- Email: [email protected]
Sustainable disposal of excess sludge: Incineration without anaerobic digestion
Journal Pre-proof Sustainable disposal of excess sludge: Incineration without anaerobic digestion Xiaodi Hao, Qi Chen, Mark C.M. van Loosdrecht, Ji Li, Han Jiang PII:
S0043-1354(19)31072-3
DOI:
https://doi.org/10.1016/j.watres.2019.115298
Reference:
WR 115298
To appear in:
Water Research
Received Date: 25 July 2019 Revised Date:
4 November 2019
Accepted Date: 8 November 2019
Please cite this article as: Hao, X., Chen, Q., van Loosdrecht, M.C.M., Li, J., Jiang, H., Sustainable disposal of excess sludge: Incineration without anaerobic digestion, Water Research (2019), doi: https:// doi.org/10.1016/j.watres.2019.115298. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.
Graphical abstract
324 86/398
Legend:
Energy deficit (kW·h /t DS) Investment/Operational costs (×104 US$/t DS)/(US$/t DS)
Anaerobic digestion 313 90/445
Anaerobic digestion
Thermal hydrolysis 109 55/392 Excess sludge
Sustainability！
Incineration
Sustainable disposal of excess sludge: incineration without
1
anaerobic digestion
2
Xiaodi Hao a, *, Qi Chen a, Mark C. M. van Loosdrecht a, b, Ji Li a, Han Jiang a
3
4
a
5
Stormwater System and Water Environment, Beijing University of Civil Engineering & Architecture,
6
Beijing 100044, P. R. China
7
b
8
Netherlands
9
____________________________
10
Sino-Dutch R&D Centre for Future Wastewater Treatment Technologies/Key Laboratory of Urban
Department of Biotechnology, Delft University of Technology, van der Maasweg 9, 2629 HZ Delft, the
*
Corresponding author: Tel: +86 131 6134 7675; Fax: +86 10 6832 2123; E-mail: [email protected]
11
12
ABSTRACT:
13
Handling excess sludge produced by wastewater treatment is a common problem worldwide. Due to limited
14
space available in landfills, as well as difficulties involved in using excess sludge in agriculture, there is a
15
need for alternative disposal methods. Although anaerobic digestion (AD) is widely used in processing
16
sludge, only partial energy recovery from methane and sludge volume reduction can be achieved, resulting
17
in a substantial amount of sludge remaining, which needs to be disposed of. Direct incineration after sludge
18
drying is one possible option, a practice that is already in place in some cities in China. A comparison
19
between direct incineration and conventional AD (with or without pretreatment by thermal hydrolysis) has
20
to be made with respect to the energy balance and investment & operational (I & O) costs. This comparison
21
reveals direct incineration to have the lowest energy deficit and I & O costs. Therefore, it is expected that
22
direct incineration without AD will become the preferred sustainable approach to handling sludge.
23
Keywords: Sludge disposal; direct incineration; anaerobic digestion; sludge drying; energy deficit;
24
investment & operational costs
25
26
1. Introduction
27
Nowadays, the main methods of treating wastewater involve activated sludge, which makes excess sludge
28
an unavoidable problem. Landfilling and agricultural & horticultural use are the most convenient and
29
economical approaches to handling sludge (Suh et al., 2002). However, lack of landfilling spaces is a
30
prevailing problem for big cities, both worldwide and in China. In addition, agricultural applications are
31
limited by the low fertilizer value of sludge and the slow release of its nutrients, which helps farmers little
32
in raising grain production. Chinese farmers would rather use chemical fertilizers for crop cultivation, and
33
even tend to give up their conventionally “ecological” habits on applying manure as fertilizer (Hao et al.,
34
2019c).
35
Under these circumstances, China has to search for alternative methods of handling excess sludge. In
36
Europe and several other areas, sludge thickening followed by anaerobic digestion (AD) is commonly
37
applied to reduce the excess sludge amounts (Khalili et al., 2017; Thomsen et al., 2017). However, AD does
38
not remove all sludge organics, with as much as 50%-70% of the organics remaining (Hao et al., 2015;
39
Fernández-Arévalo et al., 2017). For this reason, other disposal methods, such as incineration, are still
40
necessary (Yoshida et al., 2018).
41
AD, followed by incineration, has a long history in Europe and in other countries (Abuşoğlu et al.,
42
2017), but direct incineration after dewatering and drying (reducing the moisture content of sludge to
43
40%-70%) (Fytili et al., 2008; Abuşoğlu et al., 2017) is also currently practiced in some Chinese cities.
44
Which way is better, AD plus incineration or direct incineration? This question calls for a comparative
45
study concerning the energy balance and investment & operational (I & O) costs.
46
Traditionally, producing biogas is seen as a sustainable approach to energy-efficient wastewater
47
treatment. However, in these cases the boundaries for evaluation are often within the wastewater treatment
48
plant itself. If direct incineration is, overall, more effective than AD plus incineration, it could conceivably
49
become a sustainable approach to handling sludge, and China would be able to leap forward in the
50
development on its sludge disposal. It was that idea that prompted this study.
51
2. Direct incineration
52
The proposed process of direct incineration consists of three steps: mechanical dewatering, thermal drying,
53
and incineration. Dewatering is applied to reduce the moisture content of sludge, from about 99% to 80%,
54
which creates a considerable reduction in sludge volume, by about 95%. However, this sludge is still far
55
from suitable for self-supporting incineration (Abuşoğlu et al., 2017). With thermal drying, the moisture
56
content of dewatered sludge can be reduced from 80% to 40%-70% (dependent on the organic content of
57
the sludge), which allows for self-supporting incineration (Fytili et al., 2008; Abuşoğlu et al., 2017).
58
Following this, organics contained in dried sludge can be incinerated directly, and fully oxidized into CO2
59
in incinerators at a high temperature of 800-900 °C or higher, with thermal energy generation, and with
60
phosphorus and other minerals remaining in the ashes (Abuşoğlu et al., 2017). The thermal energy can be
61
used for generating electricity, for drying sludge on site, or for municipal heating/cooling systems. Finally,
62
phosphorus and other minerals or metals can be recovered from the ashes (Murakami et al., 2009).
63
2.1. Energy balance
64
2.1.1. Mechanical dewatering
65
There are already many types of mechanical dewatering processes, including press and vacuum filtration,
66
centrifugation, etc. (Colin et al., 1995). Mechanical dewatering energy consumption depends mainly on the
67
structure of sludge flocs and the complex shear force between sludge and water, as well as different types
68
of dewatering equipment and chemicals dosage, and is not dependent on the different origins of sludge
69
(Novak, 2006). The energy consumption of different processes varies from 15-179 kW·h/t DS (Wakeman,
70
2007). The belt press filter, which has an energy consumption of about 60 kW·h/t DS (the moisture content
71
of sludge is reduced from 99% to 80%), is extensively applied in China (Wakeman, 2007).
72
2.1.2. Thermal drying
73
There are two types of thermal drying processes: i) the full drying process, whereby moisture content can
74
be reduced to ≤15%, and which is generally conducted by rotating-drum dryers; ii) the partial drying
75
process where moisture content is usually reduced to 35%-50%, often operated by rotary-plate dryers
76
(Gross, 2010; Lowe, 2010). Dried sludge with ≤70% moisture content can be self-incinerated without the
77
input of any auxiliary fuels (Fytili et al., 2008; Abuşoğlu et al., 2017), and so a partial drying process is
78
sufficient in this case. The total energy consumption (Et) of the thermal drying process consists of an
79
increased solids’ temperature (Es) and water evaporation (Ew), as expressed in Eqs. 1 and 2 (Chen et al.,
80
2017). Es =(T2 -T1 )∙Cs ∙Ms ∙100
（1）
81
Where, Es is the energy consumption of the temperature increase of the solids in the dewatered sludge, kJ/t
82
DS; Cs is the specific heat capacity of dewatered sludge, 3.62 kJ/kg· oC; T1 and T2 are the temperatures of
83
the input (20 oC) and output (100 oC) sludge, respectively; and Ms is the weight of dry matter of sludge
84
(DS), which is a constant value: 10 kg/t wet sludge (99% moisture content). Ew =Cw ∙ Mw =[
Ms w1 ∙ T2 -T1 ∙100+Qg ∙Mw 1-w1
Ms Ms ]∙100 1-w1 1-w2
（2）
（3）
85
Where, Ew is the energy consumption of water evaporation, kJ/t DS, including heating water and
86
evaporating water; Cw is the specific heat capacity of water, 4.2 kJ/kg· oC; w1 and w2 are the moisture
87
content of the input (80%) and output (40%-70%) sludge, respectively; Qg is the potential vaporization heat
88
value of water, 2,260 kJ/kg; and Mw is the weight of evaporated water, kg, which can be calculated by Eq. 3
89
(Chen et al., 2017).
90 91
The most important parameter is the targeted moisture content (w2) of dried sludge after thermal drying, which can be calculated by Eq. 4 (Murakami et al., 2009; Mills et al., 2014; Samolada et al., 2014). QsL=QsH ∙ 1-w2 -Qg ∙(w2 +9wH )
（4）
92
Where, QsL is the low calorific value of sludge, set at 3.36 GJ/t DS for self-supporting incineration; QsH is
93
the high calorific value of sludge, which is the maximal potential energy contained in dried sludge, and
94
which can be calculated by Eq. 6 (Cai et al., 2010), in GJ/t DS; and wH is the mass fraction of the organic
95
bound hydrogen element in the dry matter (about 2% of the dry matter) (Murakami et al., 2009; Samolada
96
et al., 2014).
97
Because of heat loss in sludge dryers, the actual energy consumption of the drying processes is
98
obviously higher than the theoretical energy consumption. The actual energy consumption of sludge drying
99
can be calculated by Eq. 5. The heat loss efficiency (η) of different types of thermal dryers ranges from 10%
100
to 20% (Li et al., 2014) and here the highest value (20%) is used to calculate the energy consumption. Et '=Et ∙ 1+η
（5）
101
2.1.3. Incineration
102
The theoretical thermal energy released from incineration (Q) is equivalent to the high calorific value of the
103
dried sludge (QsH), which can be calculated by Eq. 6 (Cai et al., 2010). Q=2.5×105×(
100·Pv 100-Pc -5)× 100-Pc 100
（6）
104
Where, Q is the theoretical thermal energy released from incineration; Pv is the percentage of the organic
105
content; Pc is the percentage of inorganic coagulants added to the sludge during mechanical dewatering
106
(thus, 0% when organic polymer coagulants are added).
107
The organic content of sludge (Pv) in China is normally within the range of 30%-65%. In this study, Pv
108
was set at 53% (the typical mean value for China) (Cai et al., 2010). Therefore, the theoretical thermal
109
energy released from incineration is calculated at 11.9 GJ/t DS, equivalent to the mean value of thermal
110
energy released across China (11.85 GJ/t DS) (Cai et al., 2010; Winkler et al., 2013), and the moisture
111
content (w2) of dried sludge is calculated at 57.7% by Eq. 4. Based on the above equations, the actual
112
energy consumption of the thermal drying process is 9.1 GJ/t DS, approaching to data used in process
113
designs of 11.7 GJ/t DS (Chen et al., 2017). When converted to its electrical equivalent, the actual energy
114
consumption is 2,529 kW·h/t DS (1 kW·h=3,600 kJ).
115
Of course, the actual energy recovered from incineration would be lower than the value calculated
116
above, due to factors as the low efficiency of fluidization, the incomplete combustion of solids and gases,
117
heat loss from incinerators, low-temperature off-gases (all resulting in inefficiencies in the electricity
118
generation), etc. (Murakami et al., 2009; Mills et al., 2014). According to engineering data, the total
119
thermal energy loss is about 7% of the theoretical energy released from incineration (11.9 GJ/t DS), which
120
is 11.9 GJ/t DS×7%=0.8 GJ/t DS (Li et al., 2014). Thus, the actual energy recovered from incineration is
121
around 11.1 GJ/t DS, which consists of two parts: i) 84% high-temperature thermal energy (9.32 GJ/ t DS)
122
(Qiu et al., 2007; Liao et al., 2012) which can be directly converted into electricity via Combined Heat and
123
Power (CHP) (80%, 7.46 GJ/ t DS), resulting in a left-over thermal energy (20%, 1.86 GJ/t DS); and ii) 16%
124
for low-temperature waste thermal energy (1.78 GJ/t DS) (Qiu et al., 2007; Liao et al., 2012). Both low
125
thermal energy sources (3.64 GJ/t DS)) could be counted as “electrical equivalent” (based on Table S1),
126
with a thermal utilization efficiency of up to 40%. Based on the practical calculations shown above, the
127
total electrical energy (2,072 kW·h/t DS) plus electrical equivalent (408 kW·h/t DS) generated by sludge
128
incineration is about 2,480 kW·h/t DS, or approximately 80% of the total energy.
129
2.1.4. Energy balance
130
Based on the energy estimates calculated above, the energy balance of the direct incineration process is
131
shown in Fig. 1a, revealing an energy deficit of 109 kW·h/t DS.
132 Fig. 1. Energy balances of the three compared sludge disposal processes.
133 134
2.2. Investment & operational costs
135
I & O costs depend on the capacity of wastewater treatment plants (WWTPs) and applied equipment. In
136
this study, a WWTP with a treatment capacity of 500,000 m3/d was used to estimate the I & O costs
137
according to practical data from wastewater treatment processes in China (Table S2). The plant was able to
138
produce 8,000 t wet sludge/d (99% moisture content). The investment costs consisted of infrastructure and
139
equipment. The operational costs consisted of electricity, water, chemicals, transportation, wages and
140
welfare, depreciation of fixed assets, and overhaul & maintenance costs. The detailed calculations on the
141
economic aspects are listed in Table S3 (Hong et al., 2009; Piao et al., 2016). The I & O costs of the direct
142
incineration process were calculated according to the Chinese market price levels, at 55 ×104 US$/t DS and
143
392 US$/t DS, respectively, which falls in the range of the practical data from several engineering cases in
144
China.
145
2.3. Environmental impacts
146
Nitrogen oxides (NOx), dioxins and even heavy metals can all be released from sludge during incineration.
147
At temperatures above 850 °C, however, the average released concentration (471.6 mg/m3) of NOx is lower
148
than the emission standard in China (500 mg/m3) (Lin et al., 2015). Even referring to the stringent EU
149
emission standard of NOx (400 mg/m3), NOx emission can easily be controlled by several different
150
technologies (Svoboda et al., 2006). Furthermore, the decomposition rate of dioxins is much greater than its
151
formation rate above 850 °C (Thomas et al., 2008; Zhan et al., 2016), and thus the net concentration of
152
dioxins produced can be lower than the EU emission standard (0.1 ng-TEQ/m3 off-gases) without any
153
treatment (Zhou et al., 2015). In addition, heavy metals (including mercury) are also not a serious concern
154
(Xu et al., 2008), and all of these can effectively be controlled above 850 °C (Yao et al., 2005). In short, the
155
3T+E (Temperature, Time, Turbulence and Excess) control strategy on incineration can greatly help to
156
reduce off-gases pollutants concentrations of incineration (Zhou et al., 2015). Research reports from the
157
United States of America, Germany, Great Britain and Spain have all reached similar conclusions: there is
158
no evidence that sludge incineration and even waste incineration will endanger the environment or human
159
health (Federation, 2009).
160
3. Anaerobic digestion followed by incineration
161
3.1. Conventional anaerobic digestion plus incineration
162
AD plus incineration generally includes five steps: gravity thickening, AD, dewatering, thermal drying and
163
incineration. In gravity thickening, the moisture content of sludge is reduced from 99% to 97%. During AD,
164
the fraction of organic carbon converted into CH4 is 30%-50%. It is worth noting that AD would also
165
reduce the amount of energy recovered from incineration (Mills et al., 2014).
166
After AD, the organic content (Pv) in digested sludge will be reduced to 37% (the converted organic
167
efficiency is set at 30%) and the targeted moisture content (w2) of thermal drying is calculated at 41.3%
168
(for self-supporting incineration). As a result, the total energy balance for AD plus incineration can be
169
calculated, as shown in Fig. 1b. Obviously, the energy deficit of 324 kW·h/t DS is much higher than that
170
from direct incineration (109 kW·h/t DS). Moreover, the environmental impact of greenhouse gases (CH4)
171
from AD due to leaks from digesters is almost inevitable (Daelman et al., 2012).
172
The investment costs of AD plus incineration would increase to 86 ×104 US$/t DS due to the
173
additional infrastructure and equipment of gravity thickening tanks (about 6 ×104 US$/t DS) and digesters
174
(about 37 ×104 US$/t DS). Additional complications in management can also increase the operational costs,
175
by up to 398 US$/t DS. The detailed calculations of the economic aspects are listed in Table S3 (Hong et al.,
176
2009; Piao et al., 2016).
177
3.2. Pretreatment by thermal hydrolysis prior to anaerobic digestion
178
Thermal hydrolysis could be applied prior to AD as a pretreatment technology in order to increase
179
converted organic efficiency (Ometto et al., 2014). The net energy output of AD with thermal hydrolysis
180
involved can be increased by 40% (Fdz-Polanco et al., 2008). Thermal hydrolysis could increase the
181
converted organic efficiency of AD by up to 50% (Fdz-Polanco et al., 2008), but the targeted moisture
182
content in the following thermal drying unit would have to be reduced to 21.1% for self-supporting
183
incineration, which would make it difficult to fluidize in incinerators. In this case, the moisture content of
184
the dried sludge is still considered at 41.3% (fluidizable sludge) in the calculation, which cannot attain
185
self-supporting incineration and needs an external auxiliary fuel (Samolada et al., 2014). As a result, the
186
energy deficit would reach a higher value, up to 313 kW·h/t DS, as shown in Fig. 1c, similar to the energy
187
deficit from conventional AD plus incineration (324 kW·h/t DS), and also higher than that of direct
188
incineration (109 kW·h/t DS).
189
The organic content of sludge (Pv) is an important parameter, as it directly determines the calorific
190
value of sludge and the moisture content of self-supporting incineration (w2), which affects the energy
191
balance calculations of the three compared processes. The sensitivity analysis of related key data and/or
192
factors (Fig. S1) reveals that the energy deficit decreases as the organic content of sludge rises, but that the
193
energy deficit of the direct incineration process remains the lowest in the three compared processes when
194
the organic content of raw excess sludge is above 38%. In fact, sludge with <38% organic content is not
195
suitable for anaerobic digestion either, and would be more suitable for direct landfilling, if landfilling space
196
permits.
197
For a comparision, the energy deficits and I & O costs (based on Chinese economic data) from the
198
three discussed processes are illustrated in Fig. 2. Direct incineration has the lowest value for both the
199
energy deficit and the I & O costs, indicating the advantage of direct incineration in handling sludge.
200 Fig. 2. Comparison of energy deficit (a) and I & O costs (b) for the three evaluated sludge disposal processes
201
202
4. Discussion
203
Evaluation of different sludge disposal methods based on incineration indicates that direct incineration of
204
sludge results in lower investment and operational costs (Zhang et al., 2006) and is energetically more
205
favorable. This calls into question the often-promoted practice of producing more energy in the form of
206
biogas on-site (Ometto et al., 2014). Although this might give the facility concerned a positive image as
207
being an energy-producing entity, it does not help satisfy the overall societal demand to become more
208
energy-efficient. Only when it is desirable to make a wastewater facility independent of the electricity grid
209
(i.e., no cable connection and saving the fixed costs of electricity supply), might there be an economical
210
advantage to producing biogas on-site.
211
Putting incineration into practice is, in general, not easy. The incinerators have a strong economic
212
benefit when constructed as centralized units. Wastewater treatment plants, however, should be placed near
213
existing population centers to minimize transport of wastewater and to be able to use recovered heat (e.g.
214
from effluent) for a local district heating/cooling system (Hao et al., 2019a). The excess sludge should then
215
be concentrated and dried at these local sites, after which it can be transported to a centralized facility (Hao
216
et al., 2019b). Drying sludge could be conducted with the low-value thermal energy (40-80 ℃) recovered
217
from the effluent of WWTPs (Hao et al., 2019a), which might be the most beneficial option for drying
218
sludge at low temperatures (20-80 ℃) (Font et al., 2011). Within larger utilities, a cost optimisation
219
including the effect of process unit scale and transport distance will result in the optimal placement of the
220
different process units for wastewater treatment, sludge drying and incineration.
221
Future wastewater treatment operation should endeavour to maximise resource recovery with minimal
222
energy usage (Guest et al., 2013; van Loosdrecht et al., 2014). In this framework, incineration can be an
223
essential step. Mineral recovery from the bottom ashes may be a practice that is still its infancy, but it is
224
certainly possible. For instance, incineration with full phosphate recovery (up to 90%) from the bottom
225
ashes is the only option for phosphate recovery when agricultural use of sludge is not feasible (Egle et al.,
226
2015). This would also allow recovery of other metals with often a higher content in the bottom ashes than
227
in the mined ores. However, it is also important to recover organic material (Perez-Feito et al., 2006; Lin et
228
al., 2010). In general, recovery of such materials is preferred over using organic material for energy
229
production (van Loosdrecht et al., 2014). How chemicals recovery will influence future schemes for
230
handling sludge residues remains unclear, but we also believe that in those cases a final incineration step
231
will be the best solution – certainly for the large urban areas in which most of the world’s population
232
resides.
233
5. Conclusions
234
Based on the description and discussion, some key conclusions can be drawn:
235
Land resources are insufficient for receiving sludge from big cities in China and other countries.
236
Farmers are not willing to accept sludge as fertilizer, due to the low fertilizer value and slow release of
237
nutrients.
238
The current focus on using anaerobic digestion to minimize sludge and produce energy is challenged
239
when wastewater handling is evaluated on a system scale. Direct incineration of sludge has the lowest
240
energy deficit and I & O costs, compared to conventional AD (also including pretreatment by thermal
241
hydrolysis).
242
243
Acknowledgments
244
The study was financially supported by the National Natural Science Foundation of China (51878022) and
245
also by the special fund from Beijing Advanced Innovation Center of Future Urban Design (2019).
246
247
Appendix A. Supplementary data
248
Table S1 Conversion between standard coal and electricity; Table S2 Practical investment and operational
249
costs of sludge incineration in China; Table S3 Economic comparison among the three compared sludge
250
disposal processes; Fig. S1 Sensitivity analysis of the sludge organic content (Pv) vs. energy balance
251
252
References
253
Abuşoğlu, A., Özahi, E., İhsan Kutlar, A., Al-jaf, H., 2017. Life cycle assessment (LCA) of digested sewage sludge
254
incineration for heat and power production. J. Clean. Prod. 142, 1684–1692. doi:10.1016/j.jclepro.2016.11.121
255
Cai, L., Chen, T. B., Gao, D., Yang, J., Chen, J., Zheng, G. D., Du, W., 2010. Investigation of calorific value of sewage in
256
large and middle cities of China. China water wastewater. 15, 106-108. doi:10.3724/SP.J.1088.2010.00432 (in Chinese)
257
Chen, S. Q., Wang, F., Chi, Y., Yan, J. H., Cen, K. F., 2017. Analysis on mass and energy balance of sludge drying and
258
incineration system. Chinese Journal of Environmental Engineering. 11, 515-521. doi:10.12030/j.cjee.201508185 (in
259
Chinese)
260 261
Colin, F., Gazbar, S., 1995. Distribution of water in sludges in relation to their mechanical dewatering. Water Research. 29,
2000-2005. doi:10.1016/0043-1354(94)00274-b
262 263 264 265 266 267
Daelman, M. R. J., Voorthuizen, E. M. V., Dongen, U. G. J. M. V., Volcke, E. I. P., Loosdrecht, M. C. M. V., 2012. Methane
emission during municipal wastewater treatment. Water Research, 46, 3657-3670. doi:10.1016/j.watres.2012.04.024
Egle, L., Rechberger, H., Zessner, M., 2015. Overview and description of technologies for recovering phosphorus from
municipal wastewater. Resources, Conservation and Recycling. 105, 325-346. doi:10.1016/j.resconrec.2015.09.016
Federation, W. E., 2009. Wastewater solids incineration systems. USA: Business Expert Press. 200-202.
http://agris.fao.org/agris-search/search.do?recordID=US201300138039
268
Fernández-Arévalo, T., Lizarralde, I., Fdz-Polanco, F., Pérez-Elvira, S.I., Garrido, J.M., Puig, S., Poch, M., Grau, P., Ayesa,
269
E., 2017. Quantitative assessment of energy and resource recovery in wastewater treatment plants based on plant-wide
270
simulations. Water Research. 118, 272–288. doi:10.1016/j.watres.2017.04.001
271
Fdz-Polanco, F., Velazquez, R., Perez-Elvira, S.I., Casas, C., Del-Barrio, D., Cantero, F. J., Fdz-Polanco, M., Rodriguez, P.,
272
Panizo, L., Serrat, J., Rouge, P., 2008. Continuous thermal hydrolysis and energy integration in sludge anaerobic
273
digestion plants. Water Science & Technology. 57, 1221-1226. doi:10.2166/wst.2008.072
274 275 276 277 278 279
Font, R., Gomez-Rico, M. F., Fullana, A., 2011. Skin effect in the heat and mass transfer model for sewage sludge drying.
Separation & Purification Technology. 77, 146-161. doi:10.1016/j.seppur.2010.12.001
Fytili, D., Zabaniotou, A., 2008. Utilization of sewage sludge in EU application of old and new methods-A review.
Renewable and Sustainable Energy Reviews. 12, 116–140. doi:10.1016/j.rser.2006.05.014
Gross, T. S. C., 2010. Thermal drying of sewage sludge. Water & Environment Journal. 7, 255-261.
doi:10.1111/j.1747-6593.1993.tb00843.x
280
Guest, J. S., van Loosdrecht, M. C. M., Skerlos, S. J., Love, N. G., 2013. Lumped pathway metabolic model of organic
281
carbon accumulation and mobilization by the alga chlamydomonas reinhardtii. Environmental Science & Technology, 47,
282
3258-3267. doi:10.1021/es304980y
283
Hao X.D., Li J., van Loosdrecht M. C. M., Jang, H., Liu, R. B., 2019a. Energy recovery from wastewater: Heat over organics.
284 285 286 287 288 289 290 291 292 293 294 295 296
Water Research. 161, 74-77. doi:10.1016/j.watres.2019.05.106
Hao, X. D., Liu, R. B., Huang, X., 2015. Evaluation of the potential for operating carbon neutral WWTPs in China. Water
Research. 87, 424–431. doi:10.1016/j.watres.2015.05.050
Hao, X. D., Wang, X. Y., Liu, R. B., Li, S., van Loosdrecht M. C. M., 2019b. Environmental impacts of resource recovery
from wastewater treatment plants. Water Research. 160, 268-277. doi:10.1016/j.watres.2019.05.068 Hao, X. D., Yu, J. L., Fu, K. M., Jiang, H., 2019c. Resource of “Nutrient Water” ignored by rural wastewater treatment. China water & wastewater. 35, 5-12. (in Chinese) Hong, J., Hong, J., Otaki, M., Jolliet, O., 2009. Environmental and economic life cycle assessment for sewage sludge treatment processes in Japan. Waste Management. 29, 696–703. doi:10.1016/j.wasman.2008.03.026 Khalili, A., Jamshidi, S., Khalesidoust, M., Vesali, M., 2017. Evaluation of sewage sludge for incineration (case study: Arak
wastewater treatment plant). Environmental Energy and Economic Research. 1, 249–258. doi:10.22097/eeer.2017.47251
Liao, C. H., Zhao, J. N., Wang, L., 2012. Introduction and analysis of performance evaluating index of combined heat and
power system at home and abroad. Gas & Heat. 32, 4-9. doi:10.3969/j.issn.1000-4416.2012.01.002 (in Chinese)
297
Lin, X. Q., Yan, M., Dai, A. H., Zhan, M. X., Fu, J. Y., Li, X. D., Chen, T., Lu, S. Y., Alfons, B., Yan, J. H., 2015.
298
Simultaneous suppression of PCDD/F and NOX during municipal solid waste incineration. Chemosphere, 126, 60-66.
299
doi:10.1016/j.chemosphere.2015.02.005
300
Lin, Y. M., Kreuk, M. D., van Loosdrecht, M. C. M., Adin, A., 2010. Characterization of alginate-like exopolysaccharides
301
isolated from aerobic granular sludge in pilot-plant. Water Research, 44, 3355-3364. doi:10.1016/j.watres.2010.03.019
302
Li, S., Li, Y., Lu, Q., Zhu, J., Yao, Y., Bao, S., 2014. Integrated drying and incineration of wet sewage sludge in combined
303
bubbling and circulating fluidized bed units. Waste Management. 34, 2561–2566. doi:10.1016/j.wasman.2014.08.018
304
Lowe, P., 2010. Developments in the thermal drying of sewage sludge. Water & Environment Journal. 9, 306-316.
305 306
doi:10.1111/j.1747-6593.1995.tb00944.x
Mills, N., Pearce, P., Farrow, J., Thorpe, R. B., Kirkby, N. F., 2014. Environmental & economic life cycle assessment of
307
current
308
doi10.1016/j.wasman.2013.08.024
&
future
sewage
sludge
to
energy
technologies.
Waste
Management.
34,
185-195.
309
Murakami, T., Suzuki, Y., Nagasawa, H., Yamamoto, T., Koseki, T., Hirose, H., Okamoto, S., 2009. Combustion
310
characteristics of sewage sludge in an incineration plant for energy recovery. Fuel Processing Technology. 90, 778–783.
311
doi:10.1016/j.fuproc.2009.03.003
312
Novak, J. T., 2006. Dewatering of sewage sludge. Drying Technology. 24, 1257-1262. doi:10.1080/07373930600840419
313
Ometto, F., Quiroga, G., Penika, P., Whitton, R., Jefferson, B., Villa, R., 2014. Impacts of microalgae pre-treatments for
314
improved anaerobic digestion: thermal treatment, thermal hydrolysis, ultrasound and enzymatic hydrolysis. Water
315
Research. 65, 350-361. doi:10.1016/j.watres.2014.07.040
316
Perez-Feito, R., Noguera, D. R., 2006. Recovery of polyhydroxyalkanoate from activated sludge in an enhanced biological
317
phosphorus removal bench-scale reactor. Water Environment Research. 78, 770-775. doi:10.2175/106143006X99803
318
Piao, W., Kim, Y., Kim, H., Kim, M., Kim, C., 2016. Life cycle assessment and economic efficiency analysis of integrated
319
management
320
doi:10.1016/j.jclepro.2015.11.012
321 322
of
wastewater
treatment
plants.
Journal
of
Cleaner
Production.
113,
325–337.
Qiu, Z. H., Dai, H. Y., Kan, A. K., 2007. Energy conservation analysis of gas combined heat and power system (CHP).
Energy research and information. 23, 237-242. doi:10.3969/j.issn.1008-8857.2007.04.010 (in Chinese)
323
Samolada, M. C., Zabaniotou, A. A., 2014. Comparative assessment of municipal sewage sludge incineration, gasification
324
and pyrolysis for a sustainable sludge-to-energy management in Greece. Waste Management. 34, 411-420.
325
doi:10.1016/j.wasman.2013.11.003
326 327 328
Suh, Y., Rousseaux, P., 2002. An LCA of alternative wastewater sludge treatment scenarios. Resource, Conservation,
Recycling. 35, 191-200. doi:10.1016/S0921-3449(01)00120-3
Svoboda, K., Baxter, D., Martinec, J., 2006. Nitrous oxide emissions from waste incineration. Chemical Papers. 60, 78-90.
329 330 331
doi:10.2478/s11696-006-0016-x
Thomas, V. M., Mccreight, C. M., 2008. Relation of chlorine, copper and sulphur to dioxin emission factors. Journal of
Hazardous Materials. 151, 164-170. doi:10.1016/j.jhazmat.2007.05.062
332
Thomsen, M., Seghetta, M., Mikkelsen, M.H., Gyldenkærne, S., Becker, T., Caro, D., Frederiksen, P., 2017. Comparative life
333
cycle assessment of biowaste to resource management systems-A Danish case study. Journal of Cleaner Production. 142,
334
4050–4058. doi:10.1016/j.jclepro.2016.10.034
335 336 337 338
van Loosdrecht, M. C. M., Brdjanovic, D., 2014. Anticipating the next century of wastewater treatment. Science. 344,
1452-1453. doi:10.1126/science.1255183
Wakeman, R. J., 2007. Separation technologies for sludge dewatering. Journal of Hazardous Materials. 144, 614–619.
doi:10.1016/j.jhazmat.2007.01.084
339
Winkler, M.K.H., Bennenbroek, M.H., Horstink, F.H., van Loosdrecht, M.C.M., van de Pol, G.J., 2013. The biodrying
340
concept: An innovative technology creating energy from sewage sludge. Bioresource Technology. 147, 124–129.
341
doi:10.1016/j.biortech.2013.07.138
342 343 344 345
Xu, G. R., Zou, J. L., Li, G. B., 2008. Stabilization of heavy metals in ceramsite made with sewage sludge. Journal of
Hazardous Materials. 152, 56-61. doi:10.1016/j.jhazmat.2007.06.072
Yao, H., Naruse, I., 2005. Control of trace metal emissions by sorbents during sewage sludge combustion. Symposium on
Combustion. 30, 3009-3016. doi:10.1016/j.proci.2004.07.047
346
Yoshida, H., ten Hoeve, M., Christensen, T. H., Bruun, S., Jensen, L.S., Scheutz, C., 2018. Life cycle assessment of sewage
347
sludge management options including long-term impacts after land application. Journal of Cleaner Production. 174,
348
538–547. doi:10.1016/j.jclepro.2017.10.17
349 350
Zhan, M. X., Fu, J. Y., Chen, T., Lin, X. Q., Li, X. D., Yan, J. H., Alfons, B., 2016. Suppression of dioxins by S-N inhibitors
in pilot-scale experiments. Environmental Science and Pollution Research. 23, 63-77. doi:10.1007/s11356-016-6401-2
351
Zhang, Y. A., Gao, D., Chen, T. B., Zheng, G. D., Li, Y. X., 2006. Economical evaluation of different techniques to treatment
352
and
353
doi:10.3969/j.issn.1674-5906.2006.02.007 (in Chinese)
354 355
dispose
sewage
sludge
in
Beijing.
Ecology
and
Environment.
15,
234-238.
Zhou, H., Meng, A., Long, Y., Li, Q., Zhang, Y., 2015. A review of dioxin-related substances during municipal solid waste
incineration. Waste Management. 36, 106-118. doi:10.1016/j.wasman.2014.11.011
Fig. 1. Energy balances of the three compared sludge disposal processes.
Fig. 2. Comparison of energy deficit (a) and I & O costs (b) for the three evaluated sludge disposal processes
Highlights • Landfills and agricultural use of excess sludge is becoming increasingly constrained • Alternative sludge disposal routes must be sought • Anaerobic digestion (AD) is not a final solution for excess sludge • Direct incineration without AD has the lowest energy deficit and I & O costs • Focus on more efficient dewatering instead of AD is recommended
S0043-1354(19)31072-3
DOI:
https://doi.org/10.1016/j.watres.2019.115298
Reference:
WR 115298
To appear in:
Water Research
Received Date: 25 July 2019 Revised Date:
4 November 2019
Accepted Date: 8 November 2019
Please cite this article as: Hao, X., Chen, Q., van Loosdrecht, M.C.M., Li, J., Jiang, H., Sustainable disposal of excess sludge: Incineration without anaerobic digestion, Water Research (2019), doi: https:// doi.org/10.1016/j.watres.2019.115298. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.
Graphical abstract
324 86/398
Legend:
Energy deficit (kW·h /t DS) Investment/Operational costs (×104 US$/t DS)/(US$/t DS)
Anaerobic digestion 313 90/445
Anaerobic digestion
Thermal hydrolysis 109 55/392 Excess sludge
Sustainability！
Incineration
Sustainable disposal of excess sludge: incineration without
1
anaerobic digestion
2
Xiaodi Hao a, *, Qi Chen a, Mark C. M. van Loosdrecht a, b, Ji Li a, Han Jiang a
3
4
a
5
Stormwater System and Water Environment, Beijing University of Civil Engineering & Architecture,
6
Beijing 100044, P. R. China
7
b
8
Netherlands
9
____________________________
10
Sino-Dutch R&D Centre for Future Wastewater Treatment Technologies/Key Laboratory of Urban
Department of Biotechnology, Delft University of Technology, van der Maasweg 9, 2629 HZ Delft, the
*
Corresponding author: Tel: +86 131 6134 7675; Fax: +86 10 6832 2123; E-mail: [email protected]
11
12
ABSTRACT:
13
Handling excess sludge produced by wastewater treatment is a common problem worldwide. Due to limited
14
space available in landfills, as well as difficulties involved in using excess sludge in agriculture, there is a
15
need for alternative disposal methods. Although anaerobic digestion (AD) is widely used in processing
16
sludge, only partial energy recovery from methane and sludge volume reduction can be achieved, resulting
17
in a substantial amount of sludge remaining, which needs to be disposed of. Direct incineration after sludge
18
drying is one possible option, a practice that is already in place in some cities in China. A comparison
19
between direct incineration and conventional AD (with or without pretreatment by thermal hydrolysis) has
20
to be made with respect to the energy balance and investment & operational (I & O) costs. This comparison
21
reveals direct incineration to have the lowest energy deficit and I & O costs. Therefore, it is expected that
22
direct incineration without AD will become the preferred sustainable approach to handling sludge.
23
Keywords: Sludge disposal; direct incineration; anaerobic digestion; sludge drying; energy deficit;
24
investment & operational costs
25
26
1. Introduction
27
Nowadays, the main methods of treating wastewater involve activated sludge, which makes excess sludge
28
an unavoidable problem. Landfilling and agricultural & horticultural use are the most convenient and
29
economical approaches to handling sludge (Suh et al., 2002). However, lack of landfilling spaces is a
30
prevailing problem for big cities, both worldwide and in China. In addition, agricultural applications are
31
limited by the low fertilizer value of sludge and the slow release of its nutrients, which helps farmers little
32
in raising grain production. Chinese farmers would rather use chemical fertilizers for crop cultivation, and
33
even tend to give up their conventionally “ecological” habits on applying manure as fertilizer (Hao et al.,
34
2019c).
35
Under these circumstances, China has to search for alternative methods of handling excess sludge. In
36
Europe and several other areas, sludge thickening followed by anaerobic digestion (AD) is commonly
37
applied to reduce the excess sludge amounts (Khalili et al., 2017; Thomsen et al., 2017). However, AD does
38
not remove all sludge organics, with as much as 50%-70% of the organics remaining (Hao et al., 2015;
39
Fernández-Arévalo et al., 2017). For this reason, other disposal methods, such as incineration, are still
40
necessary (Yoshida et al., 2018).
41
AD, followed by incineration, has a long history in Europe and in other countries (Abuşoğlu et al.,
42
2017), but direct incineration after dewatering and drying (reducing the moisture content of sludge to
43
40%-70%) (Fytili et al., 2008; Abuşoğlu et al., 2017) is also currently practiced in some Chinese cities.
44
Which way is better, AD plus incineration or direct incineration? This question calls for a comparative
45
study concerning the energy balance and investment & operational (I & O) costs.
46
Traditionally, producing biogas is seen as a sustainable approach to energy-efficient wastewater
47
treatment. However, in these cases the boundaries for evaluation are often within the wastewater treatment
48
plant itself. If direct incineration is, overall, more effective than AD plus incineration, it could conceivably
49
become a sustainable approach to handling sludge, and China would be able to leap forward in the
50
development on its sludge disposal. It was that idea that prompted this study.
51
2. Direct incineration
52
The proposed process of direct incineration consists of three steps: mechanical dewatering, thermal drying,
53
and incineration. Dewatering is applied to reduce the moisture content of sludge, from about 99% to 80%,
54
which creates a considerable reduction in sludge volume, by about 95%. However, this sludge is still far
55
from suitable for self-supporting incineration (Abuşoğlu et al., 2017). With thermal drying, the moisture
56
content of dewatered sludge can be reduced from 80% to 40%-70% (dependent on the organic content of
57
the sludge), which allows for self-supporting incineration (Fytili et al., 2008; Abuşoğlu et al., 2017).
58
Following this, organics contained in dried sludge can be incinerated directly, and fully oxidized into CO2
59
in incinerators at a high temperature of 800-900 °C or higher, with thermal energy generation, and with
60
phosphorus and other minerals remaining in the ashes (Abuşoğlu et al., 2017). The thermal energy can be
61
used for generating electricity, for drying sludge on site, or for municipal heating/cooling systems. Finally,
62
phosphorus and other minerals or metals can be recovered from the ashes (Murakami et al., 2009).
63
2.1. Energy balance
64
2.1.1. Mechanical dewatering
65
There are already many types of mechanical dewatering processes, including press and vacuum filtration,
66
centrifugation, etc. (Colin et al., 1995). Mechanical dewatering energy consumption depends mainly on the
67
structure of sludge flocs and the complex shear force between sludge and water, as well as different types
68
of dewatering equipment and chemicals dosage, and is not dependent on the different origins of sludge
69
(Novak, 2006). The energy consumption of different processes varies from 15-179 kW·h/t DS (Wakeman,
70
2007). The belt press filter, which has an energy consumption of about 60 kW·h/t DS (the moisture content
71
of sludge is reduced from 99% to 80%), is extensively applied in China (Wakeman, 2007).
72
2.1.2. Thermal drying
73
There are two types of thermal drying processes: i) the full drying process, whereby moisture content can
74
be reduced to ≤15%, and which is generally conducted by rotating-drum dryers; ii) the partial drying
75
process where moisture content is usually reduced to 35%-50%, often operated by rotary-plate dryers
76
(Gross, 2010; Lowe, 2010). Dried sludge with ≤70% moisture content can be self-incinerated without the
77
input of any auxiliary fuels (Fytili et al., 2008; Abuşoğlu et al., 2017), and so a partial drying process is
78
sufficient in this case. The total energy consumption (Et) of the thermal drying process consists of an
79
increased solids’ temperature (Es) and water evaporation (Ew), as expressed in Eqs. 1 and 2 (Chen et al.,
80
2017). Es =(T2 -T1 )∙Cs ∙Ms ∙100
（1）
81
Where, Es is the energy consumption of the temperature increase of the solids in the dewatered sludge, kJ/t
82
DS; Cs is the specific heat capacity of dewatered sludge, 3.62 kJ/kg· oC; T1 and T2 are the temperatures of
83
the input (20 oC) and output (100 oC) sludge, respectively; and Ms is the weight of dry matter of sludge
84
(DS), which is a constant value: 10 kg/t wet sludge (99% moisture content). Ew =Cw ∙ Mw =[
Ms w1 ∙ T2 -T1 ∙100+Qg ∙Mw 1-w1
Ms Ms ]∙100 1-w1 1-w2
（2）
（3）
85
Where, Ew is the energy consumption of water evaporation, kJ/t DS, including heating water and
86
evaporating water; Cw is the specific heat capacity of water, 4.2 kJ/kg· oC; w1 and w2 are the moisture
87
content of the input (80%) and output (40%-70%) sludge, respectively; Qg is the potential vaporization heat
88
value of water, 2,260 kJ/kg; and Mw is the weight of evaporated water, kg, which can be calculated by Eq. 3
89
(Chen et al., 2017).
90 91
The most important parameter is the targeted moisture content (w2) of dried sludge after thermal drying, which can be calculated by Eq. 4 (Murakami et al., 2009; Mills et al., 2014; Samolada et al., 2014). QsL=QsH ∙ 1-w2 -Qg ∙(w2 +9wH )
（4）
92
Where, QsL is the low calorific value of sludge, set at 3.36 GJ/t DS for self-supporting incineration; QsH is
93
the high calorific value of sludge, which is the maximal potential energy contained in dried sludge, and
94
which can be calculated by Eq. 6 (Cai et al., 2010), in GJ/t DS; and wH is the mass fraction of the organic
95
bound hydrogen element in the dry matter (about 2% of the dry matter) (Murakami et al., 2009; Samolada
96
et al., 2014).
97
Because of heat loss in sludge dryers, the actual energy consumption of the drying processes is
98
obviously higher than the theoretical energy consumption. The actual energy consumption of sludge drying
99
can be calculated by Eq. 5. The heat loss efficiency (η) of different types of thermal dryers ranges from 10%
100
to 20% (Li et al., 2014) and here the highest value (20%) is used to calculate the energy consumption. Et '=Et ∙ 1+η
（5）
101
2.1.3. Incineration
102
The theoretical thermal energy released from incineration (Q) is equivalent to the high calorific value of the
103
dried sludge (QsH), which can be calculated by Eq. 6 (Cai et al., 2010). Q=2.5×105×(
100·Pv 100-Pc -5)× 100-Pc 100
（6）
104
Where, Q is the theoretical thermal energy released from incineration; Pv is the percentage of the organic
105
content; Pc is the percentage of inorganic coagulants added to the sludge during mechanical dewatering
106
(thus, 0% when organic polymer coagulants are added).
107
The organic content of sludge (Pv) in China is normally within the range of 30%-65%. In this study, Pv
108
was set at 53% (the typical mean value for China) (Cai et al., 2010). Therefore, the theoretical thermal
109
energy released from incineration is calculated at 11.9 GJ/t DS, equivalent to the mean value of thermal
110
energy released across China (11.85 GJ/t DS) (Cai et al., 2010; Winkler et al., 2013), and the moisture
111
content (w2) of dried sludge is calculated at 57.7% by Eq. 4. Based on the above equations, the actual
112
energy consumption of the thermal drying process is 9.1 GJ/t DS, approaching to data used in process
113
designs of 11.7 GJ/t DS (Chen et al., 2017). When converted to its electrical equivalent, the actual energy
114
consumption is 2,529 kW·h/t DS (1 kW·h=3,600 kJ).
115
Of course, the actual energy recovered from incineration would be lower than the value calculated
116
above, due to factors as the low efficiency of fluidization, the incomplete combustion of solids and gases,
117
heat loss from incinerators, low-temperature off-gases (all resulting in inefficiencies in the electricity
118
generation), etc. (Murakami et al., 2009; Mills et al., 2014). According to engineering data, the total
119
thermal energy loss is about 7% of the theoretical energy released from incineration (11.9 GJ/t DS), which
120
is 11.9 GJ/t DS×7%=0.8 GJ/t DS (Li et al., 2014). Thus, the actual energy recovered from incineration is
121
around 11.1 GJ/t DS, which consists of two parts: i) 84% high-temperature thermal energy (9.32 GJ/ t DS)
122
(Qiu et al., 2007; Liao et al., 2012) which can be directly converted into electricity via Combined Heat and
123
Power (CHP) (80%, 7.46 GJ/ t DS), resulting in a left-over thermal energy (20%, 1.86 GJ/t DS); and ii) 16%
124
for low-temperature waste thermal energy (1.78 GJ/t DS) (Qiu et al., 2007; Liao et al., 2012). Both low
125
thermal energy sources (3.64 GJ/t DS)) could be counted as “electrical equivalent” (based on Table S1),
126
with a thermal utilization efficiency of up to 40%. Based on the practical calculations shown above, the
127
total electrical energy (2,072 kW·h/t DS) plus electrical equivalent (408 kW·h/t DS) generated by sludge
128
incineration is about 2,480 kW·h/t DS, or approximately 80% of the total energy.
129
2.1.4. Energy balance
130
Based on the energy estimates calculated above, the energy balance of the direct incineration process is
131
shown in Fig. 1a, revealing an energy deficit of 109 kW·h/t DS.
132 Fig. 1. Energy balances of the three compared sludge disposal processes.
133 134
2.2. Investment & operational costs
135
I & O costs depend on the capacity of wastewater treatment plants (WWTPs) and applied equipment. In
136
this study, a WWTP with a treatment capacity of 500,000 m3/d was used to estimate the I & O costs
137
according to practical data from wastewater treatment processes in China (Table S2). The plant was able to
138
produce 8,000 t wet sludge/d (99% moisture content). The investment costs consisted of infrastructure and
139
equipment. The operational costs consisted of electricity, water, chemicals, transportation, wages and
140
welfare, depreciation of fixed assets, and overhaul & maintenance costs. The detailed calculations on the
141
economic aspects are listed in Table S3 (Hong et al., 2009; Piao et al., 2016). The I & O costs of the direct
142
incineration process were calculated according to the Chinese market price levels, at 55 ×104 US$/t DS and
143
392 US$/t DS, respectively, which falls in the range of the practical data from several engineering cases in
144
China.
145
2.3. Environmental impacts
146
Nitrogen oxides (NOx), dioxins and even heavy metals can all be released from sludge during incineration.
147
At temperatures above 850 °C, however, the average released concentration (471.6 mg/m3) of NOx is lower
148
than the emission standard in China (500 mg/m3) (Lin et al., 2015). Even referring to the stringent EU
149
emission standard of NOx (400 mg/m3), NOx emission can easily be controlled by several different
150
technologies (Svoboda et al., 2006). Furthermore, the decomposition rate of dioxins is much greater than its
151
formation rate above 850 °C (Thomas et al., 2008; Zhan et al., 2016), and thus the net concentration of
152
dioxins produced can be lower than the EU emission standard (0.1 ng-TEQ/m3 off-gases) without any
153
treatment (Zhou et al., 2015). In addition, heavy metals (including mercury) are also not a serious concern
154
(Xu et al., 2008), and all of these can effectively be controlled above 850 °C (Yao et al., 2005). In short, the
155
3T+E (Temperature, Time, Turbulence and Excess) control strategy on incineration can greatly help to
156
reduce off-gases pollutants concentrations of incineration (Zhou et al., 2015). Research reports from the
157
United States of America, Germany, Great Britain and Spain have all reached similar conclusions: there is
158
no evidence that sludge incineration and even waste incineration will endanger the environment or human
159
health (Federation, 2009).
160
3. Anaerobic digestion followed by incineration
161
3.1. Conventional anaerobic digestion plus incineration
162
AD plus incineration generally includes five steps: gravity thickening, AD, dewatering, thermal drying and
163
incineration. In gravity thickening, the moisture content of sludge is reduced from 99% to 97%. During AD,
164
the fraction of organic carbon converted into CH4 is 30%-50%. It is worth noting that AD would also
165
reduce the amount of energy recovered from incineration (Mills et al., 2014).
166
After AD, the organic content (Pv) in digested sludge will be reduced to 37% (the converted organic
167
efficiency is set at 30%) and the targeted moisture content (w2) of thermal drying is calculated at 41.3%
168
(for self-supporting incineration). As a result, the total energy balance for AD plus incineration can be
169
calculated, as shown in Fig. 1b. Obviously, the energy deficit of 324 kW·h/t DS is much higher than that
170
from direct incineration (109 kW·h/t DS). Moreover, the environmental impact of greenhouse gases (CH4)
171
from AD due to leaks from digesters is almost inevitable (Daelman et al., 2012).
172
The investment costs of AD plus incineration would increase to 86 ×104 US$/t DS due to the
173
additional infrastructure and equipment of gravity thickening tanks (about 6 ×104 US$/t DS) and digesters
174
(about 37 ×104 US$/t DS). Additional complications in management can also increase the operational costs,
175
by up to 398 US$/t DS. The detailed calculations of the economic aspects are listed in Table S3 (Hong et al.,
176
2009; Piao et al., 2016).
177
3.2. Pretreatment by thermal hydrolysis prior to anaerobic digestion
178
Thermal hydrolysis could be applied prior to AD as a pretreatment technology in order to increase
179
converted organic efficiency (Ometto et al., 2014). The net energy output of AD with thermal hydrolysis
180
involved can be increased by 40% (Fdz-Polanco et al., 2008). Thermal hydrolysis could increase the
181
converted organic efficiency of AD by up to 50% (Fdz-Polanco et al., 2008), but the targeted moisture
182
content in the following thermal drying unit would have to be reduced to 21.1% for self-supporting
183
incineration, which would make it difficult to fluidize in incinerators. In this case, the moisture content of
184
the dried sludge is still considered at 41.3% (fluidizable sludge) in the calculation, which cannot attain
185
self-supporting incineration and needs an external auxiliary fuel (Samolada et al., 2014). As a result, the
186
energy deficit would reach a higher value, up to 313 kW·h/t DS, as shown in Fig. 1c, similar to the energy
187
deficit from conventional AD plus incineration (324 kW·h/t DS), and also higher than that of direct
188
incineration (109 kW·h/t DS).
189
The organic content of sludge (Pv) is an important parameter, as it directly determines the calorific
190
value of sludge and the moisture content of self-supporting incineration (w2), which affects the energy
191
balance calculations of the three compared processes. The sensitivity analysis of related key data and/or
192
factors (Fig. S1) reveals that the energy deficit decreases as the organic content of sludge rises, but that the
193
energy deficit of the direct incineration process remains the lowest in the three compared processes when
194
the organic content of raw excess sludge is above 38%. In fact, sludge with <38% organic content is not
195
suitable for anaerobic digestion either, and would be more suitable for direct landfilling, if landfilling space
196
permits.
197
For a comparision, the energy deficits and I & O costs (based on Chinese economic data) from the
198
three discussed processes are illustrated in Fig. 2. Direct incineration has the lowest value for both the
199
energy deficit and the I & O costs, indicating the advantage of direct incineration in handling sludge.
200 Fig. 2. Comparison of energy deficit (a) and I & O costs (b) for the three evaluated sludge disposal processes
201
202
4. Discussion
203
Evaluation of different sludge disposal methods based on incineration indicates that direct incineration of
204
sludge results in lower investment and operational costs (Zhang et al., 2006) and is energetically more
205
favorable. This calls into question the often-promoted practice of producing more energy in the form of
206
biogas on-site (Ometto et al., 2014). Although this might give the facility concerned a positive image as
207
being an energy-producing entity, it does not help satisfy the overall societal demand to become more
208
energy-efficient. Only when it is desirable to make a wastewater facility independent of the electricity grid
209
(i.e., no cable connection and saving the fixed costs of electricity supply), might there be an economical
210
advantage to producing biogas on-site.
211
Putting incineration into practice is, in general, not easy. The incinerators have a strong economic
212
benefit when constructed as centralized units. Wastewater treatment plants, however, should be placed near
213
existing population centers to minimize transport of wastewater and to be able to use recovered heat (e.g.
214
from effluent) for a local district heating/cooling system (Hao et al., 2019a). The excess sludge should then
215
be concentrated and dried at these local sites, after which it can be transported to a centralized facility (Hao
216
et al., 2019b). Drying sludge could be conducted with the low-value thermal energy (40-80 ℃) recovered
217
from the effluent of WWTPs (Hao et al., 2019a), which might be the most beneficial option for drying
218
sludge at low temperatures (20-80 ℃) (Font et al., 2011). Within larger utilities, a cost optimisation
219
including the effect of process unit scale and transport distance will result in the optimal placement of the
220
different process units for wastewater treatment, sludge drying and incineration.
221
Future wastewater treatment operation should endeavour to maximise resource recovery with minimal
222
energy usage (Guest et al., 2013; van Loosdrecht et al., 2014). In this framework, incineration can be an
223
essential step. Mineral recovery from the bottom ashes may be a practice that is still its infancy, but it is
224
certainly possible. For instance, incineration with full phosphate recovery (up to 90%) from the bottom
225
ashes is the only option for phosphate recovery when agricultural use of sludge is not feasible (Egle et al.,
226
2015). This would also allow recovery of other metals with often a higher content in the bottom ashes than
227
in the mined ores. However, it is also important to recover organic material (Perez-Feito et al., 2006; Lin et
228
al., 2010). In general, recovery of such materials is preferred over using organic material for energy
229
production (van Loosdrecht et al., 2014). How chemicals recovery will influence future schemes for
230
handling sludge residues remains unclear, but we also believe that in those cases a final incineration step
231
will be the best solution – certainly for the large urban areas in which most of the world’s population
232
resides.
233
5. Conclusions
234
Based on the description and discussion, some key conclusions can be drawn:
235
Land resources are insufficient for receiving sludge from big cities in China and other countries.
236
Farmers are not willing to accept sludge as fertilizer, due to the low fertilizer value and slow release of
237
nutrients.
238
The current focus on using anaerobic digestion to minimize sludge and produce energy is challenged
239
when wastewater handling is evaluated on a system scale. Direct incineration of sludge has the lowest
240
energy deficit and I & O costs, compared to conventional AD (also including pretreatment by thermal
241
hydrolysis).
242
243
Acknowledgments
244
The study was financially supported by the National Natural Science Foundation of China (51878022) and
245
also by the special fund from Beijing Advanced Innovation Center of Future Urban Design (2019).
246
247
Appendix A. Supplementary data
248
Table S1 Conversion between standard coal and electricity; Table S2 Practical investment and operational
249
costs of sludge incineration in China; Table S3 Economic comparison among the three compared sludge
250
disposal processes; Fig. S1 Sensitivity analysis of the sludge organic content (Pv) vs. energy balance
251
252
References
253
Abuşoğlu, A., Özahi, E., İhsan Kutlar, A., Al-jaf, H., 2017. Life cycle assessment (LCA) of digested sewage sludge
254
incineration for heat and power production. J. Clean. Prod. 142, 1684–1692. doi:10.1016/j.jclepro.2016.11.121
255
Cai, L., Chen, T. B., Gao, D., Yang, J., Chen, J., Zheng, G. D., Du, W., 2010. Investigation of calorific value of sewage in
256
large and middle cities of China. China water wastewater. 15, 106-108. doi:10.3724/SP.J.1088.2010.00432 (in Chinese)
257
Chen, S. Q., Wang, F., Chi, Y., Yan, J. H., Cen, K. F., 2017. Analysis on mass and energy balance of sludge drying and
258
incineration system. Chinese Journal of Environmental Engineering. 11, 515-521. doi:10.12030/j.cjee.201508185 (in
259
Chinese)
260 261
Colin, F., Gazbar, S., 1995. Distribution of water in sludges in relation to their mechanical dewatering. Water Research. 29,
2000-2005. doi:10.1016/0043-1354(94)00274-b
262 263 264 265 266 267
Daelman, M. R. J., Voorthuizen, E. M. V., Dongen, U. G. J. M. V., Volcke, E. I. P., Loosdrecht, M. C. M. V., 2012. Methane
emission during municipal wastewater treatment. Water Research, 46, 3657-3670. doi:10.1016/j.watres.2012.04.024
Egle, L., Rechberger, H., Zessner, M., 2015. Overview and description of technologies for recovering phosphorus from
municipal wastewater. Resources, Conservation and Recycling. 105, 325-346. doi:10.1016/j.resconrec.2015.09.016
Federation, W. E., 2009. Wastewater solids incineration systems. USA: Business Expert Press. 200-202.
http://agris.fao.org/agris-search/search.do?recordID=US201300138039
268
Fernández-Arévalo, T., Lizarralde, I., Fdz-Polanco, F., Pérez-Elvira, S.I., Garrido, J.M., Puig, S., Poch, M., Grau, P., Ayesa,
269
E., 2017. Quantitative assessment of energy and resource recovery in wastewater treatment plants based on plant-wide
270
simulations. Water Research. 118, 272–288. doi:10.1016/j.watres.2017.04.001
271
Fdz-Polanco, F., Velazquez, R., Perez-Elvira, S.I., Casas, C., Del-Barrio, D., Cantero, F. J., Fdz-Polanco, M., Rodriguez, P.,
272
Panizo, L., Serrat, J., Rouge, P., 2008. Continuous thermal hydrolysis and energy integration in sludge anaerobic
273
digestion plants. Water Science & Technology. 57, 1221-1226. doi:10.2166/wst.2008.072
274 275 276 277 278 279
Font, R., Gomez-Rico, M. F., Fullana, A., 2011. Skin effect in the heat and mass transfer model for sewage sludge drying.
Separation & Purification Technology. 77, 146-161. doi:10.1016/j.seppur.2010.12.001
Fytili, D., Zabaniotou, A., 2008. Utilization of sewage sludge in EU application of old and new methods-A review.
Renewable and Sustainable Energy Reviews. 12, 116–140. doi:10.1016/j.rser.2006.05.014
Gross, T. S. C., 2010. Thermal drying of sewage sludge. Water & Environment Journal. 7, 255-261.
doi:10.1111/j.1747-6593.1993.tb00843.x
280
Guest, J. S., van Loosdrecht, M. C. M., Skerlos, S. J., Love, N. G., 2013. Lumped pathway metabolic model of organic
281
carbon accumulation and mobilization by the alga chlamydomonas reinhardtii. Environmental Science & Technology, 47,
282
3258-3267. doi:10.1021/es304980y
283
Hao X.D., Li J., van Loosdrecht M. C. M., Jang, H., Liu, R. B., 2019a. Energy recovery from wastewater: Heat over organics.
284 285 286 287 288 289 290 291 292 293 294 295 296
Water Research. 161, 74-77. doi:10.1016/j.watres.2019.05.106
Hao, X. D., Liu, R. B., Huang, X., 2015. Evaluation of the potential for operating carbon neutral WWTPs in China. Water
Research. 87, 424–431. doi:10.1016/j.watres.2015.05.050
Hao, X. D., Wang, X. Y., Liu, R. B., Li, S., van Loosdrecht M. C. M., 2019b. Environmental impacts of resource recovery
from wastewater treatment plants. Water Research. 160, 268-277. doi:10.1016/j.watres.2019.05.068 Hao, X. D., Yu, J. L., Fu, K. M., Jiang, H., 2019c. Resource of “Nutrient Water” ignored by rural wastewater treatment. China water & wastewater. 35, 5-12. (in Chinese) Hong, J., Hong, J., Otaki, M., Jolliet, O., 2009. Environmental and economic life cycle assessment for sewage sludge treatment processes in Japan. Waste Management. 29, 696–703. doi:10.1016/j.wasman.2008.03.026 Khalili, A., Jamshidi, S., Khalesidoust, M., Vesali, M., 2017. Evaluation of sewage sludge for incineration (case study: Arak
wastewater treatment plant). Environmental Energy and Economic Research. 1, 249–258. doi:10.22097/eeer.2017.47251
Liao, C. H., Zhao, J. N., Wang, L., 2012. Introduction and analysis of performance evaluating index of combined heat and
power system at home and abroad. Gas & Heat. 32, 4-9. doi:10.3969/j.issn.1000-4416.2012.01.002 (in Chinese)
297
Lin, X. Q., Yan, M., Dai, A. H., Zhan, M. X., Fu, J. Y., Li, X. D., Chen, T., Lu, S. Y., Alfons, B., Yan, J. H., 2015.
298
Simultaneous suppression of PCDD/F and NOX during municipal solid waste incineration. Chemosphere, 126, 60-66.
299
doi:10.1016/j.chemosphere.2015.02.005
300
Lin, Y. M., Kreuk, M. D., van Loosdrecht, M. C. M., Adin, A., 2010. Characterization of alginate-like exopolysaccharides
301
isolated from aerobic granular sludge in pilot-plant. Water Research, 44, 3355-3364. doi:10.1016/j.watres.2010.03.019
302
Li, S., Li, Y., Lu, Q., Zhu, J., Yao, Y., Bao, S., 2014. Integrated drying and incineration of wet sewage sludge in combined
303
bubbling and circulating fluidized bed units. Waste Management. 34, 2561–2566. doi:10.1016/j.wasman.2014.08.018
304
Lowe, P., 2010. Developments in the thermal drying of sewage sludge. Water & Environment Journal. 9, 306-316.
305 306
doi:10.1111/j.1747-6593.1995.tb00944.x
Mills, N., Pearce, P., Farrow, J., Thorpe, R. B., Kirkby, N. F., 2014. Environmental & economic life cycle assessment of
307
current
308
doi10.1016/j.wasman.2013.08.024
&
future
sewage
sludge
to
energy
technologies.
Waste
Management.
34,
185-195.
309
Murakami, T., Suzuki, Y., Nagasawa, H., Yamamoto, T., Koseki, T., Hirose, H., Okamoto, S., 2009. Combustion
310
characteristics of sewage sludge in an incineration plant for energy recovery. Fuel Processing Technology. 90, 778–783.
311
doi:10.1016/j.fuproc.2009.03.003
312
Novak, J. T., 2006. Dewatering of sewage sludge. Drying Technology. 24, 1257-1262. doi:10.1080/07373930600840419
313
Ometto, F., Quiroga, G., Penika, P., Whitton, R., Jefferson, B., Villa, R., 2014. Impacts of microalgae pre-treatments for
314
improved anaerobic digestion: thermal treatment, thermal hydrolysis, ultrasound and enzymatic hydrolysis. Water
315
Research. 65, 350-361. doi:10.1016/j.watres.2014.07.040
316
Perez-Feito, R., Noguera, D. R., 2006. Recovery of polyhydroxyalkanoate from activated sludge in an enhanced biological
317
phosphorus removal bench-scale reactor. Water Environment Research. 78, 770-775. doi:10.2175/106143006X99803
318
Piao, W., Kim, Y., Kim, H., Kim, M., Kim, C., 2016. Life cycle assessment and economic efficiency analysis of integrated
319
management
320
doi:10.1016/j.jclepro.2015.11.012
321 322
of
wastewater
treatment
plants.
Journal
of
Cleaner
Production.
113,
325–337.
Qiu, Z. H., Dai, H. Y., Kan, A. K., 2007. Energy conservation analysis of gas combined heat and power system (CHP).
Energy research and information. 23, 237-242. doi:10.3969/j.issn.1008-8857.2007.04.010 (in Chinese)
323
Samolada, M. C., Zabaniotou, A. A., 2014. Comparative assessment of municipal sewage sludge incineration, gasification
324
and pyrolysis for a sustainable sludge-to-energy management in Greece. Waste Management. 34, 411-420.
325
doi:10.1016/j.wasman.2013.11.003
326 327 328
Suh, Y., Rousseaux, P., 2002. An LCA of alternative wastewater sludge treatment scenarios. Resource, Conservation,
Recycling. 35, 191-200. doi:10.1016/S0921-3449(01)00120-3
Svoboda, K., Baxter, D., Martinec, J., 2006. Nitrous oxide emissions from waste incineration. Chemical Papers. 60, 78-90.
329 330 331
doi:10.2478/s11696-006-0016-x
Thomas, V. M., Mccreight, C. M., 2008. Relation of chlorine, copper and sulphur to dioxin emission factors. Journal of
Hazardous Materials. 151, 164-170. doi:10.1016/j.jhazmat.2007.05.062
332
Thomsen, M., Seghetta, M., Mikkelsen, M.H., Gyldenkærne, S., Becker, T., Caro, D., Frederiksen, P., 2017. Comparative life
333
cycle assessment of biowaste to resource management systems-A Danish case study. Journal of Cleaner Production. 142,
334
4050–4058. doi:10.1016/j.jclepro.2016.10.034
335 336 337 338
van Loosdrecht, M. C. M., Brdjanovic, D., 2014. Anticipating the next century of wastewater treatment. Science. 344,
1452-1453. doi:10.1126/science.1255183
Wakeman, R. J., 2007. Separation technologies for sludge dewatering. Journal of Hazardous Materials. 144, 614–619.
doi:10.1016/j.jhazmat.2007.01.084
339
Winkler, M.K.H., Bennenbroek, M.H., Horstink, F.H., van Loosdrecht, M.C.M., van de Pol, G.J., 2013. The biodrying
340
concept: An innovative technology creating energy from sewage sludge. Bioresource Technology. 147, 124–129.
341
doi:10.1016/j.biortech.2013.07.138
342 343 344 345
Xu, G. R., Zou, J. L., Li, G. B., 2008. Stabilization of heavy metals in ceramsite made with sewage sludge. Journal of
Hazardous Materials. 152, 56-61. doi:10.1016/j.jhazmat.2007.06.072
Yao, H., Naruse, I., 2005. Control of trace metal emissions by sorbents during sewage sludge combustion. Symposium on
Combustion. 30, 3009-3016. doi:10.1016/j.proci.2004.07.047
346
Yoshida, H., ten Hoeve, M., Christensen, T. H., Bruun, S., Jensen, L.S., Scheutz, C., 2018. Life cycle assessment of sewage
347
sludge management options including long-term impacts after land application. Journal of Cleaner Production. 174,
348
538–547. doi:10.1016/j.jclepro.2017.10.17
349 350
Zhan, M. X., Fu, J. Y., Chen, T., Lin, X. Q., Li, X. D., Yan, J. H., Alfons, B., 2016. Suppression of dioxins by S-N inhibitors
in pilot-scale experiments. Environmental Science and Pollution Research. 23, 63-77. doi:10.1007/s11356-016-6401-2
351
Zhang, Y. A., Gao, D., Chen, T. B., Zheng, G. D., Li, Y. X., 2006. Economical evaluation of different techniques to treatment
352
and
353
doi:10.3969/j.issn.1674-5906.2006.02.007 (in Chinese)
354 355
dispose
sewage
sludge
in
Beijing.
Ecology
and
Environment.
15,
234-238.
Zhou, H., Meng, A., Long, Y., Li, Q., Zhang, Y., 2015. A review of dioxin-related substances during municipal solid waste
incineration. Waste Management. 36, 106-118. doi:10.1016/j.wasman.2014.11.011
Fig. 1. Energy balances of the three compared sludge disposal processes.
Fig. 2. Comparison of energy deficit (a) and I & O costs (b) for the three evaluated sludge disposal processes
Highlights • Landfills and agricultural use of excess sludge is becoming increasingly constrained • Alternative sludge disposal routes must be sought • Anaerobic digestion (AD) is not a final solution for excess sludge • Direct incineration without AD has the lowest energy deficit and I & O costs • Focus on more efficient dewatering instead of AD is recommended