Carbonization of sewage sludge as an adsorbent for organic pollutants

Carbonization of sewage sludge as an adsorbent for organic pollutants

21

Lingjun Kong*,†, Minhua Su*, Kaimin Shih†, Diyun Chen* *Guangdong Provincial Key Laboratory of Radionuclides Pollution Control and Resources, School of Environmental Science and Engineering, Guangzhou University, Guangzhou, China, †Department of Civil Engineering, The University of Hong Kong, Hong Kong SAR, China

1

Introduction

Sludge is a solid byproduct generated from water and the wastewater treatment process. Generally, it is composed of organic compounds, heavy metals, and inorganic compounds. It is classified into sewage sludge, industrial sludge, and water-working sludge (Xu et al., 2015). The characteristics of sludge are different from their sources. Sewage sludge is a mixture of exhausted biomass generated from aerobic and anaerobic microorganisms during municipal sewage treatment and inorganic materials, including sand and chemical reagents. The quantity of sludge production has increased vastly due to the rapid development of urbanization in China. Many toxic substances, including pathogens, organic containments, and heavy metals, are found in sludge (Dong et al., 2013). Public concern regarding sludge disposal has grown greatly in the last decades. Thus, it is a great challenge to manage and treat sludge efficiently. The common options for sludge disposal are landfill, land application, and incineration, but each option has important limitations with increasing environmental and legislative constraints. Although landfill is the most widely used option due to its advantages in operation and low cost, it is not a development strategy because a large amount of land is required and it does not include material recovery (Hadi et al., 2015). In addition, land application as a fertilizer is also limited because pathogens, metals, and low concentrations of antibiotics are found in the sludge. Considering incineration, a considerable reduction in the volume and high-energy output were obtained (Gomez-Pacheco et al., 2012). However, sulfur and nitrogen oxides, dioxins, and heavy metals were emitted from the incineration plants, thus resulting in public unacceptability (Bennett and Knapp, 1982). Therefore, these challenges require sustainable development techniques for sludge disposal. One of the most promising options is resource recycling. Because sludge contains a lot of organic compounds, it could be a potential precursor of carbonaceous adsorbents. Carbonization of waste sludge for biochar has

Industrial and Municipal Sludge. https://doi.org/10.1016/B978-0-12-815907-1.00021-0 © 2019 Elsevier Inc. All rights reserved.

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received much attention in recent years because the sludge could be stabilized and reduced during carbonization. During the carbonization process, the sludge was heated to 500–800°C under an inert atmosphere; the noncarbon elements were decomposed. Also, it leads to the grouping together of the free atoms of elementary carbon as elementary graphitic crystallites. A rudimentary pore structure were observed, meanwhile, many pores were blocked by the tarry substance during carbonization process. To enhance the value of sludge carbon as an efficient adsorbent, activation processes were considered to improve its porosity. Physical activation and chemical activation are widely considered at the present. Therefore, conversion of sludge into adsorbent for removing pollutants from environments is being concerned because adsorption is one of the most concerned decontamination technologies (Smith et al., 2009). In the last decades, organic pollutants have been widely discharged in many industrial processes. Thus, this work addresses the important view of conversion of sludge into adsorbents as a substitute for pollutant removal from aqueous solutions and effluent gases.

2

Materials and method

2.1 Materials The raw sewage sludge was obtained from a sewage treatment plant. It was a byproduct of sewage disposal. Chemical activation reagents, including zinc chloride (ZnCl2), potassium hydroxide (KOH), sodium hydroxide (NaOH), and sulfuric acid, (H2SO4) were analytical grade. Physical activation reagents were CO2 and steam.

2.2 Preparation of sewage sludge derived adsorbents Sludge-derived carbon adsorbents were prepared by a carbonization process coupled with an activation process. The dried sludge in the presence of activation reagents was carbonized by an electrical furnace set at a determined temperature range from 500 to 900°C. During the carbonization process, the atmosphere is free of oxygen. Generally, nitrogen was conducted to sweep the air in the furnace. The carbonized samples were washed by deionized water and an acid solution to remove the activation reagents and residual minerals in the sludge carbon.

2.3 Analytical methods 2.3.1 XRF X-ray fluorescence was applied to analyze the mineral content in the sludge-derived adsorbents. The instrument has a titanium target X-ray tube and a high-resolution detector. The samples were studied in a solid phase after grinding and sieving to use the matrices with similar physical properties (Bandosz and Block, 2006).

Carbonization of sewage sludge as an adsorbent for organic pollutants

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2.3.2 XPS The surface characteristics of the sludge carbon were analyzed by X-ray photoelectron spectroscopy (XPS) using an X-ray photoelectron spectrometer. The binding energies of all core levels were referenced to the C 1s C-C bond at 284.6 eV (Wang et al., 2017).

2.3.3 XRD X-ray diffraction measurements were conducted by the standard powder diffraction procedure. The sludge carbon was ground in the presence of methanol using a small agate mortar. The sludge carbon was ground by hand, ensuring particle sizes between 5 and 10 μm, which prevents line broadening in diffraction peaks. The mixtures were smear mounted onto the zero background quartz window of a Phillips specimen holder and allow to air dry. Samples were analyzed by Cu KR radiation generated in an X-ray diffractometer. A quartz standard slide was run to check for instrument wander and to obtain the accurate location of 2θ peaks.

2.3.4 Nitrogen adsorption-desorption The sorption and desorption of nitrogen was carried out using an auto physical adsorption analyzer. Before the experiments, the samples were dried and outgassed at 120°C for 2 h to constant vacuum (10
6 Torr). From the isotherms, the surface areas (BET method), total pore volume (from the last point of isotherm at relative pressure equal to 0.99), micropore volume, mesopore volume, and pore size distributions were calculated using corresponding adsorption models.

2.3.5 SEM The morphologies of the sludge carbon were observed using a field emission scanning electron microscope (SEM). The element compositions were acquired with an energy dispersive x-ray spectrometer (EDS) (Inca300, Oxford). The element mapping was analyzed by SEM-mapping analysis.

2.3.6 FTIR The FTIR spectra were observed through a MAGNA 560 IR Spectrometer (Nicolet, United States). The number of scans and the scanning resolution were set to 250 cm
1 and 4 cm
1, respectively. A background measurement was made of the air within the sample chamber to correct the subsequent sample spectrum by the FTIR software before analysis. Prior to analysis, it was necessary to mount the samples in a KBr medium. Thus, samples were dried, mixed with KBr (Spectrosol, VWR International Ltd., Poole, UK) at a ratio of ca. 1:150 (w/w), and crushed using an agate pestle and mortar. Approximately 150 mg of the sludge carbon–KBr mixture was loaded into a pneumatic, 13 mm die press (Specac, UK), and 10 metric tons of pressure was applied for 10 min to form a disk. The disk was conducted for further analysis.

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2.3.7 Thermogravimetric analysis Thermalgravimetric analysis was carried out using a TG Instrument thermal analyzer. The instrument settings were of a determined heating rate and a nitrogen atmosphere. For each measurement, about 25 mg of a ground adsorbent sample was used. The gravimetric results were recorded during the heating process.

2.4 Adsorption experiments Batch adsorption experiments were conducted to investigate the adsorption behavior of organic pollutants from aqueous. For each test, a determined sludge carbon was thrown into organic pollutants containing solution. The isotherm adsorption experiments were conducted at a determined temperature controlled by a water bath. Langmuir and Freundlich models were evolved to fit the experimental results. For the kinetic adsorption experiments, the solution was drawn at determined time intervals and the residual concentration of adsorbates was analyzed. Pseudo first-order and pseudo second-order models were conducted to fit the adsorption kinetics. The equilibrium adsorption capacity (qe, mg g
1) and the adsorption capacity at different time t (qt, mg g
1) were calculated as follows: qe ¼

ðC0
Ce ÞV m

(1)

qt ¼

ðC0
Ct ÞV m

(2)

where C0 (mg L
1), Ce (mg L
1), and Ct (mg L
1) represented the initial concentration, the equilibrium concentration, and the concentration at time t, respectively. V (L) represents the volume of the solution, and m (g) represents the adsorbent mass. The adsorption experiments to VOC were conducted in a fixed-bed reactor. The adsorption experimental setup consisted of a VOCs vapor generator, a glass adsorption column filled with adsorbents, and a gas analysis system, as shown in Fig. 1. First, the nitrogen steam from a compressor was divided into two steams using two mass flow controllers. One was conducted to the bubble saturator containing pure liquid pollutant. Another was conducted to adjust the VOC concentration. Then the mixed steams were passed through the adsorption column filled with adsorbents. The VOC concentration was analyzed with a gas chromatograph (GC, Shimadzu GC-2014) equipped with a flame ionization detector (FID). The residual concentration of VOC at the outlet was measured by a GC-FID at determined time intervals. The amount of the adsorbed VOC at equilibrium was calculated as follows: w¼

  ðt F c0
ci dt m 0

(3)

Carbonization of sewage sludge as an adsorbent for organic pollutants

479

Fig. 1 The experimental diagram of dynamic adsorption of toluene (A: N2; B: wave; C: flow meter; D: toluene; E: water; F: adsorbents; G: detector of flow pressure; H: absorption for offgas; I: GC-FID).

where w is the adsorption amount (mg/g); F is the volume flow rate (mL min
1); ci and c0 are the outlet and inlet concentration of adsorbate (g mL
1), respectively; m is the mass of the adsorbent (g); and t is the adsorption equilibrium time (min).

3

Results and discussion

3.1 Effect of activation process on pore characteristics 3.1.1 Physical activation A two-step process was carried out in physical activation. In the first step, the precursor was pyrolyzed at low temperature in the absence of air. The resulted char had low porosity. At the second step, it evolved the selective carbon atom burn-off under controlled gasification atmosphere. In this process, carbon dioxide and steam are the most addressed physical activation reagents. To achieve the burn-off process, the carbon structure is exposed to a carbon dioxide and steam atmosphere at about 700–1100°C. The tarry substance on the carbon surface was partly oxidized, it was served to fabricate pores. In addition, steam activation could fabricate microporosity and mesoporosity. The mechanism could be ascribed to the coupling effect of devolatilization and loss of fixed carbon during the reaction of carbon with steam: C + H2O ! CO + H2. The porosity and BET surface area increased with activation temperatures. However, more particles were burned out with the temperature increased higher than 850°C, leading to the collapse of the pore network. Thus, the BET surface area and porosity declined (Gonza´lez et al., 2009). Carbon dioxide is also active for enhancing the porosity of carbon due to the gasification of the carbon atom in the interior of the particle. It is more competitive than steam activation if the temperature exceeds the maximum temperature of 800°C. A high activation temperature and a long dwell time are required. It is reported that the expansion of opened micropores and the opening of closed micropores were both observed during carbon dioxide activation. The functional groups, including dOH,

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dNH and C]O functionalities, were decreased with increased pyrolysis temperature, indicating a decrease in the acidity of the char ( Jindarom et al., 2007). To further increase the porosity of the adsorbent, a washing process before and after carbonization was conducted to remove the inorganic fraction. HCl washing prior to CO2 activation produced an adsorbent with a BET surface area of 269 m2/g. It is much higher than 7 m2/g for that without washing (Ros et al., 2006). Compared to the results obtained by HCl postwashing, the increase in BET surface area is less effective than that prior to activation (Fitzmorris et al., 2006). The improvement of the BET surface area in the presence of the washing process is ascribed to the reduction of inorganic matter. Thus, it is not hard to understand that HCl washing prior to activation is more effective than postactivation washing because the inorganic fraction could be converted into mineral-like compounds. Therefore, the minerals could be encapsulated into the carbon phase, increasing the stability of the inorganic fraction (Bagreev et al., 2001). Therefore, it is evident that improved washing efficiency for removing the inorganic fraction is necessary. Alvarez et al. (2016) reported that CO2 activation for preparing sewage sludge char could increase its surface area. A new washing process in the presence of Na2CO3 solution was conducted to enhance the silica extraction. The ash content of sludge carbon decreased from 44.7% to 28.0% after washing. Thus, the values of the BET surface area increased from 43 to 235 m2/g and 311 m2/g after postwashing in the presence of HCl and HCl coupled with Na2CO3 solution, respectively.

3.1.2 Chemical activation Chemical activation is another technique for enhancing the BET surface area. Table 1 summarizes the different chemical reagents with various efficiencies on enhancing the BET surface area. Compared to the results in the presence of physical activation, a higher surface area and a lower activation temperature were required by chemical activation. The details in the chemical activation process are discussed as follows.

KOH Potassium hydroxide is one of the most widely concerned activation reagents. It clearly can be seen from Table 1 that potassium hydroxide activation has the most significant effect on enhancing the surface area. The obtained surface area values are different from the activation temperatures, the source of sludge, the mass ratio of potassium hydroxide to sludge. Three kinds of sludge were reported by LilloRo´denas et al. (2008) to prepare activated carbon in the presence of potassium hydroxide. During these processes, the tested sludge was carbonized, and further activated by potassium hydroxide. In the absence of the activation process, the surface areas of three chars were all not higher than 50 m2/g. Once the activation process was presented, the surface area was highly increased to more than 1000 m2/g, but the values were different. It is not hard to understand that the difference is ascribed to the properties of chars. The surface area depends on the ash content of the char. The lower the ash content of char, the higher the surface area is. The activation reaction occurred between the char and the activation reagent. The mass ratio of potassium hydroxide

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481

Table 1 Effect of chemical activation on SBET of sludge-derived adsorbents References Lillo-Ro´denas et al. (2008) Lillo-Ro´denas et al. (2008) Monsalvo et al. (2011) Monsalvo et al. (2011) Hunsom and Autthanit (2013) Kong et al. (2013b) Hunsom and Autthanit (2013) Ros et al. (2007) Ros et al. (2006) Zhang et al. (2005) Chen et al. (2002) Kong et al. (2013a,b) Martin et al. (2003) Zhang et al. (2005) Zhang et al. (2005)

Chemical reagent

Mass ration

Temperature

SBET (m2/g)

Post treatment

KOHa

1:1

700°C

1058

HCl

KOH1

1:1

700°C

1301

HCl

KOH

1:1

750°C

950

HCl

KOH

3:1

750°C

1832

800°C

853

HCl

800°C

853

HCl-CA

K2CO3

800°C

810

HCl

NaOH NaOH ZnCl2

700°C 700°C 650°C

725 689 555

HCl HCl HCl

ZnCl2

500°C

647.4

HCl

ZnCl2

500°C

517.4

CA-HCl

H2SO4

700°C

253

NaOH

H2SO4

650°C

408

NaOH

H3PO4

650°C

289

NaOH

KOH KOH

1:1

a

With different sludge precursor.

to sludge was also a key factor determining the value of the surface area. Although a high ratio of potassium hydroxide resulted in a high surface area, the yield is low in the presence of a high mass ratio. During potassium hydroxide activation, a high temperature of 700°C is required. Generally, carbonization processes was evolved as one step and two steps to prepare the activation carbon. For the one-step process, it is steam from the dehydration that KOH could react directly with organic compounds during carbonization, especially react with oxygen containing functional groups, favoring the development of porosity. It is proposed that dehydration occurred by reacting with KOH and NaOH, leading to the rise in the C:H and C:O ratios (Hsu and Teng, 2000). To remove the residual chemical activation reagent, postwashing was required. The traditional washing solution was HCl solution. However, the surface area depends

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Industrial and Municipal Sludge

on the inorganic compound content in the activated char, so it is important to enhance the efficiency to remove inorganic compounds. Citric acid coupled with the HCl washing process was conducted to enhance the postremoval efficiency (Kong et al., 2013a). Obviously, postwashing could remove the minerals from the sludgederived activated carbon. Only water dissolvable ions such as K+ and Na+ ions were removed by deionized water washing. HCl solution with a pH value of 1 was used to wash SC by protonation besides water dissolution. Parts of water-insoluble ions could be removed in the presence of the HCl solution. However, the removal efficiency is unfavorable. K
, Na
, and parts of Ca
, Mg
, and Mn
containing minerals could be removed by HCl washing while the others could not be removed easily due to the formation of stable minerals such as oxides, phyllosilicates, hydrate, or stable crystals fixed on SC. Interestingly, besides K, Na, Ca, Mg, and Mn, more Fe and Al-containing minerals in SC were removed by CA washing comparing to HCl washing. The enhanced removal efficiencies of Fe, Al, and Mn were ascribed to chelation besides protonation. A mixed solution of HCl-CA with a pH value of 1 was conducted to wash the SC. Results showed that HCl-CA-SC had the highest demineralization efficiency. The removal efficiencies and removal rates of Ca, Mg, Fe, Mn, and Al washed by HClCA were the highest compared to other washing solutions. The promoting demineralization mechanisms of HCl-CA washing could be concluded as: (a) enhancing the dissolution rate far from equilibrium through protonation; and (b) forming soluble metals complexes with organic ligands in the presence of CA. Thus, the resulting DW-SC, HCl-SC, CA-SC, and HCl-CA-SC washed by deionized water, HCl, citric acid, and a mixture solution of HCl-CA had different ash contents that were related to the mineral removal efficiencies. In addition, the values of SBET were in agreement with their ash content, as shown in Table 2. The HCl-CA-SC with the lowest ash content had the highest SBET. Considering the total pore volume and micropore volume after postwashing, it is not hard to see that the micropore volume could be increased but the postwashing process mainly resulted in an increase in the total pore volume. Thus, the HCl-CA-SC had the lowest ratio of Vmicro/VT. All these results indicated that postwashing could influence the mineral content, further enhancing the SBET and pore volume. Table 2 Pore structure parameters of SC, DW-SC, HCl-SC, CA-SC, and HClCA-SC Pore structure Yield (%)

Ash (%)

SBET (m2g21)

Vt (m3g21)

Vmicro (m3g21)

Vmicro/VT

Adsorbents SC DW-SC HCl-SC CA-SC HCl-CA-SC

100 71.26 57.74 45.52 37.46

75 64 57 46 32

116.8 318.5 325.4 385.8 605.2

0.09 0.20 0.25 0.28 0.46

0.05 0.09 0.09 0.10 0.15

0.602 0.443 0.362 0.339 0.316

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NaOH NaOH was another activation reagent. Its effect on generating a high surface area was obvious. The activation mechanism was ascribed to the dehydration. Also, postwashing was required to remove the residual activation reagent. Considering the result of the surface area, the resulted SBET obtained from NaOH activation is not as high as KOH activation. But it is more effective in the activation of coal pitch-derived carbon fibers in the presence of NaOH (Macia´-Agullo´ et al., 2004). More attention may be paid on synthesis of sludge-derived activated carbon in future.

ZnCl2 Zinc chloride activation for preparing activated carbon from paper mill sludge was reported by Khalili et al. (2000). The increase in the mass ratio of ZnCl2 to sludge from 0.75 to 2.5 resulted in a 600% increase in the mesopore volume. A mass ratio of ZnCl2 to sludge less than 1.0 and greater than 1.5 resulted in the production of micro- and mesoporous carbons, respectively. After that, the activation of sewage sludge by ZnCl2 at 650oC was reported for the activation of anaerobically digested and undigested sewage sludge. The undigested sludge with higher carbon content and lower ash content had a higher BET surface area and pore volume (Tay et al., 2001). To further enhance the SBET, a coupling template and ZnCl2 were evolved to prepare sludge-derived porous carbon in our previous study. The fabrication of hierarchical pores was reported in two steps. For the one-step process, 40.0 g sludge was first impregnated into 0.25 mol citric acid solution, being dried and carbonized in a programmable tube electric furnace (SKF-210, Hangzhou Lantian Instrument Co., China) at a rate of 20°C min
1 to 500°C in the presence of N2 and holding for 1 h. The obtained sample named SCCA was further impregnated into 0.5 mol ZnCl2 solution overnight and activated for 1 h at 500°C. After cooling to room temperature, the product was ground to less than 80 mesh, washed with citric acid-hydrochloric acid solution to remove metal ions and then rinsed with deionized water until the pH reached 7 (Kong et al., 2013b). The obtained product was dried at 105°C overnight, and named SCCA/Zn. The bulk density of SCCA was 0.58 kg dm
3 and the bulk density of SCCA/Zn was 0.36 kg dm
3, which were lower than that (1.02 kg dm
3) of SC. In addition, the SBET of SCCA and SCCA/Zn increased to 117.6 and 867.6 m2 g
1, respectively. Obviously, the SBET of SCCA/Zn was higher than that of SC activated by ZnCl2, indicating that the coupling use of the template and activation reagent caused a synergistic effect on fabricating pores. Pores in a wide size distribution can be seen on the surface of SCCA/Zn. Significantly, a lot of micropores were located on the surface of macropores, leading to an appearance of “pores in pores” as obtained by SEM analysis. Fabrication of hierarchical pores can be described as follows: First, CA is decomposed into plenty of CO2 and H2O vapors in pyrolysis. The released CO2 and H2O vapors favored fabricating abundantly hierarchical macropores, providing more interspace and macropores for ZnCl2 activation to fabricate abundant micropores and mesopores, as shown in Fig. 2.

484

Industrial and Municipal Sludge

Fig. 2 The diagram of pore fabrication process coupling CA and zinc chloride activation.

Besides, the one-step process coupling template and activation reagent was conducted to prepare the hierarchical porous carbon. A total of 40.0 g of sludge was impregnated into a 100 ml mixed solution containing 0–2 mol CA and 0.5 mol ZnCl2, then stirred overnight to form homogenous sol-gels. The sol-gels were dried and carbonized, subsequently washed with hydrochloric acid solution to remove metal ions, then rinsed with deionized water until the pH reached 7, and finally dried at 105oC overnight. The product with the maximum BET surface area was named SCCA-Zn. The maximum BET surface area of SCCA-Zn was 792.4 m2 g
1. As being similar to the results described above, the resulting SCCA-Zn possessed hierarchical pores, and lots of well-developed pores with various sizes were observed on the surface of SCCA-Zn. Coupling the template CA and ZnCl2 by one step and two steps could be an efficient method for fabricating hierarchical pores with great surface area. For the sludge char powers, it is difficult to separate them from the aqueous solution in application. The granular sludge chars were easily separated from aqueous solution, which could address the above issues. However, the interior surface of these granular char acted less because adsorption only occurred on the surfaces of granular char. Therefore, the development of a new sludge char with high availability and easy separation was highly concerned. Hollow-like spherical sludge chars were prepared by coating the sludge pastes impregnated with ZnCl2 in different mass ratios onto a cotton ball to form a kind of cotton/sludge sphere. The diagram is described in Fig. 3. The layer of the sludge paste was about 1 mm, and then the sludge-coated spheres were dried at 105oC. Finally, the

Fig. 3 The diagram of preparing hollow-like sludge char sphere.

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485

obtained spheres were put into a ceramic ark and heated in a programmable tube electric furnace at a rate of 20°C min
1 to 500°C in the presence of N2 holding for 2 h (Rozada et al., 2008). After cooling to room temperature, the product was washed with hydrochloric acid solution to remove inorganic compounds and then rinsed with deionized water until the pH value reached 7. The resulting granular char was named HSC. The SBET of HSC was 1008.4m2/g. Generally, the SBET of HSC could be equal to the sum of the SBET of cotton carbon fiber and sludge char in the HSC according to their mass percentage. Considering that the mass percentage of the sludge char and cotton carbon fiber in HSC were 87.8% and 12.2%, respectively, while the SBET of cotton carbon fiber was 2150.6 m2/g, the SBET of HSC can be calculated as 675.1 m2/g. The calculated SBET of HSC is clearly less than the experimental SBET of HSC. The fact shows that hollow-like spheres possessed the advantage of great surface areas. The apparent density of HSC was 0.1008 g/cm3. It was only 15% of that of SC. The low apparent density resulted in low resistant loss in the filling absorbent column, as shown by a pressure drop of 38.8 mm H2O column (Wu et al., 2014). Further, we report a process for preparing hollow spherical sludge carbon via using different sized polystyrene (PS) foam waste as the sacrificial hard templates and as an inner core, and the sludge paste impregnated with ZnCl2 was coated onto the PS surface. The obtained PS@sewage sludge was pyrolyzed. During pyrolyzing, the PS could be decomposed, leaving a hollow structure, and the sludge was converted into carbon, forming a hollow spherical sludge carbon. ZnCl2 acted as the activation reagent to fabricate micropores. Thus, the apparent density of hollow sludge carbon with a shell thickness of 0.2 mm was 0.0905 g/cm3, only yielding 27.0% that of the solid sludge carbon. The low apparent density resulted in a less resistant loss in the adsorbent filling column and a large external surface. This is the advantage of the hollow spherical sludge carbon as described above.

H2SO4 Chemical activation in the presence of H2SO4 was another attempt widely concerned for increasing the SBET of sludge-derived carbon. The SBET of sludge carbon increased from 137 to 408 m2/g while the micropore and total pore volume increased from 0.016 and 0.227 to 0.026 and 0.523 cm3/g, respectively, in the presence of H2SO4 (Zhang et al., 2005). However, the SBET and pore volume of sludge carbon in the presence of H2SO4 were lower than those in the presence of ZnCl2, indicating that ZnCl2 activation was more efficient than H2SO4 activation. Besides, the resulting sludge carbon activated by H2SO4 was washed by NaOH solution. An experimental design methodology was used to optimize the experimental conditions of sorbent preparation from sewage sludge by chemical activation with sulfuric acid at 700°C. The optimal SBET of sludge carbon was 353 m2/g (Rio et al., 2005). Notably, sulfuric acid concentration is not higher than 3 M because destruction of the material structure was observed when the concentration was higher than 3 M during thermal treatment (Rio et al., 2003). Pan et al. (2011) reported that the micropores of the resulting carbon were destroyed and enlarged during H2SO4 activation. Thus, the pores in these absorbents were mainly composed of mesopores and macropores, even supermacropores.

486

Industrial and Municipal Sludge

3.2 Adsorption of organic pollutants from aqueous 3.2.1 Adsorption of dyes from aqueous solution Because the activation reagents and minerals are founded in the sludge-derived adsorbents, postwashing played an important role in demineralization and further enhancing SBET. Thus, the effect of postwashing on the adsorption capacities of sludge-derived adsorbents was investigated. Methylene blue and acid orange II were conducted as adsorbates. The effect of contact time on the adsorption capacities of acid orange II and methylene blue was studied by a kinetic experiment. The initial dye concentration was 500 mg L
1. The pH value of acid orange II and methylene blue in solution was 5.85 and 3.93, respectively, without adjusting and the additive dosage of produced activated carbon was 1.5 g L
1. The adsorption capacities of acid orange II and methylene blue on sludge carbon and demineralized sludge-derived carbons were increased rapidly in the first 30 min, and then slowed down toward equilibrium within 180 min. The maximum adsorption capacities of sludge carbons, deionized water-washed sludge carbon (DW-SC), HCl-washed sludge carbon (HCl-SC), citric acid-washed sludge carbon (CA-SC), and HCl and citric acid mixtures washed carbon (HCl-CA-SC) were 116, 147, 182, 239, and 319 mg g
1, respectively, for acid orange II, and 175, 177, 192, 224, and 250 mg g
1, respectively, for methylene blue, as shown in Fig. 4. They both increased with demineralization efficiency and SBET. This could be explained by the increased specific surface area and the total pore volume, especially for mesopores and macropores, because opening and widening of micropores, mesopores, and interstices channels in sludge carbon by demineralization could provide more adsorption sites for removing acid orange II and methylene blue (Malik et al., 2007; Rodrı´guez et al., 2009). Thus, the demineralization of SC played an important role in increasing their adsorption capacities to acid orange II and methylene blue.

Fig. 4 Comparison of adsorption capacities of AOII and MB onto SC derived adsorbents washed by various washing reagents.

Carbonization of sewage sludge as an adsorbent for organic pollutants

487

Comparing the adsorption capacities of sludge carbonaceous adsorbents, SC and DW-SC had higher adsorption abilities to cationic dye (methylene blue) than to anionic dye (acid orange II) while HCl-CA-SC had higher adsorption abilities to anionic dye than to cationic dye. It can be explained by the fact that SC and DW-SC had a favorable ability to uptake H+ on their surfaces due to their basic surfaces (pHa of 10.17 and 8.12). Then, SC and DW-SC were negatively charged when the pH value was higher than the pHzpc (pH at which the surface has a net zero charge) due to their basic pHa, and therefore SC and DW-SC had higher affinity to cationic dyes than to anionic dyes due to electrostatic attracting interaction between cationic dyes and SC or DW-SC. The surfaces of HCl-SC and CA-SC were changed to neutral and the content of minerals decreased after washing with HCl and CA, then HCl-SC and CA-SC had similar adsorption amounts and adsorption kinetics for methylene blue and acid orange II due to the neutral surfaces and the weakened ion exchange reaction. In turn, the surface of HClCA-SC (pHa ¼ 4.26) was protonated by acid washing and possessed a positive charge below pHzpc by dissociation of OH-, presenting higher adsorption capacities of 319 mg g
1 to acid orange II than 250 mg g
1 to methylene blue. This was consistent with the results of a previous study. Thus, postwashing with HCl-CA was effective for demineralization to enhance pore structures, and further significantly increased the adsorption capacities for the removal of anionic dyes. The isotherm adsorption experimental data was fitted by Langmuir and Freundlich models to evaluate the adsorption characteristics. The correlation coefficients indicated almost perfect agreement with the experimental data fitted by the Langmuir model and the Freundlich model. However, the calculated maximum adsorption capacities of acid orange II and methylene blue from the Langmuir model were larger than the experimental adsorption capacities, suggesting that the isotherm adsorptions were not fitted well by the Langmuir model but were fitted well by the Freundlich model. The filling column was conducted to investigate the adsorption performances of the resulted hollow-like granular sludge derived adsorbent (HSC). Methylene blue was conducted as a pollutant, and granular activated carbon (GAC) and sludge carbon (SC) were evolved as comparative adsorbents to investigate the adsorption behavior, as shown in Fig. 5. Adsorption breakthrough curves obtained under various experimental conditions were presented. The half breakthrough time (when the Ct/C0 was 0.5) of GAC, SC, and HSC was 13.9, 10.6, and 20.2 h, respectively. The calculated adsorption capacity of HSC was 186.42 mg/g. It was about 8.4 and 12.2 times those of GAC and SC, although the SBET of GAC was equal to HSC. Such great disparity between the adsorption capacities of HSC and those of GAC and SC to methylene blue can be contributed to the hollow-like structure of HSC because the hollow-like structure took the advantage in penetrability, as discussed above. First, the Thomas model was described in Eq. (4) to evaluate the adsorption dynamic parameter to further understand the adsorption performance of HSC.   C0 kTh q0 w
kTh C0 t
1 ¼ ln Ct Q

(4)

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Industrial and Municipal Sludge

Fig. 5 Comparison of half breakthrough time and adsorption capacity of MB on GAC, SC, and HSC. Table 3 Dynamic adsorption model parameters of methylene blue on varied adsorbents Thomas model Adsorbents

kTh (mL/min mg)

q0 (mg/g)

R2

GAC SC HSC

1.253 1.227 1.547

22.16 14.67 169.40

0.983 0.979 0.985

where kTh is the Thomas rate constant (mL/h mg), q0 is the equilibrium adsorption capacity (mg/g), and C0 and Ct are the initial concentration and outlet concentration(mg/L) of MB at time t (h), respectively. w is the adsorbent mass (g) and Q is the flow rate (mL/h). The values of kTh and q0 of the tested absorbents obtained by fitting the Thomas model are shown in Table 3. The calculated q0 of GAC, SC, and HSC were 22.16, 14.67, and 169.40 mg/g, respectively. They were all consistent with their experimental values. The fact shows that the adsorption breakthrough curves of HSC to MB could be well described by the Thomas model (Z€ umriye and Ferda, 2004; Theodoridis et al., 2014). It is well known that the Thomas model is suitable for adsorption processes where the external and internal diffusions are absent. Therefore, the fitting result indicated that the adsorption of HSC to MB was not limited by diffusion. The result is consistent with the high surface permeability of hollow-like spherical structures for HSC.

3.2.2 Adsorption of benzene derivatives from aqueous solution Because a kind of sludge carbon with hierarchical pores (SCCA-Zn) was prepared by coupling template and activation reagent, the resulted sludge carbon (SCCA-Zn) was conducted as an adsorbent for removing benzenes in aqueous solution. Fig. 6 presents

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489

Fig. 6 Sorption kinetics of OH-BA, PHEN, BA, and Cl-PHEN on SCCA-Zn.

the sorption amounts of four benzene derivatives on SCCA-Zn at pH 4 as a function of contact time. Obviously, the sorption equilibrium of PHEN and Cl-PHEN was achieved at the first 60 min, but sorption of BA and OH-BA was continued after 60 min. The discrepancy was ascribed to the different sorption mechanism. Sorption isotherms of four benzene derivatives on SCCA-Zn at pH 4 and 25°C are investigated in Fig. 7 to understand the mechanism. Apparent two-step stepwise adsorption curves were observed. First, a rapid nonlinear increase in the low

Fig. 7 Sorption isotherms of four benzene derivatives on SCCA-Zn (Kong et al., 2014).

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Industrial and Municipal Sludge

equilibrium concentration range occurred, following by a slow linear increase in the high equilibrium concentration range. The phenomena were also found by other studies (Mohamed et al., 2011; Chen et al., 2012a,b; Yuan et al., 2004; Chen et al., 2008). A dual reaction domain mode could be conducted to describe the sorption performances quantitatively (Chen et al., 2008): QT ¼ QA + QP

(5)

where QT is the total sorption amount (mmol g
1), QA is the adsorption amount, and QP is the partition sorption amount (mmol g
1). In the range of high solution concentration, the interaction between benzene derivatives and SCCA-Zn was mainly ascribed to partition sorption because the surface adsorption was preferentially saturated (And and Kile, 1998; Cornelissen and Gustafsson, 2004). In this case, the dual reaction domain model (5) can be transformed to the following equation: QT ¼ QA + QP ¼ QA + KP Ce

(6)

where QA is the saturated adsorption capacity, KP is the partition sorption coefficient, and Ce is the solute equilibrium concentration. The sorption isotherms in high concentration can be fitted well with Eq. (6) and the regression parameters are presented in Table 4 (Kong et al., 2014). The fitted results illustrated that partition coefficients (KP) of the tested benzene derivatives to SCCA-Zn followed the order of Cl-PHEN > BA >PHEN > OH-BA. The order was consistent with their octanol-water partition coefficients (log Kow), which indicated the hydrophobic. A higher value of log Kow suggested hydrophobic adsorbates. More important, the values of KP and log Kow were linear related (KP ¼ 0.0213 log Kow
0.018, R2 ¼ 0.942), as shown in Fig. 8. The above results indicated that sorption at a high solute concentration range was ascribed to partition adsorption, which was dependent on hydrophobic interactions between these benzene derivatives and the highly carbonized aromatic matrix in the SCCA-Zn. Except from the partition adsorption, the other factors contributed to sorption should be considered to understand the sorption mechanism. Therefore, it is necessary to investigate the corresponding saturated adsorption (QA). According to Eq. (6), the Table 4 Partition sorption coefficients and saturated adsorption of benzene derivatives on the sludge-derived biochars from aqueous solution Adsorbate

Log Kow

KP(L g

OH-BA PHEN BA Cl-PHEN

1.26 1.46 1.95 2.45

0.010 0.014 0.030 0.036

21

)

QA (mmol g21)

R

QT (mmol g21)

1.831 1.367 2.138 1.672

0.977 0.988 0.998 0.997

2.90 2.01 3.49 3.34

Total sorption amounts at initial concentration of 50 mmol L
1.

a

a 2

Carbonization of sewage sludge as an adsorbent for organic pollutants

491

Fig. 8 Liner fitting of partition coefficient KP plotted as log Kow.

adsorption apart from partition sorption contributed to the whole isotherm can be calculated as Eq. (7): QA ¼ QT
QP ¼ QT
KP Ce

(7)

The saturated adsorption capacities (QA) of the tested benzene derivatives on SCCA-Zn were calculated from Eq. (7). The nonlinear adsorption isotherms excluding partition sorption can be fitted well with the Langmuir equation, as shown in Fig. 9 and Table 5. Obviously, the QA at plateaus (Qmax A ) calculated from the Langmuir equation was nearly similar to the QA calculated from Eq. (7). They followed the order of BA > OH-BA > Cl-PHEN > PHEN, and the order was different from the KP and log Kow. Although pore-filling adsorption was widely known due to its rich pores, the discrepancy of QA suggested that another mechanism could contribute to the sorption of these benzene derivatives besides hydrophobic interaction and pore filling. It was interesting to find that the QA of the benzene derivatives with carboxyl were greater than that of the benzene derivatives without carboxyl, no matter how many the values of their log Kow were. This fact indicated that the carboxyl groups may play an important role in the adsorption of these benzene derivatives on SCCA-Zn. Considering the fact that the elemental carbon and highly carbonized aromatic matrix contribute to the hydrophobicity (Chen et al., 2012a,b), thus, the nonhydrophobic interaction was supposed to drive the adsorption by inorganic ash in SCCA-Zn. To confirm the assumption, the inorganic ash of SCCA-Zn was obtained by burning SCCA-Zn at atmosphere (650oC), and used to adsorb four derivatives benzenes. The sorption capacities of SCCA-Zn ash for benzoic acid, p-hydroxybenzoic acid, p-chlorophenol, and phenol were 0.052, 0.058, 0.025, and 0.023 mmol g
1, respectively. The order of adsorption capacities was not in agreement with their log Kow as well. However, the adsorption capacities to benzoic acid and p-hydroxybenzoic acid with the carboxyl group were obviously greater than those

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Fig. 9 Sorption isotherms of OH-BA, PHEN, BA, and Cl-PHEN onto SCCA-Zn except partition sorption: (A) Langmuir model, (B) Freundlich model, (C) linear fitting of Langmuir model, and (D) linear fitting of Freundlich model. Table 5 Linear regression parameters of sorption isotherms of organic pollutants to SCCA-Zn from aqueous solution Langmuir

Freundlich

Adsorbates

Qmax A

k

R2

n

kf

R2

OH-BA PHEN BA Cl-PHEN

1.868 1.388 2.185 1.993

1.777 0.795 0.846 1.487

0.999 0.999 0.999 0.999

4.016 5.094 4.072 4.184

0.882 0.750 1.042 0.914

0.899 0.885 0.861 0.787

to phenol and p-chlorophenol without the carboxyl group. The preferable adsorption of carboxyl-containing benzene derivatives was ascribed to the inorganic ash in the SCCA-Zn. Because the inorganic ash was mainly composed of SiO2 (82% wt.%), the effect of the amounts of SiO2 on the sorption isotherms of the sludge char should be investigated. However, the SBET for the adsorbents in various SiO2 contents was different;

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493

Fig. 10 BET surface area normalized adsorption capacity (QSA) of the benzene derivatives as a function of SiO2 content.

the SBET normalized adsorption capacities as a function of SiO2 content were investigated in Fig. 10 to further understand the adsorption mechanism. The QSA of benzene derivatives without the carboxyl group changed slightly with SiO2 content, but the QSA of benzene derivatives with the carboxyl group increased obviously with SiO2 content. The positive effect of SiO2 on the adsorption could be ascribed to the important interaction between carboxyl and SiO2 in SCCA-Zn. A quantum chemistry calculation equipped with density functional theory and DMol3 code embedded in the materials studio software was conducted to understand the interaction between carboxyl and SiO2 in SCCA-Zn (Giannakas et al., 2013). The benzoic acid as a representative carboxyl-containing pollutant and one of the crystal faces of SiO2 (100) as an interacting surface were presented in this work to investigate the interaction. Adsorption energy Eb, defined as E (SiO2 (100)+benzoic acid) – [ESiO2 (100) + E(benzoic acid)], was evolved to investigate the interaction, where E(SiO2 (100)+benzoic acid) is the total energy of the benzoic acidadsorbed surface SiO2 (100), ESiO2 (100) is the energy of pristine SiO2 (100), and E(benzoic acid) is the energy of the isolated benzoic acid. The Eb value below 0 indicated exothermic adsorption, leading to stable minimal energy configurations. Four possible interaction modes were observed between benzoic acid and SiO2 (100), as shown in Fig. 11. Comparing the possible adsorption configurations, the Eb calculated from mode (a) and (b) is higher than that calculated from mode (c) and (d), indicating that the high Eb favors adsorption. Clearly, the preferential adsorption between the benzoic acid and SiO2 (100) surface followed the rule: mode ˚ for configuration (a) and mode (b) > mode (c) and mode (d). A SidO1 bond (1.83 A ˚ ˚ for (a) and 1.87 A for configuration (b)) and O3-H hydrogen bond (1.61 A

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Industrial and Municipal Sludge

Fig. 11 The charge distributions of adsorption configurations of benzoic acid on SiO2_quartz_beta (100) surface: Mode (a), Mode (b), Mode (c), Mode (d).

˚ for configuration (b)) between the benzoic acid and the configuration (a) and 1.69 A surface were formed, corresponding to the elongation of the neighboring CdO1 bond ˚ for mode (a) and 1.27 A ˚ for mode (b), both from 1.22 A ˚ of origin molecule, to 1.29 A ˚ ˚ for and the elongation of H-O2 bond from 0.98 to 1.02 A for mode (a) and 1.01 A mode (b). The Si atom on the active surface was two-fold coordination, unlike four-fold coordinated Si in the bulk structure. As shown in Fig. 11, the O atom of benzoic acid could attach Si for forming chemical bond SidO. In addition, it is easy to form a hydrogen bond between the O and H atoms due to the large electronegativity of the O atom. Obviously, the chemical bond was much stronger than the hydrogen bond. Thus, chemisorption and physisorption contributed to the sorption of benzoic acid on the SiO2 (100) surface; the chemisorption configuration with one Si-O bond and a hydrogen bond formed as mode (a) and mode (b) was favorable.

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3.3 Adsorption of volatile organic pollutants The sludge-derived adsorbent was also evolved for removing volatile organic pollutants. SCCA/Zn was first conducted as an adsorbent to investigate the adsorption breakthrough to toluene performed in dry air (0% relative humidity) at 298 K, in which other SC, SCZn, SCCA, and ACF were evolved to investigate the adsorption behavior comparatively. The breakthrough time for SC, SCZn, SCCA, and ACF were 3, 10, 8, and 35 min, respectively, as shown in Fig. 12. Fortunately, the breakthrough time for SCCA/Zn was 42 min, which was much greater than those of other chars. Importantly, it was about 133% longer than the sum of the breakthrough times of SCZn and SCCA, indicating that SCCA/Zn displayed the greatest toluene adsorption ability among the test adsorbents. The linear regression analysis of adsorption breakthrough curves by the YoonNelson model ln[c/(c0
c)] vs. t was performed (Long et al., 2011) to describe the adsorption breakthrough curves, where c0 and c were the inlet and outlet concentration (mg L
1), respectively. The adsorption rate constant (k0 ) was calculated from the slope of the ln[c/(c0
c)] versus time (t) plot. They were 0.065, 0.075, 0.131, 0.183, and 0.203 min
1 for SC, SCZn, SCCA, SCCA/Zn, and ACF, respectively. As expected, SCCA/Zn possessed the fastest adsorption rate. However, it was unexpected that the k0 value of SCCA was greater than that of SCZn, although SCCA had a smaller SBET and Smicro than SCZn. The unexpected result is possibly ascribed to more macropores of SCCA, and these macropores could provide many inner passageways for toluene transportation, which can speed their adsorption. The adsorption capacities of SC, SCCA SCZn, SCCA/Zn, and ACF to toluene calculated from their breakthrough curves were 0.25, 0.32, 0.49, 0.83, and 0.62 g g
1, respectively, as shown in Fig. 12. They followed the order: SCCA/Zn > ACF > SCZn > SCCA > SC. However, the adsorption capacities of SCCA/Zn and ACF are not

Fig. 12 Comparison of adsorption breakthrough time and adsorption capacities of toluene on sludge derived adsorbents with ACF (initial concentration ¼ 70 mg L–1; flow rate ¼ 200 mL min–1; T ¼ 298 K).

496

Industrial and Municipal Sludge

Fig. 13 Adsorption breakthrough curves of the adsorbents for toluene (initial concentration: 70 mg L–1, flow rate: 200 mL min–1, T ¼ 298 K).

consistent with the order of their surface areas. The adsorption capacity of the SCCA/Zn was 34% greater than that of the ACF, although the surface area of SCCA/Zn was 867.6 m2 g
1, which was lower than that of ACF. The divergence denotes that other factors besides pore filling contribute to adsorption. Compared to ACF, the sludge char was not completely carbonized, and it still contained some solid tar oil (Bernardoa et al., 2012; Keiluweit et al., 2010). First, the char was washed by acetone for 24 h to remove the tar oil on the surface of SCCA/Zn (Bernardoa et al., 2012). The breakthrough curve of SCCA/Zn partly removed the tar oil and the pictures of the removed tar and acetone are shown in Fig. 13. The adsorption capacity of the oil-removed SCCA/Zn was 0.52 g g
1. It deceased by 37% compared with the adsorption capacity of SCCA/Zn. Interestingly, the adsorption capacity of tar oil-removed SCCA/Zn was recovered to 0.81 g g
1 once the oil-removed SCCA/Zn was tar oil reloaded by dipping. The adsorption capacity of the reloaded SCCA/Zn was restored to 96%, confirming that the adsorption of toluene on SCCA/Zn was solid tar oil dependent. Because the tar oil contains large amounts of multiple aromatic hydrocarbons and a few types of aliphatics (Sa´nchez et al., 2009), the surface of SCCA/Zn is hydrophobic. The surface of SCCA/Zn possessed a strong hydrophobic property, being confirmed by contact angles of 136.5 degrees, being greater than the contact angles of ACF (76.0 degrees), as shown in Fig. 14. The favorable adsorption of toluene on SCCA/Zn is partly dependent on the tar oil, based on the principle of similarity and intermiscibility (Bernardoa et al., 2012; Li et al., 2010a,b; Pan and Xing, 2008; Sa´nchez et al., 2009). Thus, the adsorption of toluene on SCCA/Zn could be ascribed to not only pore adsorption but also hydrophobic partition adsorption. The solid tar oil not only locates in these hierarchical pores but also covers the nonmicroporous surfaces. Besides, the sludge-derived carbonaceous adsorbent was prepared from pyrolyzed sewage sludge, following by acid washing and alkaline hydroxide activation for VOC

Carbonization of sewage sludge as an adsorbent for organic pollutants

497

Fig. 14 Water drop shapes on ACF and SCCA/Zn.

removal. Air streams with low concentrations (<100 ppm, v/v) of toluene, methyl ethyl ketone, and limonene were used in dynamic adsorption/desorption experiments. Adsorption capacities to toluene, methyl ethyl ketone, and limonene were 350, 220, and 640 mg per g of alkaline-activated sludge, respectively. The adsorption efficiency and rate were favorable due to the presence of mesopores for sludge-based adsorbents. The sludge adsorbent favored the adsorption to polar VOCs such as methyl ethyl ketone. Adsorption of toluene and limonene is controlled by the size of the molecules and their ability to access the inner (micro) pores. Toluene adsorption was found to be dependent on the narrow micropore (<0.7 nm) volumes while limonene adsorption was determined by the total pore volume. The VOC adsorption process is not fully reversible.

4

Conclusion

Sewage sludge could be converted into carbonaceous adsorbents by carbonization and activation. Acid washing was evolved to remove the inorganic compounds, and further to increase the SBET. Physical activation and chemical activation were widely paid more attention. Carbon dioxide and steam were the most widely concerned physical activation reagents. Potassium hydroxide, sodium hydroxide, and zinc chloride were the highly concerned chemical activation reagents. Activation by potassium hydroxide and sodium hydroxide was usually conducted at about 800°C while activation by zinc chloride was carbonized at about 500°C. The resulting sludge-derived adsorbents were considered to be efficient adsorbents for removing organic pollutants from aqueous and atmosphere. Physical adsorption driven by pore filling and chemisorption being ascribed to complexation were the main adsorption mechanisms. The high mineral content in the sludge-derived adsorbent led to the SBET, further limiting the wide application of sludge-derived adsorbents in pollutant decontamination. Therefore, the sludge could be converted into carbonaceous adsorbents in the view of resource recovery by carbonization, but it is a challenge to the wide application of sludge resource utilization as adsorbents.

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Acknowledgments This research was supported by the Nature Science Foundations of China (Nos. 51508116, 21407155), the Nature Science Foundations of Guangdong Province (2016A030310265), the Science and Technology Research Programs of Guangzhou City (201607010311), the Project of Guangdong Provincial Key Laboratory of radioactive contamination control and resources (2017B030314182), and the Hong Kong Scholarship (XJ2016037).

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Further reading Linares-Solano, A., 2006. High surface areamaterials prepared from sewage sludge-based precursors. Chemosphere 65 (1), 132–140. Ros, A., Lillo-Rodenas, M.A., Fuente, E., Montes-Moran, M.A., Martin, M.J., Smith, K.M., Fowler, G.D., Pullket, S., Graham, N.J.D., 2009. Sewage sludge-based adsorbents: a review of their production, properties and use in water treatment applications. Water Res. 43 (10), 2569–2594.