intermittently. The system can adapt to the different requirements of leachate quality and quantity by adjusting the series-parallel mode. The construction of a DTRO membrane system mainly contains equipment building, accessorial workshop, and pool construction of small construction scale and high construction speed. The installation and debugging can be achieved in 2 weeks after receiving the equipment. The DTRO membrane system with perfect monitoring and control functions is fully automatic. PLC can automatically adjust according to the sensor parameters, timely send out an alarm signal, forming protection to the system. Operators just need to consult the error code according to the operation manual troubleshooting, with no high requirement of experience on personnel operating. Also, the DTRO membrane system is of integrity, the accessory structures and facilities are also some small structures, covering a small area.

2.11.6 Membrane Combination Technology 2.11.6.1 MBR 1 Two-Membrane (NF/RO) Process MBR 1 two-membrane (NF/RO) technology is a strong integrated process. With a designed treatment capacity of 80 m3 leachate per day, a leachate treatment plant has adopted the MBR 1 two-membrane (NF/RO) process, as shown in Fig. 2.91, which has also been applied widely in China. Table 2.47 presents the fluctuation range of the influent and effluent in different stages. The concentrated liquid resulted from an RO system, around 20
30% of the influent quantity, has not any cost-effective solution so far. Evaporation may be a choice, but is costly. Incineration with refuse at an incineration plant

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157

Regulating tank Lift pump Bag filter

Denitrification tank

MBR

Sludge and nitrification liquor recycling

Nitration tank

Tubular UF

NF system

RO system

Cleaning system

Meet the discharge standard

Concentrated solution tank and sludge tank

FIGURE 2.91 Process flow diagram of MBR 1 NF/RO system for leachate treatment.

has been developed quickly and practiced, especially the leachate generated in incineration plants.

2.11.6.2 Two-Stage Disk-Tube Reverse Osmosis (DTRO) System In two-stage DTRO system, leachate is first introduced into the regulating tank, and pH of effluent from the storage tank is adjusted by acid in order to avoid scaling caused by carbonates or other inorganic salts. Then the leachate is filtrated in grit filter and cartridge filter to decrease SS concentrations. Then, the leachate is fed to the first stage of the DTRO system for reverse osmosis filtration. When the effluents flow to the second stage of the DTRO system, the concentrated liquid is discharged to the storage tank for the further leachate recirculation. The permeate effluents from the second stage is discharged to the degassing tower and CO2 or other gases dissolved in leachate are stripped until pH 6
9, and recycled for other uses or discharged into

TABLE 2.47 Quality of Influent and Effluent in Different Stages for MBR 1 Two-membrane (NF/RO) Process for Leachate Treatment Stage

COD (mg/L)

BOD5 (mg/L)

SS (mg/L)

NH3-N (mg/L)

TN (mg/L)

pH

Influent

6300
8100

1200
2000

600
1000

450
800

500
1000

6.5
7.6

MBR

320
900

60
85

2
10

3
15

7
28

7.2
8.3

NF

25
92

10
18

0
3

2
10

4
20

6.5
7.7

RO

10
38

3
10

0
2

0
7

0
10

6.1
7.2

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159

TABLE 2.48 Design Parameters of Two Stage DTRO System for Leachate Treatment Designed parameters

Stage 1

Stage 2

Recovery rate of purified leachate QR0 (%)

80

90

108.7

86.7

Production of purified leachate Qp (m /d)

86.7

78

Membrane column nR0 /branch

46

9

9.405

9.405

Membrane filtration area SR0,t (m )

433

85

Operating pressure (Mpa)

5

3.5

Designed maximum operating pressure (MPa)

7.5

6

High-pressure pump

1

1

Inner online pumps

2

0

3

Influent flow Qd (m /d) 3

2

Single membrane filtration area SR0 (m ) 2

water receivers. The concentrated liquid is returned to the inlet side of the first stage of the DTRO system. The design parameters of the two stages of DTRO are presented in Table 2.48. The results showed that COD, BOD5, NH3-N, TN, and TP of the effluents were below 33.25, 23.94, 7.84, 9.8, and 0.1 mg/L, respectively, and SS can be around 0 mg/L.

2.11.6.3 Mobile Processing Equipment When the facilities and equipment fails, rainy season leachate increases, and equipment overloads; leachate treatment facilities are not perfect, the original leachate treatment equipment is not enough capacity, etc., it can be emergency treated by mobile devices, which is a new technology (Fig. 2.92). Mobile processing equipment advantages include easy transfer; high degree of automation, easy operation, less demand for running personnel, stable water quality, and is hardly affected by external factors. Membrane components are easy to maintain, with low operating costs.

2.12 MECHANICAL VAPORIZATON COMPRESSION (MVC) PROCESS 2.12.1 Introduction to Mechanical Vaporizaton Compression (MVC) Mechanical Vaporizaton Compression (MVC) is a technology developed at the time the traditional technology faced technical bottlenecks, and gradually got attention because of its remarkable economic, environmental, and social

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Two-stage DTRO + curling RO Concentrated liquid recharge

Original leachate tank

First-stage reverse osmosis

Concentrate storage pool

Second-stage reverse osmosis Landfill

Adjusting tank

Third-stage reverse osmosis

Clean water reserviors

Effluent water discharge

FIGURE 2.92 Process flow of mobile processing equipment for emergency.

benefits. MVC is developed by the American ship desalination institute, and has been applied for nearly 40 years. Currently, thousands of systems are operating in different industries and fields around the world. MVC— Deionization Ion Exchange (DI)—ammonium salt recovery technology, uses efficient MVC device to treat landfill leachate. The moisture and ammonia in the landfill leachate are separated from other substances, and the contaminant remained in the concentrate, as shown in Fig. 2.93. All the volatility of heavy metals and inorganic and organic substances are weaker than water, so they will be kept in the concentrate; only some volatile organic acids and ammonia and other pollutants enter steam, and eventually exist in distilled water. The volume of leachate can be concentrated to 2%B 10% of the original volume by the evaporation process. The ammonia nitrogen is trapped in the water by ion exchange resin so the ammonia nitrogen in discharged water could meet the standard. At the same time, the MVC exhausts ammonia and other volatile gases. Ammonia can be absorbed by the residual hydrochloric acid in the regenerated liquid of DI system, and then the saturated waste liquid is condensed and then crystallized into ammonium chloride crystal. MVC concentrate usually recharges to landfill. The main processing unit of the MVC is divided into the following parts. 1. Pretreatment system Removal of SS in leachate is the main purpose of the pretreatment process, tiny fiber could be removed by filter. At the same time, the

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FIGURE 2.93 Mechanical Vaporizaton Compression (MVC)
DI technological flow chart.

FIGURE 2.94 Working principle diagram of Mechanical Vaporizaton Compression (MVC).

content of the element such as Ca, Mg, which are easily to form the scale is reduced to prevent packaging the heat exchange tube in the evaporation process, and ensure the efficiency of the evaporation and normal operation of the pump. 2. MVC device The working principle of the MVC unit is shown in Fig. 2.94. The device utilizes the characteristics that the pressure and temperature is gradually enhanced when the vapor is compressed. Steam enters the

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Pollution Control Technology for Leachate From Municipal Solid Waste

evaporator’s heat exchanger tube. When the cold water is sprayed outside the tube, the steam condenses in the tube to form condensed water, and the heat is transmitted to the spray water outside the tube for continuous evaporation. Before entering the evaporator, leachate has the multistage heat exchange with the heat medium of MVC system, and then enters the MVC device. Using the principle of flash, ammonia and water evaporate at the temperature of about 103 C and are discharged from MVC device after condensing into distilled water. The material that is turned into a gas, condensed, and then exhaust out of the MVC device. The MVC makes full use of waste heat of the effluent, and increases the utilization efficiency of electric power and makes the temperature of discharged and distilled water 3 B 5 C higher than the water. The ammonia in distilled water is needed for further treatment by the ion exchange system. MVC concentrate recharges to landfill. 3. DI system The DI system adopts the large aperture and strong acid cation exchange resin, which is not easy to be blocked and has a smaller surface area, and it is not easy to absorb organic matter and easy to clean. The ion exchange reaction occurs when the condensate passes through the resin, and the NH1 4 in the water exchange with the hydrogen ion in the resin and then be removed. When the ammonia ions in the resin are saturated and the hydrogen ions of the resin are depleted, the resins can be regenerated using strong acidic reagents such as hydrochloric acid, and the regenerated resins could be reused. The regenerated liquid can be crystallized into ammonium salt products. 4. Crystallographic and evaporative ammonium recovery system The crystallographic and evaporative system is a forced circulation crystallization device of steam compression riser. The technology is very mature in application. The system uses the different solubility of ammonium chloride in different temperatures to condense the regenerative liquid and form crystal precipitation at 30 C or so, and the remnants of hydrochloric acid flow back to the acid tank and could be used for resin regeneration. The ammonium chloride is widely used in the production of battery, coating, and medicine and agriculture industries, and so on, which can produce economic benefits. “MVC 1 DI” technology is mature and has fewer components, and the evaporation (MVC) process can fully recover energy. Ion exchange (DI) process removes and recovers ammonia nitrogen in the leachate and turns into ammonium chloride crystals, not only solves the difficulty of high ammonia nitrogen in leachate and the problem of high cost, but also realizes the zero discharge of pollutants. The method of external evaporation is helpful to observe the operation, cleaning, and descaling. The technology also has some advantages, such as simple operation management, small land occupation, low

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running cost, less concentrate, and it is not affected by landfill age and seasonal change. It has strong adaptability, and the treatment facilities can be made into mobile units to save investment and operation cost. But there are some problems that need to be solved. The exhaust emissions are secondary pollution. Cation resin system need to be regenerated by hydrochloric acid after a certain time, and the waste liquid of regeneration is also secondary pollution. And, blocking problems occurs in the process of exchanging the large particle objects, and, also, replacement will be inconvenient and expensive if resin is discarded. The waste water containing hydrochloric acid in an ion-exchange system can cause corrosion and scaling. About 10% or so evaporation of the concentrate requires further treatment, which may be more difficult to treat than the concentrate produced by the membrane filtration. A very serious foam problem (it can only be solved by adding defoaming agent, which has a high cost) will occur when COD is relatively high, and it will directly affect the water quality.

2.12.2 Yunzaobang Waste Transfer Station Leachate Treatment by MVC Demonstration project with a designed leachate treatment scale of 150 t/d had been done in Yunzaobang waste transfer station to provide running data for leachate evaporation studies, to find the technical difficulties and solutions of MVC technology.

2.12.2.1 Pretreatment System The objective of the pretreatment system is to reduce the SS in leachate to below 50 mg/L, and the hardness (in CaCO3) to less than 100 mg/L. There were two stages of the test, the preprocessing of first phase used polyaluminum ferric chloride flocculent precipitation, but it can not satisfy the requirement of SS for the water inlet of MVC device, therefore ultrafiltration membrane filtration was added in the project of the second stage. The aperture of ultrafiltration membrane was 0.05B0.1 μm, the operating pressure was about 0.3 MPa, and the flux in stable operation was 200 L/(m2  h). The ultrafiltration membrane can replace the function of sedimentation tank, shorten pretreatment time, and further reduce the SS and hardness of the water inlet. Then, 2000 L leachate was pumped from the storage pool into the dosing flocculation barrel, and then 7.2 L 50% sodium hydroxide solution was added to adjust pH to 8.0. Next, 25 L 80% polyaluminum ferric chloride solution was added, and then they were fast mixed and allowed to stand for 10 minutes, then pumped to the ultrafiltration system. The leachate of the waste transfer station contained oil substance, and was easy to contaminate and clog the membrane hole. The method of reducing oil pollution was to adjust the pH of the leachate in the ultrafiltration membrane circulation tank

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to about 9.5 to improve the oil resistance of the membrane. Washing was needed at the end of daily operation in order to reduce corrosion and scaling of the equipment. 2 t water is required for each cleaning. After cleaning, 0.2 t water was injected into the hot well inside the main body of MVC to serve as the water source to preheat the machine at the next boot time.

2.12.2.2 Evaporation Device The MVC evaporator includes a degassing tower, MVC evaporator, compressor, heat exchanger, steam generator, and so on. Leachate gets out of the pretreatment system and then gets into the degassing tower, and sprayed from the upside of degassing tower, when hot steam is gushed up from the bottom of the tower. Some volatile organic compounds and parts of ammonia in leachate are removed and exhausted in the air from the leachate in the process of heat transfer. The leachate was heated to about 20 C. The preliminary preheated leachate goes through a heat exchanger, and the heat of the distilled water and concentrate is recovered, the leachate is heated to 95 B 100 C, so could evaporate as quickly as possible when it enters the main body of MVC device. Leachate gets into the MVC, spurts from the nozzle, and forms a thin film on the surface of circular pipes. The steam in the circular pipes is pressurized and heated by the compressor to 103 6 1 C. After heat exchange, most water and the organic matters, whose boiling point is less than or closed to the water, forms steam, less volatile organic matters and heavy metal ions form concentrate and enter the hot well, and water vapor in the piping is cooled to liquid and distilled water. The concentrate and distilled water are respectively passed through the heat exchanger to preheat the leachate from the degassing tower. The test device was not equipped with DI (ion exchange) and other follow-up equipment, and it was only used to study the efficiency and processing effect of evaporator. 2.12.2.3 Physical and Sensory Indicators of Discharged Water The physical and sensory indicators mainly refer to color and odor. Water discharged by MVC is colorless and transparent, as shown in Fig. 2.95. The color determined by platinum-cobalt methods was between 0 B 5 Hazen. The water had no visible particles and no precipitation after standing for 30 days or more, and had a slightly burnt smell. Concentrate discharged by MVC was black or brown. It was cloudy and there were a large number of impurities, such as suspended materials. The color determined by platinumcobalt methods was between 20 B 25,000,000 Hazen, and it had an obvious burnt smell. 2.12.2.4 Rate of Clean Water The rate of clean water refers to the percentage of clean water and the total amount of treated leachate of the device during the operation. The rate of

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165

Clear water rate of MVC (%)

FIGURE 2.95 Leachate samples for transfer station leachate treatment by Mechanical Vaporizaton Compression (MVC) (from left to right: MVC clean water, water after pretreatment, water after degassing tower, MVC concentrate).

100 90 80

93.7

92.1

92.0

25th

26th

27th

79.3

76.9

65.2

70 56.6

60 48.4

50

39.1

40 30 19th

20th

21st

22nd

23rd Time (d)

24th

FIGURE 2.96 Clean water rate of Mechanical Vaporizaton Compression (MVC) for leachate from transfer station.

clean water of the MVC device during operation is shown in Fig. 2.96. From the field observation, sometimes water emission was greater than the amount of leachate flowing into the device at the same time, as around 0.2 m3 tap water was stored in the hot well for convenience of starting up the machine next time in the process of shutdown cleaning. If the temperature of the MVC reached the temperature of evaporation, the leachate vapor became clean water and discharge in the heat exchange process.

2.12.2.5 SS SS could be removed from leachate completely by MVC device. As shown in Fig. 2.97, the SS after the pretreatment is basically between 481 and

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Pollution Control Technology for Leachate From Municipal Solid Waste

FIGURE 2.97 SS concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

706 mg/L. The SS value changed without a certain rule after the degassing tower, and its value fluctuated between 196 and 1983 mg/L. After entering MVC, the distilled water and the concentrate were formed after evaporation and concentration of the leachate. The SS of clean water was reduced to 0 mg/L, while the SS in the concentrate reached above 602 mg/L, exceeding the concentration of the leachate. This indicated that the SS fluctuation of the leachate had no effect on the SS value of the clean water; it can always stay close to the concentration of 0 mg/L, and the removal rate of SS will reach 100%.

2.12.2.6 pH The pH of fresh leachate was between 4.0 and 4.5. The sodium hydroxide was added in the pretreatment and ultrafiltration process to adjust the pH to about 9.5. After pretreatment, the increased pH value of the leachate was 0.09B1.71. The increased pH value of the degasification tower after the MVC was between 0.21 and 1.26. In the whole MVC process, the changes of pH value didn’t show a mathematical regularity, nor did it show the changes of the nature, as shown in Fig. 2.98. Although there were many factors that can result in the changes of pH, but it can be concluded that pH of the leachate won’t be drastically changed by MVC, the pH of concentrate and water was between 9.21 B 10.08.

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FIGURE 2.98 pH of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

2.12.2.7 COD The COD of the clean water is between 158B784 mg/L, and the COD can be stable between 300B750 mg/L in 80% of the time, and the removal rate of COD is 95.39%B99.15%, as shown in Figs. 2.99 and 2.100. COD of fresh leachate in the waste transfer station was 13,000 B 28,000 mg/L. COD of leachate was between 14,504 and 25,480 mg/L after pretreatment, because much of the particulate matter and suspended material was removed by coagulation precipitation and ultrafiltration. The value of COD was between 14,504 and 39,395 mg/L after treatment with degassing tower, which was not very different from that of pretreatment, and the purpose of ther degassing tower was not to remove the COD. After entering MVC, the leachate was finally discharged in two forms, concentrate and water, after evaporation and condensation. It can be seen from Fig. 2.100 that MVC has a higher removal rate of COD, the fluctuation of water quality would not affect the removal rate of MVC, and the removal rate was more than 95 percent overall. However, the high removal rate does not mean that the COD can meet the emission standard. Clean water discharged by MVC contained a high concentration of organic matters, which were mostly small molecule volatile organic compounds and needed further processing, such as RO, NF, DI, or the physical and chemical process, such as activated carbon adsorption to reach corresponding emission or recycling standards. The COD of the concentrates was between 25,786 B74,480 mg/L, and the value of COD increased by 1.47
3.19 times compared with the leachate after pretreatment. Although

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100.00 99.15

1150 96.92

99.00 98.46

97.15

97.84

97.87

900

97.04

96.75

98.00 97.00 96.00

650

95.39

95.00 94.00

400

93.00

150

92.00

19th

20th

21st

22nd

23rd

24th

25th

26th

COD removal rate of MVC/%

COD of MVC clean water/(mg/L)

FIGURE 2.99 COD concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

27th

Time (d)

FIGURE 2.100 COD value and removal rate of clean water for leachate over operating time using Mechanical Vaporizaton Compression (MVC).

the mass of concentrates was greatly reduced compared with the RO/NF membrane treatment process, the concentration had been increased and further treatment was needed.

2.12.2.8 TOC TOC concentration in clean water was 14 B 245 mg/L, compared to the pretreatment, the removal rate of TOC was 96.18% B 99.83%, TOC in the clean water was below 200 mg/L in 90% time. It can be seen from Fig. 2.101 that TOC of fresh leachate after pretreatment was 5 675 B 9

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FIGURE 2.101 TOC concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

FIGURE 2.102 TOC value and removal rate of clean water for leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

952 mg/L, after degassing tower was 4 889 B 18 260 mg/L, and there was little change before and after the degassing treatment. After entering MVC, the leachate was discharged in two forms, concentrate and water. The TOC value in clean water was between 14 and 245 mg/L, as shown in Figs. 2.101 and 2.102.

2.12.2.9 NH3-N and TN The NH3-N concentration of clean water was 2.20B3.10 mg/L, and the removal rate of NH3-N was 76.35%B82.56%. The TN concentration of

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6.00

0.86

0.84

5.00 0.83

0.83

0.84

0.82

4.00

0.82

0.80

3.00

0.80

0.80 0.79

0.78

0.77

2.00

0.76

0.76

1.00

NH3-N of removal rate/%

NH3-N of MVC clean water/(mg/L)

FIGURE 2.103 NH3-N concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

0.74

0.00

0.72

19th

20th

21st

22nd

23rd

24th

25th

26th

27th

Time (d)

FIGURE 2.104 NH3-N value and removal rate of clean water for leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

clean water was 4B133 mg/L, and the removal rate of TN was 91.41%B 99.52%, as shown in Figs. 2.103
2.106. It can be seen from Fig. 2.103 that the ammonia nitrogen concentration was 10.37B18.04 mg/L after pretreatment, and the total nitrogen concentration was 827B2 076 mg/L. After degassing tower, ammonia nitrogen concentration was 14.01 B 16.70 mg/L, the total nitrogen concentration was 606B2556 mg/L. The concentration of ammonia nitrogen and total nitrogen was changed, because in the degassing tower. On one hand, large molecules of nitrogen organics would decompose into small molecules in high temperature; on the other hand, there was a

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171

FIGURE 2.105 TN concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

99.47% 98.66%

TN of MVC clean water/(mg/L)

98.00%

99.52%

250

98.98% 200

96.00%

94.94%

93.56% 93.43%

150

100.00%

94.00%

92.77% 92.00%

91.41%

100

90.00%

50

TN removal rate of MVC/%

300

88.00%

0

86.00%

19th

20th

21st

22nd

23rd

24th

25th

26th

27th

Time (d)

FIGURE 2.106 TN and removal rate of clean water for leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

certain amount of ammonia gas emitted from the leachate. The NH3-N concentration in MVC clean water was 2.20B3.10 mg/L, and the removal rate of ammonia nitrogen was 76.35%B84.04%. The TN concentration in clean water was 4B133 mg/L, with a TN removal rate of 91.41%B 99.52%. Ammonia nitrogen concentration in the MVC concentrate was between 12.60 and 32.60 mg/L, and it was 1.3B2.5 times of the ammonia nitrogen concentration in the concentrated solution after the pretreatment when the operating conditions were stable. The total nitrogen concentration in the

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FIGURE 2.107 TP concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

MVC concentrate was 1 478B4 460 mg/L, and it was 1.3B4.5 times of the total nitrogen concentration in the concentrated solution after the pretreatment when the operating conditions were stable. Total nitrogen concentration was extremely high, but ammonia nitrogen concentration was lower, showed that most of the nitrogen existed in the form of organic nitrogen.

2.12.2.10 TP The total phosphorus concentration of the MVC clean water was 0B0.67 mg/L, and the removal rate of total phosphorus reached to 91.2%B 100.0%. It had excellent treatment effect of total phosphorus. It can be seen from Fig. 2.107 that the total phosphorus concentration after the pretreatment is between 4.05B16.52 mg/L, and the total phosphorus concentration in leachate has not changed significantly after the degassing tower. The total phosphorus concentration in the MVC concentrated solution was significantly higher than that of pretreatment, and its value was 7.14B17.54 mg/L, which was 1.1B2.7 times higher than the total phosphorus concentration after the pretreatment. This showed that the phosphorous substance in leachate cannot be vaporized with the water in MVC, and was present in the concentrate in the form of phosphate finally. And the flocculation and ultrafiltration of pretreatment had more obvious effects to the removal of total phosphorus in leachate. 2.12.2.11 TDS and Electric Conductivity It can be seen from Fig. 2.108 that the total dissolved solid (TDS) concentration of leachate after the flocculation and ultrafiltration is between

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FIGURE 2.108 TDS of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

11,100 B 17,500 mg/L, and the TDS value has not changed much after the degassing tower. After entering the main body of MVC, the TDS value was significantly different between the clean water and the concentrate. TDS concentration in clean water ranged from 57B129 mg/L. And it was high in the concentrated solution of MVC, reaching to 21,500 B 34,000 mg/L, indicating that MVC had a special advantage in lowering the content of solid substances in leachate. The electric conductivity reflects the amount of conductive ions in the water. It can be seen from Fig. 2.109 that the conductivity of clean water is between 88.5B123.8 μS/cm. The TDS value of clean water had a high correlation with conductivity value. The ratio of TDS to conductivity was roughly 0.64B 0.78.

2.12.2.12 Hardness Hardness had a huge impact on scaling of the main equipment of the MVC, equipment cleaning frequency and degree of corrosion were affected by the scaling speed in normal work, so the hardness index of the water was one of the key considerations of MVC. The amount of Ca and Mg elements was measured, and was converted to CaCO3 for hardness. The Ca and Mg elements concentration in each work procedure is shown in Figs. 2.110
2.113. The hardness indicators are shown in Fig. 2.112. According to the data in the Fig. 2.113, the Ca elements concentration in the leachate after pretreatment was 47.72B156.75 mg/L, the Mg element concentration was 8.81B38.84 mg/L, and the corresponding hardness was 156.17B543.57 mg/L.

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FIGURE 2.109 Correlation between TDS and conductivity for leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

FIGURE 2.110 Ca concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

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FIGURE 2.111 Mg concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

FIGURE 2.112 Hardness of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

The required total hardness of the test equipment was less than 100 mg/L, so the pretreatment process did not meet the requirements. The Ca and Mg elements’ concentration decreased to a certain extent after the degassing tower. The hardness of the leachate after degassing tower was 62.53B563.40 mg/L (in CaCO3), and the removal rate of hardness was 0%B 68.8%. Ca and Mg elements concentration of clean water was very

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FIGURE 2.113 Hardness value and removal rate of clean water of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

low, the corresponding hardness was 0.86 B 16.36 mg/L, with a removal rate of Ca and Mg elements of 95.2% B 99.9% and 94.2% B 99.9% and a removal rate of hardness of 94.6% B 99.0%. Both the Ca and Mg elements’ concentration in the concentrate showed an upward trend. Low hardness of clean water can reduce the scaling speed of the inner wall of the circulation tank, but high hardness of concentrate can quicken the scaling speed of outer wall of the circulation tank, and reduce the evaporation efficiency of the MVC. It was a problem needed to be overcome in the MVC process.

2.12.2.13 Metallic Elements There were 7 kinds of high concentration metal elements in the leachate after pretreatment, with a concentration sequence from high to low of K, Ca, Na, Mg, Fe, Al, and Mn. K concentration was more than 200 mg/L, Ca, Na concentration were between 90 B 200 mg/L, Mg, Fe concentration were between 10B60 mg/L, Al, Mn concentration were between 0.5 B 3 mg/L. Other metals concentrations were between 0 B 0.3 mg/L, with a concentration sequence from high to low of Zn, Sr, Cr, Ni, Ba, V, As, Pb, Cu, Co, Cd, Se, B, Li. The metal elements’ concentration in MVC clean water were significantly reduced, and the order of the first seven elements concentration was Ca, Mg, Fe, K, Al, Zn, and Cr. The other elements’ concentration in MVC water was very small, and the concentration range was basically 0B0.05 mg/L. The removal rate of the monovalent metal ions was far higher than that of bivalent or above metal ions in MVC. K concentration in clean water was 0B 0.99 mg/L, with a removal rate of 99.6B100.0%. The removal rate of Na reached 100.0%. The Fe concentration in clean water was 0B1.92 mg/L, with a removal rate of 96.0B100.0%. The Fe concentration in the concentrate was 4.67B39.55 mg/L, which increased by 0.20B1.65 times compared with leachate after pretreatment, possibly due to the measuring error or deposition of iron elements in alkaline environment (Fig. 2.114).

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FIGURE 2.114 Fe concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

FIGURE 2.115 Al concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

It can be seen from Fig. 2.115 that Al element concentration in the leachate after pretreatment is between 0.00B2.58 mg/L, in MVC clean water is 0B1.39 mg/L. The removal rate of Al element in the MVC process was 42.0%B 100.0%. Al element concentration in the concentrate was 0.00B3.69 mg/L. It can be seen from Fig. 2.116 that Mn element concentration in leachate after pretreatment is between 0.25 B 2.29 mg/L, in the leachate after degassing tower was 0.13 B 2.16 mg/L, in MVC clean water was 0.00 B 0.03 mg/L,

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FIGURE 2.116 Mn concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

and the removal rate of Mn element was 98.7% B 100.0%. Mn element concentration in the concentrate was 0.12B1.94 mg/L. Other metal elements concentration in MVC clean water were small. The total cadmium concentration was 0 B 0.01 mg/L, total chromium concentration was 0.11 B 0.21 mg/L, total arsenic concentration was 0.21 B 0.07 mg/L, total lead concentration was 0 B 0.22 mg/L, other metal elements concentration were almost undetectable. It can be seen from Figs. 2.117 and 2.118 that various elements’ concentration of ultrafiltration membrane concentrate is high, indicating the process of flocculation and filtration in pretreatment has obvious removal effect on various kinds of metal elements in leachate.

2.12.3 Indexes Summary of Discharged Water Treated by MVC Process The quality of discharged water after pretreatment, clean water of MVC, and concentrates of MVC is listed in Table 2.49.

2.13 ISOLATION OF DISSOLVED HUMIC ACIDS FROM NONDEGRADABLE LANDFILL LEACHATE HSs are defined as a fraction of the organic matter characterized by its yellow-to-dark color, high molecular weight, acidic properties, and high degree of decomposition. HSs are ubiquitous in the environment, occurring in soils, waters, and sediments of the ecosphere. They are a product of the decomposition of organic materials, including animal and plant tissues,

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179

FIGURE 2.117 Macroelements concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

FIGURE 2.118 Microelements concentration of leachate from transfer station over operating time using Mechanical Vaporizaton Compression (MVC).

bio-wastes, microfauna, etc. HSs can be divided into three categories (humin, humic acid, and fulvic acid) according to their solubility at different pH values. For fulvic acid (FA), it is soluble at any pH values, and its color varies from yellow to brown black and its molecular weight ranges between 175

TABLE 2.49 MVC Discharge Indices of Leachate From Transfer Station Over Operating Time Using Mechanical Vaporizaton Compression (MVC) Pollutant category

Pretreatment leachate after flocculation 1 ultrafiltration

Clean water of MVC

Concentrated liquid of MVC

pH

9B10

9B10

9B10

Color(platinum-cobalt methods)

1 250B5 500

0B2

20 000B25 000

COD (mg/L)

14 504B25 480

158B784

25 786B74 480

Suspended solids (mg/L)

481B1 292

,5

602B7 054

TN (mg/L)

827B2 076

4B133

1 478B4 460

NH3-N (mg/L)

10.37B18.04

2.20B3.10

12.60B32.60

TP (mg/L)

4.05B16.52

0B0.67

7.14B17.54

Cd (mg/L)

0B0.04

0B0.01

0B0.01

Cr (mg/L)

0.08B0.41

0B0.17

0.11B0.21

As (mg/L)

0B0.05

0B0.01

0.02B0.07

Pb (mg/L)

0B0.30

0B0.28

0B0.22

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FIGURE 2.119 Procedure for the isolation of dissolved humic acids from nondegradable landfill leachate.

and 3570 Da with highly acidic properties. For humic acid (HA), it is soluble only at pH values above 2, with dark brown-to-black color and molecular weight ranging from several hundreds to thousands of Daltons. For humin, it is insoluble at any pH values, and its characteristics and properties is not well known. It has been well proven that leachate can be biologically treated from COD 20,000 to 250B600 mg/L and further reduction is very difficult. The isolation of organic matter was conducted using XAD resin, aiming to confirm the presence of HSs in the leachate and its contribution to the COD value. Procedure for the isolation of HSs is given in Fig. 2.119. The leachate influent and effluent of the aged-refuse-based biofilter was used as a leachate sample in order to know the removal of HSs by the biofilter (Fig. 2.119). The leachate sample was first filtered by 0.45 μm polymembrane in order to remove suspended solid (SS) and humin, and then passing through XAD-8 resin (a cross-linking organic resin produced by Nanjing University in China) after acidifying, in which HA and FA was retained in the resin. The COD and TOC value of sample were analyzed by standard methods.

2.13.1 Separation of Humus (HSs) From Leachate by Resin The COD value of landfill leachate was 1810 mg/L in the influent. After the removal of SS by filtration, the COD dropped to 1737 mg/L, indicating that the SS contributed to little percent of the COD value. Then the leachate was passed through XAD-8 resin, in which humus (HA and FA) were retained. It can be seen from Table 2.50 that half of the COD value was consisted of HSs. Table 2.51 shows the COD and TOC values of another leachate effluent after the isolation of hydrophobic acid and hydrophilic acid. It can be seen

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TABLE 2.50 COD Changes in the Isolation Process of HSs of Sample 1 (BOD5 of the Influent: 0 mg/L) Items

Influent

After filtration (Before XAD-8 resin)

After XAD-8 resin

Percentage of HSs’ COD to the total (%)

COD (mg/L)

1810

1737

1071

40.8

TABLE 2.51 COD and TOC of Leachate Effluent of Sample 2 (BOD5 of the Influent: 0 mg/L) Items

Influent

After filtration (before XAD-8 resin)

After XAD-8 resin

Percentages of HSs’ COD to the total (%)

COD (mg/L)

245.9

243.1

157.6

35.9

TOC (mg/L)

98.09

97.03

74.80

23.7

that the leachate effluent COD can be reduced from 245 to 157 mg/L after removal of HSs using the resin, indicating that the HSs accounted up to 35.9% of the COD value. TOC in leachate are of great interest due to its ability to form complexes with heavy metals and to enhance solubility of hydrophobic compounds. Moreover, HSs in leachate from MSW landfills may represent up to 72% of the total DOC. In order to investigate the contribution of the HSs in leachate to its total TOC, the concentrations of TOC were measured in leachates before and after filtration by XAD-8 resin. Then, HSs carbon proportion in TOC of each leachate could be calculated. This proportion will give an idea of the representation of HSs to the total TOC in leachate and therefore its real significance in the interaction between organic matter and pollutants in the landfill. From the results presented in Table 2.51, it can be seen that in leachate with COD of 245.9 mg/L, it had a TOC of 98.09 mg/L, and the contribution of the HSs to TOC in leachate was 23.7%.

2.13.2 Elemental Composition of Humus (HSs) The results of elemental analysis of HA in leachate and standard HA from company are shown in Table 2.52. It was found that HA from leachate had

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TABLE 2.52 Elemental Composition of Humic Acid (%) Sample

C

N

S

H

C/N

Standard HA

48.73

0.686

1.504

4.843

71.09

HA

50.74

6.605

3.033

5.678

7.682

50.74% carbon, a little higher carbon content than that of standard HA. Regarding hydrogen, HA from leachate with 5.68% had little higher content compared with standard HA. Meanwhile, the contents of nitrogen of HA in leachate were almost ten times than that of standard HA, while the content of sulphur were only two times of that of standard HA. Nondegradable leachate treatment is a worldwide concern. The separation of humic acids may provide useful information for the development of innovative treatment technology of leachate. A resin was used for the selective separation of humic acids and its contribution to the COD value in landfill leachate was determined. It was found that the percentage of HSs contributed to 40% of the COD value in leachate influent, and to 35.9% of the COD value in the effluent. Nearly half the percentage of the total organic carbon constitutions consisted of HSs.