Leachate Quality

10.2 LEACHATE QUALITY Hans-Jürgen Ehrig and Rainer Stegmann

INTRODUCTION Leachate quality is the result of the waste composition, water budget, and the biological, chemical, and physical conditions in the landfill body. Until now it is not possible to predict the leachate composition of a new landfill, and it is very difficult to describe the future leachate pollution of an existing landfill. As a consequence of the high organic waste content, uncontrolled anaerobic biological processes develop during the first up to 10 years of landfill operation, whichdwith the influence of chemical/ physical processesddetermine the leachate quality particularly the organic carbon content of leachate. In addition, these processes influence the pH and the mobility of a variety of other compounds as iron, calcium, etc. Reducing the period of the anaerobic phase in a landfill results in a significant reduction of the leachate pollution. After the end of landfill operation, the leachate pollution will remain at high levels over a very long period of time (decades and longer) with only a slow decrease. Leachate is a specific wastewater with a wide range of variation in quality and quantity. The leachate production rate is a result of climatic conditions (e.g., precipitation, evaporation) and landfill specifics (e.g., infiltration, storage). Waste emplacement technology as thickness of waste layers, compaction of waste, kind of daily soil cover etc., influences the flow conditions that affect the contact between water and waste. This contact is the basis of most processes in the landfill body and the resulting leachate quality. As a consequence of the infiltrated water and the reactive organic and inorganic waste compounds, a wide range of biological, chemical, and physical processes take place in the landfill. With increasing content of organics, the influence of biological processes on leachate composition becomes more dominant. In addition, processes such as solubility, fixation, complexation, precipitation, etc., influence the leachate composition. The interactions between the different processes are extremely complex and not predictable. As a result a nearly accurate prediction of flow rate and composition is in general very difficult. The environmental effect of leachate varies between very low and hazardous also depending on the specific situation. Hazardous waste material in landfill does not automatically result in hazardous leachate pollution. Landfills are a very slow reacting system with the effect that the annual transfer rate of substances to the environment is relatively low. The comparatively small amounts of hazardous

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511

substances (e.g., batteries, medicine, detergents) that are part of municipal solid waste (MSW) do very often not affect the leachate concentration significantly, even over a very long period of time. After the landfill operation has been terminated and the landfill has been closed, the concentrations of most substances in the leachate decrease with time. From that time on no more pollution potential is added to the landfill, and due to leachate and gas production the landfill emission potential will be steadily but slowly reduced. On the other hand as a result of degradation processes and changing environmental conditions in the landfill body, compounds (e.g., ammonium, reduced sulfur components) may also after landfill closure even increase over a limited period of time.

BIOLOGICAL PROCESSES AND LEACHATE QUALITY Most MSW contains a high amount of organic degradable waste. The waste organics are degraded in the landfill by biological processes in an uncontrolled way. The kind and intensity of these processes depend on the environmental conditions such as access of oxygen, temperature, moisture content and movement, biodegradability of organics, bioavailability etc. After waste emplacement, these processes are mostly aerobic over a limited period of time. Also when the waste layers are highly compacted, air can penetrate up to 1e2 m deep into the landfill; below the top waste layers the processes change to anaerobic. The influence of aerobic processes in low compacted landfills is even higher but hardly predictable. During the anaerobic degradation process, mainly two defined groups of microorganism are degrading the organics; they should be in an equilibrium, otherwise unwanted intermediate degradation products will be produced. The control of this equilibrium is possible in landfills only to a very limited extent. The consequences may be very high organic leachate concentrations over a limited period of time. In Fig. 10.2.1 (left side) the steps of the anaerobic biological degradation are shown. In a first phase the waste organics are hydrolyzed to smaller organic products. These intermediate products are reduced in the acidogenesis step to mostly volatile fatty acids (VFAs), alcohols, and inorganic residuals. Many products of hydrolysis and acidogenesis are in the liquid phase and leave the landfill as leachate. The effect of these processes on leachate quality is presented in Fig. 10.2.1 (right side). The organic pollution [chemical oxygen demand (COD) and biological oxygen demand (BOD5)] during the first years of landfill operation is in general very high. This is especially the case if the content of degradable organics is high. During this phase, COD concentrations may in some cases increase to up to 100,000 mg/L and represent mainly easy degradable organic acids. As a result the pH values are often in the range of 4.5e6.5 (Fig. 10.2.1 right side). The next steps of the anaerobic degradation process are acetogenesis and methanogenesis. Both processes must be in equilibrium: the microorganisms active during the acetogenesis metabolize the VFA and alcohols to acetic acid and H2. These are the substrates for the methanogenic microorganism, produced as end products of fermentation biogas (40%e75% methane and 60%e25% carbon dioxide). Both processes are very sensitive, and the growth rate of the microorganisms is very low. Both groups of microorganisms may be inhibited particularly by the intermediate product VFA at higher concentrations. Hydrogen as an end product of acetogenesis

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

Figure 10.2.1 Sequence of anaerobic biodegradation (left side) and the effect of incomplete biodeg-

radation on organic leachate pollution (typical older landfill with waste disposal in 2-m lifts, waste density 0.8e1.0 t/m3 and 750 mm/year precipitation (right side) (Ehrig, 1983). BOD, biological oxygen demand; COD, chemical oxygen demand. may inhibit the microorganisms of acetogenesis if the hydrogen is not converted to methane and by these means accumulate (see also Chapter 3.1). After some years of landfill operation, the easy biodegradable organic waste compounds will be reduced. The methanogenic bacteria may develop in areas with lower concentrations of organic acids as in inorganic waste compartments (e.g., soil, plastic), etc., and reach out from there into areas with higher acid production. This means there will be in the same landfill areas of acid and methane production, where with time the methanogenic conditions will be dominant. As a consequence, at some stage there will be a sharp decrease of organic leachate concentrations (Fig. 10.2.1dright side). Somehow simplified the first phase with high organic concentrations is called acetic phase, and the subsequent phase with highly reduced organic degradable compounds is called methanogenic phase. The time until landfills reach the methanogenic phase differs in a wide range. Observations at different landfills in Germany have shown periods between 2 and 12 years until a stable methanogenic phase had been reached. Robinson (1995) reported that in the U.K. 33% of the observed landfills have reached the methanogenic phase after 2.5 years, 73% after 4 year, 80% after 7 years, and 93% after 10 years. Looking at the scheme of anaerobic degradation in Fig. 10.2.1, the end productdwith exception of some inorganic and difficult or not degradable organic substancesedis mainly biogas. Leachate from landfills in the methanogenic stage shows relatively high COD but low BOD concentrations.

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The residual COD consists to a high degree of humic- and fulvic-like acids and may reach values between several hundred mg/L and 5000 mg/L. But in opposite to the organic leachate compounds produced in the acetic phase, these residual organics are less biodegradable or nonbiodegradable. These differences in biodegradability are used to characterize the anaerobic phase by means of the BOD5/COD ratio. The acetic phase is characterized by a BOD5/COD ratio >0.4 and the methanogenic phase by a BOD5/COD ratio
0.1; a ratio between 0.4 and 0.1 characterizes the transitional phase. This phase can often be observed at landfills because the collected leachate represents a mix of the different leachate qualities from the different landfill sections that are of different age. Leachate samples from boreholes in the upper part of a landfill in the methanogenic phase in some cases show high organic leachate concentrations typical for the acetic phase. In such cases the fully established anaerobic microorganism community of the lower layers reduces the intermediate products from above layers. The change of the leachate quality as a result of different biological phases in the landfill is a great challenge when planning a leachate treatment plant. Owing to this fact, it needs the treatment of two different wastewaters. This problem leads to considerations to reduce the time period of the acetic phase or even to prevent it. In the following, some options are presented: • Recirculating the leachate to a landfill part with fully established methanogenic phase (prevents overloading, this part acts as an anaerobic filter) (see Chapters 10.3, 12). • Installation of a 2-m layer of composted waste with a coarse structure as the first lift at the bottom of the landfill that will act as a biological filter. • Degradation of the easy biodegradable organic waste prior to landfilling, i.e., mechanical biological treatment (MBT), or on site composting before compaction. • Separate collection and treatment of the food and yard waste. • Improving the environmental conditions for microorganisms with increased moisture content (controlled leachate recirculation). • Slow built-up of the landfill to final height. Fig. 10.2.2 (left side) shows the COD leachate concentration pattern of six different landfills. These are operated in three different modes where always two landfills were operated in the same way. In the past, in Europe, waste was mainly emplaced in 2-m layers; this has been changed to placement and high compaction of waste in thin layers (50 cm). As a consequence of the old operation mode, in many cases high organic leachate concentrations over several years have been observed. At landfills where the waste is compacted in thin layers, the density will be higher. In additiondas the results in Fig. 10.2.2 showdthis operation mode has a positive environmental effect. Compared to the emplacement in 2-m lifts the COD concentrations in the leachate are lower and the duration of the acetic phase is shorter. The waste input of both landfill examples is relative low, which results in a relatively slow built-up of the landfill to final height. In Fig. 10.2.2 (right side) an estimated uncontrolled biological degradation of organic matter in the surface layer is shown. Measurements at landfills show that the air penetrates about 1e2 m into the landfill. Having installed 2-m layers, some parts of each layer have only limited access to air. But if there is a 0.3- to 0.5-m layerdeven with a higher densitydthe entire lift has contact with air as

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

Figure 10.2.2 Effect of waste lift height and leachate recirculation on organic leachate qualitydtwo landfills of each type (left side). Estimated uncontrolled aerobic biodegradation in surface landfill layers dependent on the increase of landfill height (right side) (Ehrig, 1989). BOD, biological oxygen demand; COD, chemical oxygen demand.

long as it is exposed to the atmosphere. The lower the increase of landfill height with time, the longer is the exposure of the waste to air with the consequence of a relatively fast aerobic degradation of the easily degradable organics. After covering such a layer with new waste, the environmental conditions change to an anaerobic milieu. But since the organic matter is partly aerobically degraded, the growth of acidogenic microorganisms is reduced and less organic acids are produced. As a consequence, the environmental conditions (e.g., pH, organic acid concentrations) for the development of methanogenic microorganisms will improve. If daily cover is used (which is not practiced in Germany at most cases), the soil should have a relatively high permeability. At several landfills, leachate is/has been recirculated back to the landfill. In the early days, landfill operators practiced it for economic reasons because they expected as long as the total amount of leachate was recirculated there are no leachate treatment costs. But experience shows that on the longer run this practice caused a lot of problems due to elevated leachate production, water built up in the landfill, and seepage of leachate through the slopes. As a consequence the amount of recirculated leachate should be controlled, and the process has to be monitored. In addition, only treated leachate with low organic acid concentrations should be used (see Chapter 12). On the other hand, Fig. 10.2.2 (left side) shows that the intensity and time of the acetic phase will be reduced during a controlled recirculation. This is due to the fact that the water flux through the landfill body (i.e., products of microbial activity are transported faster out of the landfill body) and also the moisture content of waste increase. Owing to the partly reduction of VFA in the recirculated leachate, the pH increases with additional

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positive effects on anaerobic degradation in lower layers of the landfill body. At very fast growing landfills (10 m/year), leachate recirculation has shown no effect because the concentration of organic acids is high, and the increase of the water addition is not sufficient to leach out significant amounts of organic acids. In addition the aerobic phase in the surface layer is too short. At high precipitation rates (e.g., >1200 mm/a), recirculation is not effective any more (more information on leachate recirculation is given in Chapter 12). At several landfills the leachate from the acetic phase section of the landfill (young part) is recirculated to older sections of the landfill where the methanogenic phase is already established. Some indications on the treatment capacity of the latter section are presented in Chapter 10.5. Another possibility to reduce or prevent the production of high-strength acetic leachate is the emplacement of a layer of composted waste as a first lift at the bottom of the landfill. Owing to the reduced organic content of the composted waste (it should have a coarse structure) and its positive environmental conditions for methanogenic microorganisms (e.g., pH), the high organic concentrations of the leachate from the upper layers can be partly anaerobically degraded. It should be considered that composted waste is not a uniform material compared to a fixed film reactor with exact given flow path. Also for this reason a 2-m layer of composted waste should be realized. At very fast growing landfills of 10 m/year and more, the thickness of composted waste should be increased up to 4 m.

LEACHATE QUALITY As a consequence of the waste composition, the environmental conditions and the landfill operation leachate quality can vary in a wide range. Tables 10.2.1e10.2.4 present leachate quality data from landfills in several countries. Box 10.2.1 shows median, 75% quantile, and maximum leachate data from 55 MSW landfills in North Rhine Westphalia (Germany). The most important leachate compounds are organic substances represented by the COD and BOD5 and nitrogen compounds measured as ammonium and organic nitrogen. At first sight the frequently high COD and BOD5 values during the acetic phase seem to be the most important challenge. But these organics can be relatively easily biologically degraded. High COD leachate concentrations during the methanogenic phase require specific treatment for elimination. In addition to the COD the organic quality of wastewater is often measured as TOC (total organic carbon). To compare both values a relationship of the COD/TOC ratio amounts to approximately 2.7e3.3 and may be used to convert COD into TOC or vice versa. The COD/TOC ratio describes the oxidation stage of organic substances and it will be higher for non oxidised compounds (e.g. hydrocarbons). Inorganic nitrogen is a degradation product of different organic materials as there are garden and yard waste, textiles, paper, and food waste. In contrast to the organic pollution the nitrogen concentrations do not decrease during phase changes. In many cases a slight increase with landfill age can be observed. One reason for the different behavior of both components is the way they leave the landfill and their aggregate state. The products of biodegraded organics are in the liquid stage as acids and alcohols or in the gaseous stage as CH4 and CO2. The products of the biodegradation of organic

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

CHAPTER 10 j Leachate Quality

Table 10.2.1 Leachate analysis of German landfills Parameter

(1) Aver.

pH

(1) Min.

(1) Max.

(2) Aver.

(2) Min.

(2) Max.

6.1

4.5

7.5

8

7.5

9

(3) 1e5 years

(3) 6e10 years

(3) 11e20 years

(3) 21e30 years

(4) Min.

(4) Max.

7.3

7.5

7.6

7.7

5.4

9

BOD5

mg/L

13,000

4,000

40,000

180

20

550

2,285

800

275

185

6

16,000

COD

mg/L

22,000

6,000

60,000

3,000

500

4,500

3,810

2,485

1,585

1,160

22

22,700

NH4

mgN/L

(5)

750

30

3,000

405

600

555

445

0,4

7000

NO3

mgN/L

(5)

3

0.1

50

3.6

7.6

12

9

200

NO2

mgN/L

(5)

0.5

0

25

0.06

0.63

0.5

0.8

11.7

Organic N

mgN/L

(5)

600

10

4,250

Total P

mg/L

(5)

6

0.1

30

Alkalinity (as CaCO3)

mg/L

(5)

6,700

300

11,500

AOX

mg/L

(5)

2,000

320

3,500

2,765

1,930

1505

1,130

20

7,500

Cl

mg/L

(5)

2,100

100

5,000

1,300

2,135

1,760

1,025

13

28,000

SO4

mg/L

80

10

420

98

146

93

83

1.1

1,810

Na

mg/L

(5)

1,350

50

4,000

815

1,125

905

645

13

4,700

K

mg/L

(5)

1,100

10

2,500

1,220

910

695

595

25

2,200

500

70

1,750

517

(Continued)

Table 10.2.1 Leachate analysis of German landfillsdcont'd SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

Parameter

(1) Aver.

(1) Min.

(1) Max.

(2) Aver.

(2) Min.

(2) Max.

(3) 1e5 years

(3) 6e10 years

(3) 11e20 years

(3) 21e30 years

(4) Min.

(4) Max.

Mg

mg/L

470

50

1,150

180

40

350

290

205

145

115

15

1,167

Ca

mg/L

1,200

10

2,500

60

20

600

375

465

325

155

12

10,000

Zn

mg/L

5

0.1

120

0.6

0.03

4

1.1

1.5

0.53

0.54

Mn

mg/L

25

0.3

65

0.7

0.03

45

3.9

1.5

1.1

0.9

0.05

43

Fe

mg/L

780

20

2,100

15

3

280

50

15

9.9

8.7

0.08

550

As

mg/L

(5)

160

5

1,600

15

21

42

14

1

370

Cd

mg/L

(5)

6

0.5

140

11

4

4

3

0.13

70

Cr

mg/L

(5)

300

30

1,600

156

224

164

177

5

2,570

Cu

mg/L

(5)

80

4

1,400

711

73

62

36

2

40,000

Hg

mg/L

(5)

10

0.2

50

Ni

mg/L

(5)

200

20

2,050

199

154

135

115

3

1,930

Pb

mg/L

(5)

90

8

1,020

156

56

67

34

5

1,300

(1) and (2) Weekly analysis at 15 landfills over 6 years (Ehrig, 1980; Ehrig and Hagedorn, 1994), (3) and (4) leachate analysis of >50 landfills over the entire landfill age (Ehrig, 2001; Krümpelbeck, 2000). Aver., average; Max., maximum; Min., minimum; (1) acetic phase; (2) methanogenic phase; (3) average leachate analysis at landfills with an age of 1e5 years, 6e10 years, 11e20 years, and 21e30 years; (4) minimum and maximum values of all landfills (3); (5) the same values as during methanogenic phase (2).

CHAPTER 10 j Leachate Quality

Table 10.2.2 Leachate analysis from different countries: (1) acetic phase and (2) methanogenic phases leachate analysis from

UK landfills (Robinson, 1995), (3) French landfill (Amokkrane et al., 1997), (4) Italian landfill (Lopez et al., 2004), (5) and (6) Greek landfills (Tatsi et al., 2003; Loizidou et al., 1992), and (7) Algerian landfill (Salem et al., 2008) Parameter

(1) Min.

pH

(1) Max.

(1) Aver.

(2) Min.

(2) Max.

(2) Aver.

(3) Aver.

(4)

(5) Aver..

(6) Aver.

(7)

5.12

7.8

6.73

6.8

8.2

7.52

8.2

8.2

6.2

7.9

8.27

BOD5

mg/L

2,000

68,000

18,632

97

1,770

374

200

2300

70,900

1,050

980

COD

mg/L

2,740

152,000

36,817

622

8,000

2,307

4,100

10,540

26,800

5,350

3,792

TOC

mg/L

1,010

29,000

12,217

184

2270

733

1,430

3,900

NH4

mgN/L

194

3,610

922

283

2,040

889

1,040

5,210

3,100

940

85.8

NO3

mgN/L

<0.2

18

1.8

0.2

2.1

0.86

124

150

55

14.6

NO2

mgN/L

0.01

1.4

0.2

<0.01

1.3

0.17

TKN

mgN/L

Phosphate

mgP/L

0.6

22.6

5

0.3

18.4

4.3

Alkalinity (as CaCO3)

mg/L

2,720

15,870

7,251

3,000

9,130

5,376

Cl

mg/L

659

4,670

1,805

570

4,710

2,074

5,420

SO4

mg/L

<5

1,560

676

<5

322

67

550

Na

mg/L

474

2,400

1,371

474

3,650

1,480

3,000

1.75

8.4

3,400

1,100

32

167

8.8

21,470

12,880

4,950

4,900

3,260

4,120

4,569

210

3,056

3,970

58.2

(6)

519

(Continued)

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

Table 10.2.2 Leachate analysis from different countries: (1) acetic phase and (2) methanogenic phases leachate analysis from UK landfills (Robinson, 1995), (3) French landfill (Amokkrane et al., 1997), (4) Italian landfill (Lopez et al., 2004), (5) and (6) Greek landfills (Tatsi et al., 2003; Loizidou et al., 1992), and (7) Algerian landfill (Salem et al., 2008)dcont'd Parameter

(1) Min.

(1) Max.

(1) Aver.

(2) Min.

(2) Max.

(2) Aver.

(3) Aver.

(4)

(5) Aver..

(6) Aver.

K

mg/L

350

3,100

1,143

100

1,580

854

880

3,460

Min.

Max.

Mg

mg/L

25

820

384

40

1,580

250

110

24.1

85.2

140

Ca

mg/L

270

6,240

2,241

23

501

151

68

15.7

Mn

mg/L

1.4

164

32.9

0.04

3.59

0.46

Fe

mg/L

48.3

2,300

654

1.6

160

27.4

0.91

2.7

Zn

mg/L

0.09

140

17.4

0.03

6.7

1.14

0.73

0.16

As

mg/L

<1

148

24

<1

485

34

Cd

mg/L

<10

100

20

<10

80

15

100

<20

Cr

mg/L

30

300

130

<30

560

90

Cu

mg/L

20

1,100

130

<20

620

130

Hg

mg/L

<0.1

1.5

0.4

<0.1

0.8

0.2

Ni

mg/L

<30

1870

420

<30

600

170

810

310

Pb

mg/L

<40

650

280

<40

1,900

200

460

<30

Aver., average; Max., maximum; Min., minimum.

0.04

2,210 390

(7)

0.41 5

16.2

8.23 1.43

<100

<30

700

1,910

200

90

280

390

670

1350

370

<100

3,490

CHAPTER 10 j Leachate Quality

Table 10.2.3 Leachate analysis from different countries (Robinson, 2007; Robinson and Carville, 2007) (1) and (2) South Africa,

(3) Hong Kong, (4) Thailand, (5) Indonesia, (6) New Zealand, and (7) Korea Parameter

(1)

pH

(2)

(3)

(3)

7.5

8.2

8.6

7.8

(4)

7.6

(4)

7.2

(4)

7.0

(5)

(6)

(6)

(7) Min.

(7) Max.

8.4

7.2

7

7.5

8.3

TOC (mg/L) 968

76

737

342

1058

1,181

1,969

1,840

4,601

BOD5

mg/L

170

550

167

117

COD

mg/L

760

4,560

2,580

873

2,700

1,560

1,980

NH4

mgN/L

435

1,554

2,563

1,156

3,032

1,424

1,350

2,000

860

1,740

NO3

mgN/L

<0.1

9.2

2.5

<0.1

<1

<1

<1

<1

<1

3

NO2

mgN/L

0.4

<0.1

<0.1

0.7

<0.4

0.3

<1

<1

120

Phosphate

mgP/L

1.4

13

27.6

22.2

1.7

15

14

12

8

14

Alkalinity (as CaCO3)

mg/L

2,422

9,652

11,500

4,940

23,910

12,505

15,970

7,840

3,713

10,113

Cl

mg/L

1,690

4,626

2,740

821

3,802

2,498

3,650

2,330

859

973

667

2,156

SO4

mg/L

15

6.4

1.6

159

1

1

6

80

Na

mg/L

K

mg/L

590

2,825

2,100

217

2,453

1,460

2,179

1,130

669

429

550

1,707

1,615

1,000

375

1,932

1,010

1,819

1,600

471

649

436

1,123 (Continued)

521

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

Table 10.2.3 Leachate analysis from different countries (Robinson, 2007; Robinson and Carville, 2007) (1) and (2) South Africa, (3) Hong Kong, (4) Thailand, (5) Indonesia, (6) New Zealand, and (7) Koreadcont'd Parameter

(1)

(2)

(3)

(3)

(4)

(4)

(4)

(5)

(6)

Mg

mg/L

80

195

31

18

121

132

182

56

Ca

mg/L

105

198

19

22

55

126

199

86

Mn

mg/L

0.86

0.24

0.6

1.65

0.47

0.4

6.56

Fe

mg/L

12/ 18.8

Zn

mg/L

0.17/ 0.12

As

mg/L

62

Cd

mg/L

<1

<50

<50

<50

Cr

mg/L

80

780

<500

Cu

mg/L

<10

<50

Hg

mg/L

<50

Ni

mg/L

120

Pb

mg/L

<4/7.3

Max., maximum; Min., minimum.

9.35

64

20

95

(6)

160

5.5

7.8

2.77

1.57

3.08

6.23

0.2

0.89

2

0.9

0.15

0.61

0.24

0.46

1.24

1.65

32

12

<20

10

20

160

250

60

70

<50

<50

386

50

50

<1,000

<1,000

<1,000

380

110

100

<1,000

<1,000

<1,000

<300

70

150

(7) Min.

(7) Max.

79

178

41

297

CHAPTER 10 j Leachate Quality

Table 10.2.4 Leachate analysis from Chinese landfills (Yue, 2016) Parameter

(1) Min.

(1) Max. Beijing

pH

(2)

(3)

(4)

Taiwan

Shanghai

Chongqing

(5) Min.

(5) Max.

Shenzhen

(6) Min.

(6) Max.

Changchun

6.8

8.5

7.9

7.4e7.8

8.21

6.2

8.0

7.2

7.5

BOD5

mg/L

90

18,200

164

1,619

1,011

1,000

36,000

465

820

COD

mg/L

2,070

33,300

2,483

8,592

4,867

3,000

60,000

2,027

3,212

NH4

mgN/L

1,100

2,930

534

2,701

400

1,500

1,013

1,044

SS

mg/L

324

4,720

584

737

100

6,000

TS

mg/L

DS

mg/L

10,200

24,000

9,463

mS/cm

18

29

24

Conductivity Volatile fatty acid

mg/L

156 9,619

672

Max., maximum; Min., minimum. (1) Li et al. (2008): Determined in six landfills of Beijing: Asuwei, Anding, Beishenshu, Liulitun, Gaoantun, and Jiaojiapo landfill. (2) Fan et al. (2006): The landfill has been operated for about 10 years with total capacity of 3.4 million m3. The waste consists of municipal solid waste (MSW) (20%e60%) and incinerator bottom ash (40%e80%). (3) Zou et al. (2011): Laogang landfill in Shanghai started operation in 1989, receiving about 5000 t of refuse per day. 61% of the refuse is construction wastes, 38% agricultural and domestic waste. (4) Liu (2007): Chang Sheng Qiao landfill in Chongqing has operated for 3.5 years, mainly receiving MSW. Because of the relative high annual average temperature, the biological stabilization rate of waste was high. (5) Jiang et al. (2002): Waste composition in Shenzhen: 60% food waste, 7% ceramics textiles, 10% paper, 13% plastics, 4% glass, 1% metal. (6) Guan (2005): The leachate sampled from Peijia landfill in Changchun was typical middle or old age leachate.

523

Box 10.2.1 Leachate quality from municipal solid waste landfills (German class II) in North Rhine Westphalia (NRW), Germany (values based on 55 class II landfills and 2200e5800 analysis; dependent on parameter) (Anonymus, 2010) Parameter

Median

75% Quantile

Maximum

SO4

mg/L

77

431

11,000

Cl

mg/L

1,153

1,998

10,668

Na

mg/L

912

1,509

10,100

K

mg/L

518

840

4,900

Ca

mg/L

127

271

5,490

Mg

mg/L

92

155

1,121

Conductivity

mS/m

1,091

3,845

42,500

COD

mg/L

1,300

3,809

49,250

TOC

mg/L

512

1,305

16,320

NH4-N

mg/L

615

1,149

8,160

AOX

mg/L

680

1,461

22,900

As

mg/L

20

111

18,000

Cd

mg/L

1.5

30

1,000

Cr

mg/L

80

322

4,120

Cu

mg/L

40

224

4,100

Hg

mg/L

0.5

1.9

40

Ni

mg/L

116

283

3,560

Pb

mg/L

13

246

5,240

Zn

mg/L

110

996

23,300

nitrogen compounds such as proteins are ammonium and organic acids, acetate as well as CO2 and H2 (Sleat et al., 1989). To migrate through the landfill, the ammonium has to dissolve in the leachate. Due to the organic content of the waste, the mobilized ammonia may partly adsorb to organic material. In addition, the solubility of some proteins is not high. Nitrogen compounds of several proteins may be bound in not easily degradable organics. All together and with the high potential of nitrogen in the waste and the limited transport may be the reason of the long-term release of N in the leachate. The concentrations of most organic trace compounds in leachate are mostly low, because the solubility of many trace compounds in leachate is low and their amount in the emplaced waste is comparably low too. At several old landfills, organic halogens are/have been measured in concentrations of some milligrams per liter. Organic halogens with a high vapor pressure leave the landfill with the

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

gas flow. But some halogens with a lower vapor pressure are in the liquid stage and may be dissolved in the leachate. Some inorganic compounds such as sulfate, iron, calcium etc., show the same tendency between acetic and methanogenic phase as COD and BOD5. Many organic substances and demolition material contain sulfate. The biological reduction of sulfate to sulfide has a lower energy gain for microorganisms than hydrolysis and acidogenesis but a higher energy gain than methanogenesis. For this reason, sulfate is reduced in nearly the same period of time as the biological phase changes. The sulfate content of demolition waste can be released over a long period of time, which may result in an increase of sulfide concentrations in leachate at old and closed landfills unless they are not precipitated as metal sulfides. The concentrations of calcium, iron, manganese etc., in the leachate are influenced by their content in the emplaced waste (e.g., demolition waste). In addition the pH influences the solubility conditions. The relatively low pH values during the acetic phase result in an increase of the abovementioned compound concentrations, and with the change to the methanogenic phase where relatively higher pH values can be measured, these concentrations decrease. The concentrations of heavy metals in leachate vary by species and with time as well as between different landfills and sometimes between different parts of the same landfill. The solubility of heavy metals depends mainly on the pH value and the redox potential. Metals may adsorb at organic waste compounds or may react with inorganic substances. A typical metal fixation is the formation of metal sulfides, which precipitate out due to their relatively low solubility in leachate. Under aerobic conditions, the sulfides may be oxidized to sulfates and metals may be solubilized again but will mainly coprecipitate with the newly formed hydroxides. Owing to high CO2 concentrations in the landfill, some heavy metals such as zinc and cadmium may also precipitate as carbonates. For these reasons the heavy metal content in the leachate is in general relatively low either under aerobic as well as anaerobic conditions. At most landfills, mainly the biodegradable organics, nonbiodegradable organics, and nitrogen compounds in the leachate are in focus. Nitrogen is measured as ammonium and organic nitrogen. The differentiation between biodegradable and nonbiodegradable organics is relatively difficult. As a representative parameter of these two groups of organics with different behavior, the COD is measured. In addition to the oxidation of the organics, the oxygen consumption of some inorganic compounds also is included in the COD. The parameter BOD5 describes the biodegradable fraction in 5 days (in some countries also BOD7 BSB 7). For this reason the not easily degradable organics in the leachate, e.g., from the methanogenic phase are not or only partly measured. A more meaningful value would be, e.g., a BOD20. The measurement of lower BOD values is often difficult. Low BOD values and high ammonium concentrations in the leachate often go together in methanogenic phase. The toxicity effect of ammonium on BOD measurement requires a high dilution of the leachate for the BOD analysis. With increasing sample dilution, the potential error of the BOD analysis may be higher. Of course also the potential nitrification in the BOD bottle has to be suppressed. The solution may be a laboratory batch treatment test using a fixed film reactor where the COD of the constant influent and the constant effluent are measured. In this case the difference of the COD concentrations would represent the biodegradable fraction.

CHAPTER 10 j Leachate Quality

525

LONG-TERM ASPECTS OF LEACHATE QUALITY At the end of landfill operation, leachate will still be produced unless the landfill surface is completely covered with an impermeable liner. As long as water infiltrates into the landfill, substances will be dissolved and transported out of the landfill. Especially for the parameters COD and nitrogen, it is necessary to estimate how long leachate treatment may be necessary. At landfills with a long-term leachate monitoring, leachate emissions can be estimated; but monitoring phases of 20e30 years are in most cases too short for long-term prediction because in many cases the leachate concentrations are still high because the landfill has not reached the final stable phase. Investigating pilot-scale MSW test cells it was possible to study the biological waste stabilization over 1e2 years; as a result, at the end of the experiments very low concentrations had been reached. Using pilot-scale data for long-term prediction of the leachate quality and quantity of full-scale landfills, it is necessary to find a parameter to adjust both timescales. Instead of a timescale the water infiltration to dry solids ratio (w/s ratio) shows a relative good agreement of the different data sets (Ehrig, 2001; Ehrig and Krümpelbeck, 2001; Krümpelbeck, 2000; Krümpelbeck and Ehrig, 2000). A typical description of such transport processes is an exponential equation (Fig. 10.2.3). Input data for the regression analysis are the leachate concentration at different times (Ct) and the time (t); the parameter k as a constant value includes the average annual infiltration/dry solids rate (mm/year/kg dry solids of a 1 m2 waste column). The result of the regression analysis is the concentration (C0) at the end of landfill operation (year ¼ 0) and a coefficient F. The parameter F describes the curve of this function and depends on the behavior of the specific pollution potential as there are solubility, transfer conditions etc., for each parameter. The coefficient differs between different parameters [average value: F (COD) ¼ 1; F (ammonium) ¼ 0.4]. Examples with different concentrations of COD and ammonium at the end of landfill operation and different w/s ratios are presented in Fig. 10.2.3. Considering the inhomogeneity of waste and landfills and the high fluctuation of leachate concentrations, the parameter F should be used for long-term prediction in a range of 25%. A similar approach has also been published by Wang et al. (2013). Considering the inhomogeneous nature of the waste in the landfills, a comparison between both estimations shows a relative good agreement. Other authors such as Weber (2002) used also an exponential model for the long-term leachate quality prediction but did not consider the effect of the same infiltration rate at landfills with different heights (e.g., 10 and 30 m). Fig. 10.2.4 shows the practical application of the equation Ct ¼ C0 * ek*t (“calculated” in Fig. 10.2.4). Average annual COD and ammonium concentrationsdstarting at the end of the operationdwere monitored in the leachate of 16 relative small German (state of Bavaria) landfills and are presented. An exponential regression was calculated to statistically describe the data set (see “Expon” in Fig. 10.2.4); in addition the exponential leachate prediction model was applied (Fig. 10.2.3). Fig. 10.2.4 shows on one side a good agreement of such estimations and on the other side their limits. Two years after the end of operation the COD concentrations drop sharply mainly due to the fact that after the end of operation most landfills in Germany had been covered with a

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

Figure 10.2.3 Calculated long-term chemical oxygen demand (COD) and ammonium leachate con-

centrations dependent on landfill height (m) and annual infiltration rate (mm) (two examples for each case: typical low and high concentrations at the end of operationdexplanation see Long-Term Aspects of Leachate Quality section) (Krümpelbeck, 2000; Ehrig, 2001).

top liner. The prediction of the leachate quality at this early stage using a statistical evaluation or the application of the exponential model would give very erratic results. Fig. 10.2.4 shows a very good agreement of calculated and measured COD values after these initial 2 years (relative stable conditions); this is not the case for the long-term behavior of ammonium. At several landfills the ammonium concentrations after some decades of operation or after the end of the operation may be relatively constant or may even increase for a certain period. Some thoughts about potential reasons for the high NH4 fluctuations are discussed above. An estimation of future tendencies would be more feasible after the end of the period of increasing or constant concentrations.

CHAPTER 10 j Leachate Quality

527

Ammonium (mgN/L)

COD (mg/L)

Time after end of operation (years)

Figure 10.2.4 Comparison of average annual COD and NH4-N concentrations in leachate of 16

Bavarian landfills after the end of landfill operationdsolid line: average values; dashed line: exponential regression using the measured values; dotted line: calculated with Ct ¼ C0 * ek*t (see Long-Term Aspects of Leachate Quality section). COD and ammonium data from Huber and Schatz (2007).

LEACHATE QUALITY FROM LANDFILLS FOR MECHANICALLY AND BIOLOGICALLY TREATED WASTE (MECHANICAL BIOLOGICAL TREATMENT LANDFILLS) As already mentioned the most critical parameters of leachate pollution are COD and ammonium; these parameters represent the high long-term emission potential. As a consequence high costs for leachate treatment over decades have to be encountered. Both parameters are predominantly influenced by the biodegradable organic waste fraction and the intensity of the biodegradation processes in the landfill. This situation and the uncontrolled production of the greenhouse gas methane during operation were the basis for the decision to ban biodegradable waste from landfilling. Several states in the EU allow only biologically or thermally treated MSW to be landfilled (see also Chapters 1.1, 1.2, 2.1, and 4.1). Several aerobic and anaerobic biological treatment processes and their combinations are used to biologically stabilize MSW to prescribed limit values. Owing to different mechanical and biological treatment processes and the degree of MSW stabilization, the leachate pollution can vary to a certain extent. There are several MBT landfills in operation but the period of operation is relatively short and most landfills are combinations of the closed raw waste and MBT landfills. For this reason the leachate are often mixtures of both types of landfills. Robinson et al. (2004) collected and analyzed leachate samples at different German landfills. The results are presented in Table 10.2.5. At these landfills more than

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

Table 10.2.5 Leachate analysis from different landfills containing more than 90% mechanical

biological treatment waste (Robinson et al., 2004) Parameter

(1)

(2)

(3)

(4)

(5)

pH

8.3

7.9

7.5

8.4

8.5

BOD5

mg/L

202

35

1-55

6

3

COD

mg/L

4670

1620

700-2500

869

1020

TOC

mg/L

1480

543

300e950

308

340

NH4

mgN/L

1130

197

0e27

34.2

1.8

NO3

mgN/L

<0.3

<0.3

15e66

6.3

4.7

NO2

mgN/L

<0.1

0.1

0.1e1.7

1.0

0.3

Phosphate

mgP/L

12.4

2.8

0.4

0.3

Alkalinity (as CaCO3)

mg/L

6120

2010

1670

895

Cl

mg/L

2270

2290

901

1090

SO4

mg/L

117

449

414

878

Na

mg/L

1520

1250

622

789

K

mg/L

728

777

393

387

Mg

mg/L

88

104

64

67

Ca

mg/L

176

329

232

255

Mn

mg/L

1.38

2.94

1.45

1.61

Fe

mg/L

19.5

13.9

2.59

1.31

Zn

mg/L

0.67

0.2

0.19

0.53

As

mg/L

<1

6

<1

4

Cd

mg/L

6

3

2

3

Cr

mg/L

494

96

42

90

Cu

mg/L

180

150

80

250 (Continued)

CHAPTER 10 j Leachate Quality

529

Table 10.2.5 Leachate analysis from different landfills containing more than 90% mechanical biological treatment waste (Robinson et al., 2004)dcont'd Parameter

(1)

(2)

(3)

(4)

(5)

Hg

mg/L

<0.1

<0.1

<0.1

<0.1

Ni

mg/L

260

82

51

97

Pb

mg/L

130

40

10

20

AOX

mg/L

790

1500

180

370

100e900

Pretreatment: (1) Passive windrows. (2) Turned windrows. (3) Turned windrows, active aeration. (4) Container. (5) Container þ windrows. Landfill age: (1) 9 years. (2)e(5) 2e5 years.

90% of the landfilled waste was MBT waste, which had been aerobically treated. The intensity of the aerobic treatment was different and increases in principle from column 1 to 5 (Table 10.2.5). With the exception of one landfill, the landfill age was relatively short amounting to 2e5 years. As a consequence the presented data may give only an indication on leachate quality from MBT landfills. But also due to this limitation it becomes obvious that with increasing biological stabilization the concentrations of nitrogen (ammonium and nitrate) as well as BOD5 and COD decrease. Comparing leachate quality data from old/very old MSW landfills (decades or longer) particularly with the NH4 concentrations from MBT landfill shows that the NH4 concentrations are significantly lower in the leachate from MBT landfills; in addition, the BOD5 values are also very strongly reduced. But the remaining nonbiodegradable organics measured as COD and TOC are still relatively high in the leachate from MBT landfills. Considering that these MBT waste landfills are only some years old with a relatively small waste volume, the COD concentrations are in the same range as in leachate from smaller old MSW landfills. This implies that leachate from MBT landfills also has to be collected and treated for a long nonpredictable time period. The higher density, different structure and composition of the MBT waste results in a higher water adsorption capacity (see Chapters 4.1, 14.1). Therefore at least over a period of several years low leachate production rates can be expected. The data in Table 10.2.5 do not show significant differences in the inorganic leachate quality from MSW and MBT landfills. During mechanical treatment the high calorific value (coarse fraction) is separated from the fraction <10 mm or smaller. The fine fraction consists of biologically degradable waste and inorganic fines as dust etc. Owing to the comparatively higher concentrations of the

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

inorganic compounds in the MBT waste higher inorganic leachate concentrations also could be expected; but the presented data here do not give any indications in this regard. The data in Table 10.2.5 include only MBT waste landfills with aerobic treatment. Theoretically there might be some differences in leachate quality whether the treatment happened under aerobic or anaerobic conditions. Dependent on the time of aerobic posttreatment after the anaerobic treatment step elevated COD and ammonium leachate concentrations may occur especially when the postcomposting step is relatively short. In case of an anaerobic wet treatment process, some compounds may be washed out. But this has to be proven by data from the field. INFLUENCE OF LANDFILL IN SITU AERATION ON LEACHATE QUALITY Many laboratory-scale investigations show that organic leachate concentrations decrease relatively fast with an increasing degree of biological stabilization. While the BOD5 values decrease to very low values (<25 mg/L), the COD will still remain elevated (>100 mg/L). Leachate quality data from laboratoryscale plants also show a decrease of the NH4-N concentrations to low values, which may be <10 mg/L. But these tests were operated at constant temperatures of 35 C, i.e., also for this reason the leachate quality data cannot be transferred to full-scale landfills, where during aeration temperatures up to 50 C will be reached in most cases. In addition a higher water flux was installed. But these data show the potential that leachate quality may be reached due to landfill aeration. The elevated temperatures may result in faster carbon degradation; organic nitrogen ammonification may be enhanced that may result in an elevated pH. At temperatures >45 C nitrification will be inhibited, and due to the higher pH, NH3 may be produced; in how far this may cause (local) inhibition processes in the landfill body is difficult to predict. After the temperatures decrease <45 C, the reduction of the ammonium can be expected. At the bottom of landfills, often high water content/levels can be found. Owing to the fact that these areas are difficult to aerate, this situation may influence the leachate quality negatively and may result in an increase of the relative low concentrationsdreached in the aerated partdof some leachate parameter (see also Ritzkowski, 2011). LEACHATE QUALITY FROM BOTTOM ASH LANDFILLS In many countries, incineration of MSW is a common practice. In Germany it is regarded as a treatment step before landfilling. Bottom ash is in Germany, Denmark, and some other countries partly used and partly landfilled either in Class I (inert waste landfills) or Class II (MSW landfill (Germany for treated waste)); fly ash is landfilled in hazardous waste landfills (German class III landfill). In many cases, ferrous material and sometimes aluminum are separated from bottom ash prior to landfilling. In other countries, bottom and fly ashes are still codisposed in mono landfills. During incineration, almost all organics are oxidized. Compared with bottom ash the leachable amount of heavy metals from fly ash is considerably higher. The high temperatures during incineration destroy chemical bonds of several inorganic components. Exothermic reactions of new chemical bonds can be observed in fresh bottom ash over a period of about 3 months. If fresh bottom ash is landfilled, increasing temperatures can be

CHAPTER 10 j Leachate Quality

531

measured. During this phase the leachability of several inorganic compounds is relatively high (see also Chapter 20). The quality of incineration and the kind of treatment of the residuals may have an effect on the quality of the leachate, but it is very difficult to verify this. Tables 10.2.6 and 10.2.7 show the variation of data from different bottom ash landfills. Some of these landfills contain fly ash or noncombustible inorganic waste also. Leachate of bottom ash landfills is often highly polluted with chloride, sulfate, sodium, and potassium. The heavy metal concentrations are in the same range as in leachate from MSW landfills. At several landfills, the COD respectively TOC concentrations are relatively high, but the amount of biodegradable compounds is very low. The data presented by Johnson et al., 1999 (Table 10.2.6) show that by separating out bulky materials from the bottom ash COD and TOC values in the leachate may be reduced. In this case also ferrous material had been separated and the bottom ash was stored over periods of 3e6 month (until the exothermic reactions have come to an end and the temperature drops down to nearly ambient values). The effect on chloride, sulfate etc., in leachate seems to be relatively limited. Notable are the low heavy metal concentrations in the leachate of this landfill.

Table 10.2.6 Leachate quality of bottom ash landfills Parameter

pH

Aver.(1)

Min.(1)

Max.(1)

Min.(2)

Max.(2)

Aver.(3)

Min.(3)

Max.(3)

Aver.(4)

7.7

7.4

8

7.2

10.5

10.3

8.7

11.3

10.5

10.3

4.4

44.8

8

TOC

mg/L

126

6

400

COD

mg/L

42

22

62

75

1100

BOD5

mg/L

859

17

1700

<2

130

Na

mg/L

5050

300

9800

1600

7300

1024

225

2098

1746

K

mg/L

1650

100

3200

600

4300

460

181

944

737

Ca

mg/L

521

401

641

32

1000

329

124

644

536

Mg

mg/L

0.09

66

15.2

2.5

26.2

18.2

Cl

mg/L

2784

290

9300

160

14,300

1672

391

3657

3042

mgN/L

72

0.3

190

1

87

NH4

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

Table 10.2.6

Leachate quality of bottom ash landfillsdcont'd

Parameter

Aver.(1)

Min.(1)

Max.(1)

Min.(2)

Max.(2)

Aver.(3)

Min.(3)

Max.(3)

Aver.(4)

640

7200

1190

529

2141

1786

<0.01

18.5 0.0047

0.0015

0.018

NO3

mgN/L

58

25

90

SO4

mg/L

1562

985

2900

AOX

mg/L

1.8

0.01

4.54

Fe

mg/L

0.75

0.5

1

Mn

mg/L

0.055

0.05

0.06

Pb

mg/L

19

0.3

75

<0.5

40

2.7

0.5

7

Cu

mg/L

123

40

220

<0.5

210

101

6.9

479

33

Zn

mg/L

66

2

250

<10

910

5.7

2.9

10.5

5.6

Cd

mg/L

2

0.9

5

<0.1

2

1.3

0.38

5.4

1.7

Ni

mg/L

20

1

70

<2

42

Cr

mg/L

17

13

25

<1

80

10.9

8.5

16.1

11.6

Hg

mg/L

0.1

<0.05

3

As

mg/L

5

25

Mo

mg/L

522

145

1315

602

W

mg/L

113

34

234

63

Sb

mg/L

32

11

57

19

V

mg/L

22

11

48

35

Aver., average; Min., minimum; Max., maximum. (1) ATV-Fachausschuss 3.6 (1997). Unknown number of bottom ash landfills (Germany). (2) Hjelmar et al. (1995). Bottom ash landfill (Denmark). (3) Johnson et al. (1999). Bottom ash landfill (Switzerland). The ash was stored in piles over 3e6 month before dumping. Ferrous material and bulky unburned organic material were removed. Landfill age during sampling time: 3e4 years. (4) average values during dry weather periods.

CHAPTER 10 j Leachate Quality

533

Table 10.2.7 Leachate quality of bottom ash landfills Average(1) 1973e75

Parameter

pH

10.1e9.1

Average(1) 2001e03

Average(2) 1990e93

Average(2) 1997e99

Average(2) 2004e06

Average(3) and Standard Deviation

8.9e8.6

10

9.3

9

7.8  0.5

TOC

mg/L

387

47

30

COD

mg/L

1540

140

103

K

mg/L

3600

210

1142

492

545

Ca

mg/L

610

14

252

212

195

Mg

mg/L

0.27

17

6

5

7

Cl

mg/L

9200

340

3476

4033

4150

NH4

mgN/L

34

0.35

160

48

18

NO3

mgN/L

0.5

5

3

SO4

mg/L

1040

3000

4300

AOX

mg/L

3.5

0.56

0.85

Fe

mg/L

0.04

0.1

0.14

1.3  1.4

Pb

mg/L

<20

30

70

10

<10

19  31

Cu

mg/L

<30

86

40

20

<5

290  270

Zn

mg/L

33

940

70

20

20

58  36

Cd

mg/L

<5

1.2

10

5

<1

1.2  2.8

Ni

mg/L

50

72

200

17

8

64  32

Cr

mg/L

33

<2

10

10

7

500  310

Hg

mg/L

<1

<0.005

3

10

As

mg/L

14

6.2

10

20

3100

1500

2400  1100

1900  380

554  193

0.26  0.25 50

(1) Hjelmar and Hansen (2005). Bottom ash (85%) and fly ash landfill (15%) (Denmark). Landfill operation 1973e1976. (2) Morf and Kuhn, 2010. Bottom ash landfill (Switzerland). Dumping on this landfill part was mostly finished in 1996. (3) Øygard et al. (2005). Landfill for bottom ash, construction, and inorganic industrial waste (Norway). Landfill operation starts 1999. Analysis of the landfill leachate throughout 2002 and 2003.

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

LEACHATE FROM OTHER LANDFILL TYPES As an example of leachate quality, data (median, 75% quantile, and maximum) from different types of landfills monitored in North Rhein Westphalia (state in Germany) over certain periods of time are presented in Boxes 10.2.2e10.2.4.

Box 10.2.2 Leachate quality from inert waste landfills (German class I) in North Rhine Westphalia (NRW), Germany (values based on 25 class I landfills and 334e1031; analysis dependent on parameter) (Anonymus, 2010) Parameter

Median

75% Quantile

Maximum

SO4

mg/L

398

1,761

6,600

Cl

mg/L

230

529

2,712

Na

mg/L

330

1,043

3,350

K

mg/L

106

305

1,540

Ca

mg/L

140

308

1,230

Mg

mg/L

18

151

1,100

Conductivity

mS/m

292

774

7,120

COD

mg/L

59

572

6,500

TOC

mg/L

35

215

2,335

NH4-N

mg/L

10

195

1,294

AOX

mg/L

50

503

8,400

As

mg/L

5

45

590

Cd

mg/L

0.7

4

34

Cr

mg/L

10

29

270

Cu

mg/L

10

68

1,370

Hg

mg/L

0.2

0.9

5.5

Ni

mg/L

20

74

1,110

Pb

mg/L

5

97

2,800

Zn

mg/L

49

964

18,900

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Box 10.2.3 Leachate quality from hazardous waste landfills (German class III) in North Rhine Westphalia (NRW), Germany (values based on 14 class III landfills and 640e2250 analysis; dependent on parameter) (Anonymous, 2010) Parameter

Median

75% Quantile

Maximum

SO4

mg/L

1,083

4,577

44,800

Cl

mg/L

6,200

27,485

189,600

Na

mg/L

2,302

19,906

89,500

K

mg/L

500

8,240

43,000

Ca

mg/L

304

610

10,200

Mg

mg/L

123

220

1,700

Conductivity

mS/m

2,100

7,680

96,300

COD

mg/L

658

3,781

35,000

TOC

mg/L

102

1,192

8,700

NH4-N

mg/L

258

1,497

9,845

AOX

mg/L

1,000

5,412

34,000

As

mg/L

150

4,376

38,000

Cd

mg/L

2

4,386

87,000

Cr

mg/L

56

2,954

50,000

Cu

mg/L

40

1,960

21,000

Hg

mg/L

0.4

9

274

Ni

mg/L

320

2,590

55,000

Pb

mg/L

8

481

8,200

Zn

mg/L

50

18,498

590,000

SOLID WASTE LANDFILLING j Concepts, Processes, Technologies j R. Cossu, R. Stegmann

Box 10.2.4 Leachate quality from sewage sludge landfills (mostly anaerobic digested sewage sludge) in North Rhine Westphalia (NRW), Germany (values based on five sewage sludge landfills and 121e222 analysis; dependent on parameter) (Anonymous, 2010) Parameter

Median

75% Quantile

Maximum

SO4

mg/L

210

569

2200

Cl

mg/L

190

479

1900

Na

mg/L

67

171

650

K

mg/L

23

57

300

Ca

mg/L

165

364

786

Mg

mg/L

30

86

250

Conductivity

mS/m

256

1022

4980

COD

mg/L

96

831

1850

TOC

mg/L

24

152

1000

NH4-N

mg/L

10

153

820

AOX

mg/L

50

114

530

As

mg/L

5

28

109

Cd

mg/L

0.1

2

11

Cr

mg/L

10

25

150

Cu

mg/L

12

495

7600

Hg

mg/L

0.1

0.2

1

Ni

mg/L

16

586

4700

Pb

mg/L

10

32

200

Zn

mg/L

20

162

1300

CONCLUSIONS Leachate production rates and quality have to be known to design a leachate treatment plant. Since the plant has to be built in the course of the landfill construction, leachate loads have to be predicted. The data presented here are examples mainly from landfills in Europe, but the quality characteristics are similar all over the world. On the other hand it can be seen from the wide range of concentrations that each landfill is different and leachate quality has to be monitored from the beginning of landfill operation. Important are the waste input quality and the way the landfill is operated. Tendencies are summarized in Table 10.2.8. Operating the landfill in a way that leachate pollution is reduced will result in a significant cost savings; in addition, with an improved way of operation, additional positive effects can be achieved (see Chapters 1.1, 2.1, 2.2, 3, 14.1, 16.2).

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Table 10.2.8 Factors of influence on tendencies in leachate concentrations with time based on average

conditions (different factors can be combined e.g. MBT waste with high or low precipitation input and/ or high or low compaction, which will result in higher or lower decrease of organic concentrations).

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Hjelmar, O., Johannessen, L.M., Knox, K., Ehrig, H.J., Flyvbjerg, J., Winther, P., Christensen, T.H., 1995. Management and Composition of Leachate from Landfills. EU commission, Management of waste, DGXI, C7 unpublished report. Hjelmar, O., Hansen, J.B., 2005. Sustainable landfill: the role of final storage quality. In: Tenth International Waste Management and Landfill Symposium, Sardinia, 3-7, October 2005. Huber, W., Schatz, S., 2007. Entwicklung der Sickerwasserzusammensetzung in der Nachsorgephase e Auswertung von bayerischen Altdeponien (Development of leachate composition in the aftercare phase e analysis at old Bavarian landfills). In: 18. Nürnberger Deponieseminar 2007. Jiang, H.T., Zhou, G.M., Gao, T.Y., 2002. The characteristics of MSW landfill leachate. Environmental Protection Science 3, 11e13 (in Chinese). Johnson, C.A., Kaeppeli, M., Brandenberger, S., Ulrich, A., Baumann, W., 1999. Hydrological and geochemical factors affecting leachate composition in MSW incinerator bottom ash. Part II: the geochemistry of leachate from landfill Lostorf, Switzerland. Journal Contaminant Hydrology 40, 239. Krümpelbeck, I., 2000. Veröffentlichungen des Lehrstuhls für Abfall- und Siedlungswasserwirtschaft der Bergischen Universität Wuppertal. Untersuchungen zum langfristigen Verhalten von Siedlungsabfalldeponien: Investigations on long term behaviour of landfills, vol. 3. Krümpelbeck, I., Ehrig, H.J., 2000. Emissionsverhalten von Altdeponien (Emission behaviour of old landfills). Deponietechnik, Hamburg. Li, C.P., Li, G.X., Luo, Y.M., Li, Y.F., 2008. Fuzzy mathematics-based ground water quality evaluation of six MSW landfills in Beijing. Environmental Science 10, 2729e2735 (in Chinese). Liu, G.Q., 2007. Study on Characteristic and Removal Efficiency of DOM in Leachate (Master thesis). Chongqing University (in Chinese). Loizidou, M., Vithoulkas, N., Kapitanios, E., 1992. Physical chemical treatment of leachate from landfill. Journal Environmental Science and Health A 27 (4), 1059. Lopez, A., Pagano, M., Volpe, A., di Pinto, A.C., 2004. Fenton’s pretreatment of mature landfill leachate. Chemosphere 54, 1005. Morf, L.S., Kuhn, E., 2010. Qualität von Sickerwasser aus Zürcher Schlackenkompatimenten (Leachate quality from landfill parts e Zürich). AWEL Amt für Abfall, Wasser, Energie und Luft, Kanton Zürich. Øygard, J.D., Gjengedal, E., Måge, A., 2005. Mass-balance estimation of heavy metals and selected anions at a landfill receiving MSWI bottom ash and mixed construction wastes. Journal of Hazardous Materials a 123, 70. Ritzkowski, M., 2011. How does landfill aeration impact on leachate composition? In: Cossu, Diaz, Stegmann (Eds.), Sardinia 2011: Proceedings Thirteenth International Waste Management and Landfill Symposium. Robinson, H., 1995. A Review of the Composition of Leachates from Domestic Wastes in Landfill Sites. Environment Agency, UK. CWM 072/95. Robinson, H.D., Knox, K., Bone, B.D., 2004. Improved Definition of Leachate Source Term from Landfills e Phase 1: Review of Data from European Landfills, Science Report P1e494/SR1. Environment Agency, UK, ISBN 1844323269. Robinson, H., June 2007. The Composition of Leachates from Very Large Landfills. An International Review, vol. 81. CWRM, p. 19. Robinson, H.D., Carville, M., June 2007. The design, commissioning and operation of leachate treatment plants at large landfills in tropical regions. In: Paper Presented to “Waste a Global Resource”, the 2007 Annual Exhibition and Conference of the UK Chartered Institution of Wastes Management. Torbay, UK. Salem, Z., Hamouri, K., Djemnaa, R., Allia, K., 2008. Evaluation of landfill leachate pollution and treatment. Desalination 220, 108. Sleat, R., Harries, C., Viney, I., Rees, J.F., 1989. Activities and distribution of key microbial groups in landfill. In: Christensen, Cossu, Stegmann (Eds.), Sanitary Landfilling. Academic Press. Tatsi, A.A., Zouboulis, A.I., Matis, K.A., Samaras, P., 2003. Coagulation-flocculation pretreatment of sanitary landfill leachate. Chemosphere 53, 737. Wang, Y., Pelkonen, M., Kaila, J., 2013. Optimization of landfill leachate management in the aftercare period. Waste Management & Research 21, 789. Weber, B., 2002. Wie lange muss die Sickerwasserbehandlung dauern? (How long must be leachate treated?). In: Tagung Deponienachsorge, 10./11.9.2002, Hamburg. Yue, D., 2016. Personal Communication. Zou, L.S., Tang, J., Ye, K.Z., 2011. Choice of landfill leachate treatment in Shanghai old port. Journal of Yunnan University of Nationalities (Natural Sciences Edition) 20, 23e26 (in Chinese).

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