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Sludge carbonization and activation: From hazardous waste to functional materials for water treatment
Accepted Manuscript Title: Sludge carbonization and activation: from hazardous waste to functional materials for water treatment Author: Fangwei Cheng Hongxi Luo Lei Hu Bin Yu Zhen Luo Maria Fidalgo PII: DOI: Reference:
S2213-3437(16)30405-5 http://dx.doi.org/doi:10.1016/j.jece.2016.11.013 JECE 1326
To appear in: Received date: Revised date: Accepted date:
25-8-2016 7-11-2016 9-11-2016
Please cite this article as: Fangwei Cheng, Hongxi Luo, Lei Hu, Bin Yu, Zhen Luo, Maria Fidalgo, Sludge carbonization and activation: from hazardous waste to functional materials for water treatment, Journal of Environmental Chemical Engineering http://dx.doi.org/10.1016/j.jece.2016.11.013 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. 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.
Sludge carbonization and activation: from hazardous waste to functional materials for water treatment Fangwei Cheng 1, Hongxi Luo 1, Lei Hu 2, Bin Yu 2, Zhen Luo 2, Maria Fidalgo3 * 1. Department of Chemical Engineering, University of Missouri-Columbia, Columbia, MO 65211, United States 2. CECEP Both Environment Eng. & Tech. Co., Ltd, Wuhan, Hubei 430071, China 3. Department of Civil & Environmental Engineering, University of Missouri-Columbia, Columbia, MO 65211, United States KEYWORDS: Sustainability, Sludge, Carbonization, Water treatment, Adsorption, Sorbents
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ABSTRACT: The utilization of sludge from wastewater treatment plants as adsorbent material was investigated. Dry sludge from Ezhou Qingyuan sewage treatment plant (Hubei, China) was heated in anaerobic conditions to produce carbonized sludge, that further chemically activated at higher temperatures and K2CO3 to enhance porosity and surface area. The materials were characterized by scanning electron microscopy (SEM), Fourier transformed infrared spectroscopy (FTIR). thermo gravimetric analysis (TGA) and nitrogen adsorption isotherms. TGA curves showed water and low molecular weight organics were lost in a first stage, with the onset of decomposition at 300℃ and up to 700℃; activation resulted in further carbonization. Nitrogen adsorption experiments yielded Type IV isotherms, characteristic of mesoporous materials. Activation greatly increased surface area, reaching up to 642 m2/g. FTIR spectra showed the formation of a carboxyl-metal complex at activation, but no further changes in functional groups with increasing reaction temperature. The adsorption capacities of carbonized and activated sludge towards Rhodamine B were investigated by batch and kinetic experiments. Adsorption increased with activation temperature, reaching a maximum at 700 ℃ with the exception of the sample carbonized at 500 ℃ where a monotonic increase in capacity was observed. Isotherms showed Langmuir-type behavior; kinetic data was successfully fitted to a pseudo second order model. The adsorption was not affected by pH changes or dissolved solids type and concentration. Zeta potential determinations showed minimal variation of surface charge in the pH range of interest. The results indicated that sludge carbonization is a promising sustainable technology for mass sludge treatment.
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INTRODUCTION The disposal of sludge from wastewater treatment plants is a major environmental concern worlwide1. It contains huge amounts of hazardous pollutants and harmful bacteria, which constitutes a serious threat to the environment and public health when treated ineffectively.1-4 In China, about 36 million tons of wet sludge (80% moisture) are produced from sewage treatment plants every year, and nearly 60% of it is not treated sufficiently.5-7 Traditionally, direct landfilling is the main method of sludge disposal.8-9 However, direct landfilling takes up arable land and increases the possibility of groundwater pollution.8, 10 Alternatively, other methods such as drying and incineration are also employed to stabilize sludge and decrease its volume.11-14 The amount of sludge is dramatically reduced, but the residue requires further disposal, and landfilling is still necessary. Ash from incineration is another type of hazardous waste and needs to be stabilized using concrete. These strategies are not sustainable, and more environmentally conscious approaches are essential for sludge treatment.8 Recently, energy recovery and sludge utilization have become a new trend in sludge treatment, which includes the sludge carbonization approach.15-17 Sludge carbonization is a process where sludge is heated up under anaerobic conditions and transformed into odorless carbon-containing particles.18 In the late 1990s, sludge carbonization techniques were developed by Japanese companies under the regulations of the Japan Sewage Works Association for emission reduction.19 By using carbonized sludge as an alternative fuel, emissions were reduced effectively by 23.5–35.8% compared with direct incineration.19-20 Hence, sludge carbonization has been industrialized, and carbonized sludge is now mainly used as an alternative fuel for power generation in Japan.21
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Compared to raw sludge, carbonized sludge is a better platform to derive sludge based materials since the composition of carbonized sludge is less complicated. Meanwhile, different modification methods can be chosen for carbonized sludge based on its carbon content as well as the target application. Carbonized sludge can be activated to produce activated carbonized sludge with high porosity and surface area.18 Low-cost, tunable properties, sustainability, and the possibility for mass production make carbonized sludge and activated carbonized sludge promising candidates in fields like air purification and water treatment.22 The combined sewage system in China, where municipal wastewater is collected with storm wastewater, introduces additional challenges for sludge treatment. The first large scale industrial sludge carbonization system in China was set up by CECEP Both Environment Eng. & Tech. Co., Ltd in 2015, with a capacity of 60 ton dry sludge/day. The carbonized was successfully applied in soil amelioration.23 However, the specific surface area of carbonized sludge produced was found to be too low to be directly efficiently used as adsorbent material. Compared with the sludge used in some previous works24, the sludge produced in the plant had a lower organic content (30% ~ 40%)25 and reflects the general situation in China’s wastewater treatment plants. Low organic content in sludge brought in challenges in preparation of activated carbon from sludge. As installed capacity for sludge carbonization increases, there is a need to investigate routes for large scale sludge carbonization-activation systems in order to guide the industry level sludge carbonization/activation in China and worldwide. The aim is to advance the sustainability of the wastewater treatment sector providing a safe disposal method for a waste via the conversion into valuable material. We report in this work a general strategy for large-scale production of activated carbonized sludge from carbonized sludge and apply them in liquid-phase adsorption. The major objectives
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of this study are (i) to investigate the effects of carbonization temperature on the thermal properties of carbonized sludge and find the optimal carbonization temperature for sequential activation, (ii) to clarify the influence of activation temperature on the specific surface area and pore structures of activated carbonized sludge, and (iii) to evaluate the performance of both carbonized sludge and activated carbonized sludge in liquid-phase Rhodamine B (RhB) removal and to study the interaction between their adsorption differences and porosity. To achieve these objectives, different carbonization temperatures were used to prepare carbonized sludge. For each carbonization temperature, the corresponding carbonized sludge was activated at a different activation temperature to obtain activated samples. Liquid-phase adsorption of RhB was then studied for all samples to assess their performance and collect data for adsorption isotherm and adsorption kinetics. In addition, different conditions such as pH and ionic strength were tested to determine their effects on adsorption. This work presents an approach for massive, sustainable, and low-cost sludge treatment with a focus on application of its products. EXPERIMENTAL METHODS Materials. Sludge was obtained from Ezhou Qingyuan sewage treatment plant (Hubei, China). Ezhou Qingyuan water treatment plant receives storm and domestic wastewater (~95%) and a smaller fraction of pre-treated wastewater from a nearby iron processing plant (5%). The facility includes a 60 t/d sludge carbonization unit on-site. Prominent properties of the raw sludge and the dry sludge used in this work and corresponding wastewater treated by the plant were summarized in Table 1. Table 1. Prominent properties of the wastewater, raw sludge and the dry sludge.25
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COD
pH Wastewater
BOD5 Total P NH3-N Total Cr Cr(VI) Total Cd
(mg/L) (mg/L) (mg/L) (mg/L) (mg/L) (mg/L) (mg/L)
7.42
52.6
15.9
0.546
6.999
0.012
H2O Organic Ash NH3-N COD SBET (wt%) (wt%) (wt%) (mg/L) (mg/L) (m2/g)
0.009
0.001
pH
Raw sludge
83
5.2
11.8
N/A
N/A
N/A
N/A
Raw sludge_1_HS26
82.4
6.6
11
N/A
N/A
N/A
N/A
Raw sludge_2_TXL26
83
6.8
11.2
N/A
N/A
N/A
N/A
Dry sludge
<1
30.3
68.9
27.17
254.48
9.13
6.96
Cd
Hg
Pb
Cr
As
Ni
Zn
Cu
(mg/kg) (mg/kg) (mg/kg) (mg/kg) (mg/kg) (mg/kg) (mg/kg) (mg/kg) Dry sludge
1.68
0.915
41.6
105
23.9
68.2
731
232
Standard*
20
15
1000
1000
75
200
4000
1500
(*CJ/T 248-2007: The disposal of sludge from municipal wastewater treatment plant – The quality of sludge used for afforestation in gardens or forests; CJ/T: Voluntary Standards issued by the Ministry of Housing and Urban-Rural Development of China.) The physicochemical properties of the sludge used in this work were similar to those from Huangshi WWTP (at Huangshi) and Tangxun Lake WWTP (at Wuhan) in the Hubei Province, also listed for comparison in Table 1 as Raw sludge 1 HS and 2 TXL, and thus, it can be considered representative of large WWTPs in China. Potassium carbonate (K2CO3), hydrochloric acid (HCl), Rhodamine B, hydrochloric acid (HCl), sodium hydroxide (NaOH), potassium nitrate (KNO3), potassium chloride (KCl), potassium sulfate (K2SO4), and potassium dihydrogen phosphate (KH2PO4) were purchased from Sinopharm Chemical Reagent Co,.Ltd. All chemical reagents were of analytical grade and
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directly used as received without further purification. Ultrapure water (18.2 MΩ·cm) was obtained in the laboratory and used for the preparation of all solutions. Preparation of carbonized sludge and activated carbonized sludge. The synthesis of carbonized sludge and activated carbonized sludge is shown in Figure 1. The process started at the water treatment plant, with the operations limited by the red line. Firstly, sludge from a secondary settling tank was sent to a storage tank and then pumped into a decanter via a sludge feed pump. Flocculants were introduced at the end of the sludge feed pipeline. Most of the water content in the sludge was removed in the decanter, and further treated at 100 ℃ for 12 h to obtain absolute dry sludge, which was used in the synthesis of the carbonized sludge and activated carbonized sludge together with the other chemical reagents.27-28
Figure 1. Preparation procedure of carbonized sludge and activated carbonized sludge; wastewater treatment plant processes are marked in red; carbonization in blue and activation process in green.
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The dry sludge obtained in the plant was taken to the laboratory to continue with the carbonization process, marked in blue in Figure 1. In a typical experiment, 12 g of absolute dry sludge was added to a quartz crucible without pretreatment, which was then placed in the middle of a tube oven. The oven was heated to 500 ℃ at a constant rate of 10 ℃/min and kept at the final temperature for 1 h.22 Nitrogen was used to provide an inert atmosphere at a flowrate of 60 mL/min. At the end of the carbonization process, the carbonized sludge was collected once the oven was cooled down to room temperature and stored for further use. In order to obtain the activated material, a 2 g amount of the as-synthesized carbonized sludge was mixed and ground with 4 g of K2CO3 as activation agent in a mortar, as schematically depicted in Figure 1, delineated in green. The mixture was added to a quartz crucible in an oven, heated to 600 ℃ at a constant rate of 10 ℃/min and kept at the final temperature for 1 h. Nitrogen at 60 mL/min was again used as a protecting gas during the thermal treatment. After the oven cooled down, the activated carbonized sludge was collected and washed 3 times with 1 mol/L HCl and ultrapure water, followed by drying at 80 ℃ overnight. A series of control experiments was done to investigate the carbonization temperatures and activation temperatures. The products were labeled as C_X for carbonized sludge samples and C_X_Y for activated carbonized sludge samples, where X and Y represent the corresponding carbonization temperature and activation temperature, respectively (e.g., C_500 is carbonized at 500 ℃, while C_500_600 is carbonized at 500 ℃ and activated at 600 ℃). Material characterization. The thermal properties of carbonized and activated carbonized sludge were studied by a thermogravimetric analysis (TGA) instrument (ZCT-1, JYGK, heating rate = 20 ℃/min). The specific surface area and pore volume of all samples were measured by
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isothermal nitrogen adsorption-desorption analysis (V-sorb 2800P, GAPP). The morphology of the activated carbonized sludge was studied by Scanning Electron Microscopy (SEM) (JSMIT300, JEOL). Surface functional groups of the activated carbonized sludge were determined by Fourier transform infrared spectroscopy (FTIR) analysis (Nicolet 6700, Thermo Scientific), while surface charge was investigated by zeta potential measurements using a zeta potential analyzer (Zetasizer Nano ZS, Malvern). Adsorption performance study. Batch adsorption tests. To evaluate the adsorption performance of the prepared sludges, RhB dye was chosen to represent organic pollutants encountered in dyeing industry in an adsorption test. In a typical experiment, 25 mg of treated (carbonized/activated carbonized) sludge was added to a 100 mL flask containing 25 mL 100 mg/L RhB solution. The suspension was left to equilibrate for 40 min under continuous stirring (600 rpm) at room temperature (25 ℃). After the test, the treated sludge was separated from the solution by centrifugation at 6000 rpm for 10 min, and the concentration of the RhB solution was measured by UV-vis absorbance, using a spectrophotometer (UV-2800, Unico). The removal rate was measured by Eqn. 1: h r
where
and
o
1
(Eqn. 1)
represent the initial RhB concentration and that at time t, respectively.29
Adsorption isotherm studies. Adsorption isotherms were generated for the condition of highest % RhB removal (as determined by the batch adsorption tests), at room temperature (25 ℃) by adding different amounts of treated sludge (2.5, 5, 7.5, 10, 12.5, 15, 20, 25, 30, and 40 mg) to several 100-mL flasks each containing 25 mL of 100 mg/L RhB solution. The pH of the solution
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was 7 and the ionic strength fixed at 300 mmol/L by addition of KNO3. Separation, RhB concentration determination and percentage removal calculations were performed as described above. Adsorption kinetics. The kinetics of the adsorption process was studied for the treated sludge of highest % RhB removal. In this experiment, 200 mg of the treated sludge were added to 200 mL of a 200 mg/L RhB solution. The pH of the solution was 7 and the ionic strength was controlled by KNO3 (300 mmol/L). The adsorption test was conducted for 40 min under continuous stirring (600 rpm) at room temperature (25 ℃), with samples of the solution taken every 8 min to determine their concentrations by UV-vis spectrophotometry. The stability of the adsorbent was determined by performing the kinetics tests for three consecutive cycles, both with and without regeneration. When regeneration was incorporated, treated sludge was separated after each adsorption cycle and heated to 300 ℃ for 30 min under nitrogen. Effects of pH, ionic strength and anions. The effect of pH on adsorption was studied for the sample with the highest % RhB removal adjusting the pH of the RhB solution between 4 and 10 by addition of 0.1 M HCl or NaOH. Variable concentrations of inorganic salts were added to the solution to investigate the influences of ionic strength on removal capacity.
RESULTS AND DISCUSSION Thermal properties of sludge and carbonized sludge. The sludge carbonization process was analyzed by TGA, as shown in Figure 2. Dry sludge was heated from 30 to 800 ℃ with a programed temperature rise in an anaerobic atmosphere with nitrogen as a protecting gas.
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Figure 2. Thermogravimetric analysis of sludge carbonization process under inert atmosphere. The curve can be divided into a low temperature region (region I) and a high temperature region (region II). The volatilization of moisture and some compounds of small organic molecules took place in region I with temperature below 300 ℃, and the corresponding weight loss was 16%. The volatilization and decomposition of higher molecular weight organic molecules, such as tar oil, began once the temperature reached 300 ℃,30 and resulted in a weight loss of 22%. As a result, the brown sludge was carbonized and turned black. The weight loss increased an additional 2% when the temperature was raised from 700 to 800 ℃, indicating that further carbonization occurred at those temperatures. The carbonization temperature for sludge processing in industry generally varies from 400 to 600 ℃; the product obtained at 400 ℃ being
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used for alternative fuel because of the relatively high organic content.30 To study the possibility of transforming carbonized sludge into an activated carbon adsorbent, the highest final temperature in this study was set to 700 ℃. Carbonized sludge samples were produced with final temperatures of 500, 600, and 700 ℃ (C_500, C_600, and C_700). The carbonized sludge contains fixed carbon and inorganic compounds. The fixed carbon content is expected to vary, since the components of the product are temperature dependent. As sludge fixed carbon content is key for its application as absorbent, it was investigated by TGA with an air supply. Based on the TGA (Figure 2), the ratio of fixed carbon to inorganic inert compounds was determined to be around 1:3. The weight percentage of fixed carbon in the carbonized sludge was evaluated as 25% for sludge carbonized at 500 ℃. The carbon content in the carbonized sludge was significantly lower than that in carbon-based materials derived from other waste precursors31-32. The results are shown in Figure 3.
Figure 3. TGA of carbonized sludge with air supply. (a) TGA curves and (b) residuals carbon content.
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As shown in Figure 3 (a), the TGA curves for samples C_500, C_600, and C_700 had similar shapes with a sharp weight loss between 300 and 550 ℃. However, with increasing carbonization temperature, the weight loss interval became narrower, and its boundaries shifted from 300 and 550 ℃ to 350 and 500 ℃. The shift in the lower boundary to the high temperature region could be attributed to the lower amount of organic residuals in the samples produced at higher carbonization temperature. Considering the constant heating rate during the analysis, the lower carbon content required less time for complete combustion and therefore shifted the upper boundary to the low temperature region. The residual carbon content decreased with carbonization temperatures, and it was determined to be 23.8%, 18.4%, and 17.3% for C_500, C_600, and C_700, respectively as shown Figure 3 (b). The results demonstrated that an increase in the carbonization temperature and residence time lowered the carbon content, indicating that more carbonization occurred. The carbon content for C_700 was only 1% lower than that for C_600, which suggests that the process was almost completed at 700 ℃. The decrease in carbon content is consistent with the hypothesis explaining the shift in the upper boundary of TGA curve to the low temperature region. Surface area and morphology. To study the influence of carbonization temperature and activation temperature on the absorption ability of samples, the surface area and pore size distribution were investigated by N2 adsorption tests at 77 K. Samples tested included carbonized sludge before and after activation with K2CO3 at temperatures of 600 ℃, 700 ℃,and 800 ℃. The BET surface are (SBET), micropore volume (Vmicro), and meso pore volume (Vmeso) are shown in Table 2.
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Table 2. Textural parameters derived from the N2 adsorption data at 77 K; BET surface are (SBET), micropore volume (Vmicro), and meso pore volume (Vmeso). Sample Code SBET
Vmicro Vmeso
Vtotal
(m2/g) (cm3/g) (cm3/g) (cm3/g) C_500
24
0.01
0.12
0.13
C_500_600
119
0.05
0.23
0.28
C_500_700
642
0.25
0.55
0.8
C_500_800
390
0.15
0.68
0.83
C_600
37
0.02
0.09
0.11
C_600_600
145
0.06
0.27
0.33
C_600_700
389
0.15
0.52
0.67
C_600_800
375
0.12
0.7
0.82
C_700
45
0.02
0.16
0.18
C_700_600
60
0.04
0.19
0.23
C_700_700
271
0.11
0.4
0.51
C_700_800
249
0.09
0.56
0.65
For the different series of carbonized sludge, higher carbonization temperature resulted in higher surface area of the carbonized sludge.33-34 The surface areas of carbonized sludge were 24 m2/g, 37 m2/g, and 45 m2/g for C_500, C_600, and C_700, respectively. The surface area was 119 m2/g for C_500_600 and 642 m2/g for C_500_700. When the activation temperature was raised from 600 to 700 ℃, the corresponding surface area increased remarkably. However, a further increase in temperature from 700 to 800 ℃ resulted in a decrease in surface area, which can be explained by samples losing part of their microporous structures at high temperature. Similar results were observed for both the C_600 and C_700 series, where 700 ℃ was again the
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activation temperature corresponding to the largest surface area. However, the surface area for activation products processed at 700 ℃ decreased as the carbonization temperature increased from 500 to 700 ℃. Since lower residual carbon content was found for carbonized sludge produced at higher temperature, less carbon was available to react with the activation agent to develop additional porosity and resulted in lower specific surface areas. The volume of micropores exhibited the same trend as the surface area, and the product activated at 700 ℃ had the largest Vmicro for each series.35 This could be explained by the reaction between carbon and K2CO3 being slow, thus developing less microporosity at the relatively low temperature of 600 ℃. On the other hand, a rapid reaction at 800 ℃ resulted in the transformation of micropores into mesopores, reducing the microporal volume and the surface area. The volume of mesopores and total pore volumes monotonically increased with the activation temperature for each series. To investigate the relation between pore structure and adsorption capability, the micropore size distribution and adsorption-desorption isotherms were analyzed. The pore size distribution of the C_500 series is presented in Figure 4 (a): narrow peaks with high intensity were found for each sample between 0.5 and 0.7 nm. At higher carbonization temperatures, the peaks became lower and wider, indicating the loss of well-defined micropore structures. Adsorption-desorption isotherms are shown in the inset graphs: all isotherms were of Type IV according to the IUPAC classification. A sharp increase in the low relative pressure region revealed the existence of micropores; hence, the activated sample could be classified as a microporous material hybrid with mesopores, which is consistent with the textural parameters in Table 2. Similar results were observed for both the C_600 and C_700 series in Figures 4 (b) and (c).
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Figure 4. Micropore size distributions: (a) C_500 series, (b) C_600 series, and (c) C_700 series. Inset: corresponding nitrogen adsorption and desorption isotherms at 77 K. The morphology of activated carbonized sludge C_500_800 was characterized by SEM and corresponding images at different magnifications are shown in Figure 5.
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Figure 5. SEM images of C_500_800; scale bars (a) 1µm and (b) 10 µm. The carbonized sludge was composed of porous particles with average size of approximately 50 µm and exhibited high surface roughness with features in the nanometer scale. An irregular surface morphology is favorable for adsorption, as it results in a large fraction of potential readily available sites to the adsorbate molecule, as opposed to those present in the inner porosity that require diffusion of the contaminant within the pore tortuosity to reach them. The inner pores of the carbonized sludge could be accessed through the pore openings observed in the surface, as the one shown in Figure 5 (a) which was approximately 2 m in diameter. Overall, the SEM images of the C_500_800 exhibited well-developed surface pore structures, suggesting an internal sponge-like structure. Moreover, the relatively large particle size (hundreds of micrometers) introduced significant advantages in adsorbent separation. FT-IR analysis. FT-IR analysis was carried out for C_500 series to further investigate the effects of activation temperature on surface functional groups; the results are presented in Figure 6.
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Figure 6. FT-IR spectra for C_500 series. Five major peaks and absorption bands were found at 3415 cm-1, 1618 cm-1, 1033 cm-1, 780 cm-1 and 471 cm-1 for carbonized sludge C_500. The strong absorption band at 3441 cm-1 can be attributed to the stretching vibration mode of the O-H bond in surface hydroxyl groups or adsorbed water molecules. The peak at 1618 cm-1 can be attributed to the stretching vibration mode of C=O bond or C=C bond. Additionally, the bending vibration of H2O adsorbed on the adsorbent surface may also contribute to the absorption here. A very intensive absorption was observed at 1033 cm-1, indicating the presence of Si-O-Si or Si-O-C bond.36 The doublet at 797 cm-1 and 778 cm-1 corresponded to the Si-O symmetric stretching mode and was the characteristic peak of -quartz.36 The sharp peak at 471 cm-1 was attribute to the asymmetric bending of Si-O-Si bond.36 A small peak found at 2225 cm-1 corresponded to the stretching
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vibration mode of C≡N or C≡N bond. Another small peak was found at 694 cm-1 and was attributed to the symmetric bending vibration mode of Si-O bond within SiO4 group. Activation with K2CO3 at 600 ℃ introduced a major peak at 2080 cm-1 to the spectrum of C_500_600, that increased with activation temperature. This peak may indicate the existence of cumulated double bonds or carboxyl metal complex (C=O--M); similar results were reported by others using K2CO3 for chemical activation of sludge.37 Another peak at 1397 cm-1 was also observed for C_500_600, which corresponded to the bending vibration mode of C-H bond. When activation temperature was 700 ℃, the doublet at 797 cm-1 and 778 cm-1 was replaced by a weak peak at 775 cm-1. Such transformation might indicate the loss of -quartz in the adsorbent at elevated temperature. Meanwhile, no peak was found around 2200 cm-1 for C_500_700. Similar spectrum was obtained for C_500_800 except for the occurrence of a new peak at 596 cm-1. The Si-O bond was found in all sample spectra and demonstrated the formation of silicate, which was in good accordance with the low organic content in sludge determined from TGA. The lack of abundance of surface groups was also related to the low organic content in the raw sludge. Carboxyl metal complex was found for all activated samples and loss of -quartz in the samples occurred at activation temperature higher than 700 ℃. No major change in surface functional groups was observed when activation temperature was further raised to 800 ℃. Hence, activation and its temperature only slightly altered the composition of the treated sludge and only had minor effects on their surface groups. Batch adsorption studies. The absorption capacity of the carbonized sludge towards RhB was investigated. Figure 7 shows the percentage removal of RhB achieved by each treated sludge sample.
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Figure 7. RhB removal ratio (%) for (a) C_500 series, (b) C_600 series, and (c) C_700 series. As shown in Figure 7 (a), the RhB removal of the activated sample of the C_500 series was 14.59% for C_500, 60.38% for C_500_600, 92.73% for C_500_700, and 96.85% for C_500_800. Comparison of C_500 and C_500_600 performance showed that activation enhanced the RhB removal ability by more than 40% since more surface area was available for adsorption. When the activation temperature was increased from 600 to 700 ℃, the RhB removal ratio was increased an additional 30%. Activation at 800 ℃ led to the highest RhB removal of 96.85%. Figure 7 (b) shows that the RhB removal for C_600, C_600_600, C_600_700, and C_600_800 was 11.45%, 60.87%, 95.78%, and 78.80%, respectively. Similar to the C_500 series, when the activation temperature was increased from 600 ℃ to 700 ℃, the RhB removal
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increased by 35%. However, further increases in activation temperature negatively affected adsorption performance. The C_700 series followed exactly the same trend as the C_600 series, as illustrated in Figure 7 (c). The RhB removal for each sample was 14.23%, 59.58%, 82.21%, and 71.34%. Overall, the RhB removal for C_500, C_600, and C_700 samples was much lower than that obtained after activation, which underscores the key role of this process in the reuse of waste sludge as an adsorbent material. Adsorbents carbonized at different temperatures showed similar removal performance (about 60%) when activated at 600 ℃. When 700 ℃ was used for activation, the adsorbent carbonized at 600 ℃ gave the best RhB removal, followed by those carbonized at 500 and 700 ℃. With activation at 800 ℃, an increase in performance was observed for only the adsorbent carbonized at 500 ℃, while a decrease in RhB removal was found for adsorbents processed at the other two temperatures. Generally, the removal performance increased with the surface area. However, C_500_800 showed slightly higher RhB removal capacities than C_500_700, which had much larger surface area. Due to the complex nature of sludge, we conjectured that other factors in addition to surface area and surface functional groups also contribute to the adsorption process for C_500_700 and C_500_800, such as mass transfer in the pore structure of the adsorbent. Adsorption isotherm studies. Since C_500_800 had the best removal performance, it was chosen for the adsorption isotherm and kinetics studies. Figure 8 (a) presents the relationship between the mass of C_500_800 and the removal of RhB. The removal increased with the adsorbent mass and reached its maximum value of 96.85% at 25 mg. The adsorption behavior was analyzed by fitting the experimental data to two widely used models: the Freundlich and Langmuir isotherm models.
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Figure 8. (a) Effect of mass of adsorbent C_500_800 on RhB removal ratio. Linear fitting curve for experimental data with (b) Freundlich model and (c) Langmuir Model. (d) Adsorption isotherms of RhB on C_500_800; [RhB]=100 mg/L, pH=7, T=25℃, [KNO3]=300 mM. The Freundlich isotherm model is expressed with the following equation:
(Eqn. 2) where
and
represent the mass of RhB adsorbed per mass of adsorbent (mg/g) and the
concentration of RhB in solution (mg/L) under equilibrium conditions.38 The Freundlich isotherm constant
is an indicator for adsorption capacity (mg/g) of the adsorbent, and 1/n is
22
the adsorption intensity. By taking the logarithm of both sides of the equation, the linear form of the Freundlich isotherm model is represented as: (Eqn. 3)
As shown in Figure 7 (b), is 0.9012. The values for
is plotted against
. The coefficient of determination R2
and 1/n are 67.27 mg/g and 0.2264, respectively. The Freundlich
isotherm model did not provide a very good fit for the experimental data points.38 The Langmuir isotherm model is described as follows: (Eqn. 4)
The Langmuir isotherm constant
(L/mg) is related to the energy of adsorption, and
the maximum monolayer adsorption capacity (mg/g).39 To determine
and
is
, the Langmuir
isotherm model is linearized to the following form: (Eqn. 5)
is plotted against
in Figure 7 (c), and the corresponding R2 is 0.9998. Compared with the
Freundlich model, the Langmuir model gave a much better description of the adsorption behavior of the adsorbent.39 The values for
and
are 170 mg/g and 0.3486 L/mg,
respectively. The experimental isotherm and those predicted by the Freundlich and Langmuir models are presented in Figure 8 (d). A low adsorbent input limits the RhB removal, while a high adsorbent doses prevent full and efficient use of its capacity. When 25 mg of adsorbent (1 g/L) was used to
23
treat a 25 mL 100 mg/L RhB solution,
was 97 mg/g, which is up to 60% of the maximum
capacity as determined by the Langmuir model. Meanwhile,
for the RhB solution was only 3
mg/L, indicating a very successful 97% removal. Adsorption kinetics.
Figure 9. Kinetic adsorption data for RhB by C_500_800: (a) without adsorbent regeneration after each cycle and (c) with regeneration. Data fitting with pseudo-second-order kinetic model for (a) and (c) are presented in (b) and (d). [RhB]=200 mg/L, pH=7, T=25℃, [KNO3]=300 mM. The kinetics of RhB removal by C_500_800 was analyzed for better understanding of its adsorption behavior over three cycles of adsorption.
24
A pseudo-second-order kinetic model was used for data fitting, and the rate equation is presented in the following differential form: (Eqn. 6)
where
(mg/g) is the RhB loading on C_500_800 at any time t, and
capacity at equilibrium conditions.40 The term force for the adsorption, and
(mg/g) is the adsorption
(mg/g) could be interpreted as the driving
(g·mg-1·min-1) is the rate constant at a given temperature T (K).
This differential equation is linearized by integrating from t = 0 to t = t: (Eqn. 7)
is plotted against t in Figures 9 (b) and (d) to determine
and
fits the data very well for each cycle. The initial adsorption rate from Eqn. 5 by setting
.40 It is clear that the model (mg·g-1·min-1) is calculated
equal to zero and substituting the values of
and
:
(Eqn. 8) In a 3-cycle test without adsorbent regeneration, the concentrations of RhB solutions after each cycle were 11.84 mg/L, 44.07 mg/L, and 75.28 mg/L, as shown in Figure 9 (a). The performance decreases remarkably after each cycle and was only 27% of the best performance under the same conditions after the 3 cycles. However, when regeneration was applied after each cycle, as shown in Figure 9 (c), the corresponding concentrations of RhB were 13.43 mg/L, 14.73 mg/L, and 27.45 mg/L. Clearly, the adsorption capacity was almost complete restored after the first cycles and decreases only 15% after the second cycle, indicating that the adsorbent could be reused.
25
The parameters predicted by the pseudo-second-order model are presented in Table 3. The equilibrium adsorption capacity qe was 94.49 mg/g, 62.22 mg/g, and 33.56 mg/g for each cycle without adsorbent regeneration and 93.26 mg/g, 91.87 mg/g, and 80.22 mg/g with regeneration. In both cases, qe for the first cycle resulted in good agreement with the value determined from the adsorption isotherm under the same conditions, thus justifying the pseudo-second-order model for the removal system. The rate constant kT and initial adsorption rate r0 decreased rapidly from the first cycle to the third cycle in the absence of adsorbent regeneration. In contrast, when regeneration is performed, kT and r0 were almost constant for the first two cycles and decreased slightly for the third cycle. Table 3. Parameters of pseudo-second-order model for RhB removal using adsorbent C_500_800. Cycle 1 Without regeneration 2 3 1 With regeneration 2 3
qe
kT
r0
94.4875 0.0034
30.1593
64.2152 0.0025
10.2282
33.5626 0.002
2.2573
93.26
27.061
0.0031
91.8657 0.003
25.2652
80.2253 0.0027
17.427
The adsorption behavior of porous materials is affected by both surface adsorption and diffusion in pores. Intraparticle diffusion can be described by the following equation41: (Eqn. 9)
26
Where
is the rate constant (mg·g-1·min-0.5) for intraparticle diffusion. The instantaneous
adsorbent loading
is plotted against
in Figure 10 for adsorption both with and without
adsorbent regeneration.
Figure 10. Intraparticle diffusion analysis for adsorption (a) without adsorbent regeneration and (b) with adsorbent regeneration. An initial steep segment with larger slope followed by a gradual linear segment with smaller slope was observed for all curves in Figure 10. The initial steep segment represented the outer surface adsorption, while intraparticle diffusion resulted in the subsequent gradual linear segment.42 The rate constant
was determined by calculating the slope for the gradual linear
segment. From Figure 10 (a), the rate constant for each cycle was 5.080, 3.687 and 2.308 mg·g1
·min-0.5 when adsorbent regeneration was not performed. By regenerating the adsorbent after
each cycle, the rate constant was 5.428, 5.305 and 5.164 mg·g-1·min-0.5, as shown in Figure 10 (b). Hence, the adsorbent regeneration was also effective in removing the adsorbate from the internal porosity, preventing the quick build-up of the mass transfer resistance after each cycle. In a pure intraparticle diffusion mechanism, qt would increase linearly with
, producing a line
27
without y-intercept.43 Therefore, the adsorption behavior for C_500_800 was governed by both surface adsorption and diffusion in pores. Effects of pH. The effects of pH on the RhB removal were studied, and the results are presented in Figure 11 (a). The pH ranged from 4 to 10 with intervals of 1, and the RhB removal for each condition was 97.30%, 97.80%, 98.42%, 98.27%, 96.93%, 97.16%, and 96.43%, demonstrating a negligible effect of solution pH on adsorption capacity. Zeta potential of C_500_800 was measured at five different pH values, namely 3, 5, 7, 9, and 12, to have a deeper insight into the pH effects on adsorption (Figure 11 (b)). The zeta potential of C_500_800 at each experiment pH were -3.97 mV, -28.33 mV, -31.5 mV, -34.33 mV, and -28.63 mV. No significant changes in zeta potential were observed when pH was higher than 5, but it showed an abrupt increase at pH=3, suggesting the proximity of the isoelectric point for C_500_800 at a slightly lower pH. The results are compatible with a surface with a minimal density of ionizable surface groups in the range of pHs used in the adsorption experiments, which may be partially responsible for the minor effects of pH on adsorption capacity. Therefore, zeta potential measurements were in good accordance with those from adsorption tests at different pH values.
Figure 11. (a) Effects of pH on adsorption and (b) Zeta potential measurements of C_500_800.
28
Effects of ionic strength and anions. The effects of ionic strength and different anions are shown in Figure 12 (a) and (b) . Similarly to the pH dependence experiments, no significant differences were observed under all conditions tested. The concentration of KNO3 was varied between 50 mmol/L and 500 mmol/L, but the RhB removal remained practically constant ranging from 97.98 % to 99.24%. Anions
,
,
, and
were tested at both 150
mmol/L and 300 mmol/L, showing a similar outcome.
Figure 12. Effects (a) ionic strength and (b) different anions on RhB removal. In order to quantify the ionic strength effect, the influence factor K was introduced: ∑
(Eqn. 10)
where y and x represent the dependent variable and independent variable. Experimental conditions for the adsorption isotherm and adsorption kinetics (pH=7, c(KNO3)=300 mmol/L) were used as references, from which % deviation in y and x were calculated to determine K. Generally, the larger the K value, the more responsive the adsorbent is to external influences. The respective K values calculated for pH and ionic strength were 0.014 and 0.0019, which suggest that the performance was unaffected by the water matrix investigated. Therefore,
29
C_500_800 is suitable for various complex systems and could be considered as a promising candidate for industrial waste water treatment. CONCLUSION Carbonized sludge and activated carbonized sludge were synthesized, characterized and studied as an adsorbent under different conditions. The specific surface area of activated carbonized sludge increased with the activation temperature, reached its maximum value at 700 ℃, and decreased with further increase in temperature, while the total pore volume increased with increasing temperature. The specific surface area for C_500_700 and C_500_800 were 642 m2/g and 390 m2/g, and their corresponding total pore volume 0.8 cm3/g and 0.83 cm3/g. The liquid-phase adsorption performance of carbonized sludge and activated carbonized sludge were evaluated by RhB removal. Compared with carbonized sludge, activation was found to be significant for liquid-phase adsorption since RhB removal increased remarkably by more than 400% for all activated carbonized sludge samples. C_500_800 gave the highest removal of 96.85%. A Langmuir-type adsorption isotherm was observed for C_500_800, and
and
were 170 mg/g and 0.3486 L/mg. Kinetics studies showed that the RhB adsorption by C_500_800 followed a pseudo-secondorder model both with and without sample regeneration. The reusability of the activated carbonized sludge was demonstrated over three cycles of adsorption and regeneration. The adsorption performance of C_500_800 was stable under a wide range of pH and ionic strength, which was supported by the surface characterization showing a lack of pH-sensitive ionizable groups under environmentally relevant pH levels. This makes the material highly suitable for various complex treatment systems in industrial wastewater, where performance is expected to
30
be maintained for a wide range of water matrixes. In summary, sludge carbonization and activation is a promising technology for mass sludge treatment due to its low cost and sustainability.
AUTHOR INFORMATION Corresponding Author *Phone 573-884-6777. Email: [email protected] Author Contributions The manuscript was prepared through contributions of all authors. All authors have approved the final version of the manuscript. ACKNOWLEDGMENT This work is supported by CECEP Both Environment Eng. & Tech. Co., Ltd, as a part of the carbonized sludge application project.
31
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S2213-3437(16)30405-5 http://dx.doi.org/doi:10.1016/j.jece.2016.11.013 JECE 1326
To appear in: Received date: Revised date: Accepted date:
25-8-2016 7-11-2016 9-11-2016
Please cite this article as: Fangwei Cheng, Hongxi Luo, Lei Hu, Bin Yu, Zhen Luo, Maria Fidalgo, Sludge carbonization and activation: from hazardous waste to functional materials for water treatment, Journal of Environmental Chemical Engineering http://dx.doi.org/10.1016/j.jece.2016.11.013 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. 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.
Sludge carbonization and activation: from hazardous waste to functional materials for water treatment Fangwei Cheng 1, Hongxi Luo 1, Lei Hu 2, Bin Yu 2, Zhen Luo 2, Maria Fidalgo3 * 1. Department of Chemical Engineering, University of Missouri-Columbia, Columbia, MO 65211, United States 2. CECEP Both Environment Eng. & Tech. Co., Ltd, Wuhan, Hubei 430071, China 3. Department of Civil & Environmental Engineering, University of Missouri-Columbia, Columbia, MO 65211, United States KEYWORDS: Sustainability, Sludge, Carbonization, Water treatment, Adsorption, Sorbents
1
ABSTRACT: The utilization of sludge from wastewater treatment plants as adsorbent material was investigated. Dry sludge from Ezhou Qingyuan sewage treatment plant (Hubei, China) was heated in anaerobic conditions to produce carbonized sludge, that further chemically activated at higher temperatures and K2CO3 to enhance porosity and surface area. The materials were characterized by scanning electron microscopy (SEM), Fourier transformed infrared spectroscopy (FTIR). thermo gravimetric analysis (TGA) and nitrogen adsorption isotherms. TGA curves showed water and low molecular weight organics were lost in a first stage, with the onset of decomposition at 300℃ and up to 700℃; activation resulted in further carbonization. Nitrogen adsorption experiments yielded Type IV isotherms, characteristic of mesoporous materials. Activation greatly increased surface area, reaching up to 642 m2/g. FTIR spectra showed the formation of a carboxyl-metal complex at activation, but no further changes in functional groups with increasing reaction temperature. The adsorption capacities of carbonized and activated sludge towards Rhodamine B were investigated by batch and kinetic experiments. Adsorption increased with activation temperature, reaching a maximum at 700 ℃ with the exception of the sample carbonized at 500 ℃ where a monotonic increase in capacity was observed. Isotherms showed Langmuir-type behavior; kinetic data was successfully fitted to a pseudo second order model. The adsorption was not affected by pH changes or dissolved solids type and concentration. Zeta potential determinations showed minimal variation of surface charge in the pH range of interest. The results indicated that sludge carbonization is a promising sustainable technology for mass sludge treatment.
2
INTRODUCTION The disposal of sludge from wastewater treatment plants is a major environmental concern worlwide1. It contains huge amounts of hazardous pollutants and harmful bacteria, which constitutes a serious threat to the environment and public health when treated ineffectively.1-4 In China, about 36 million tons of wet sludge (80% moisture) are produced from sewage treatment plants every year, and nearly 60% of it is not treated sufficiently.5-7 Traditionally, direct landfilling is the main method of sludge disposal.8-9 However, direct landfilling takes up arable land and increases the possibility of groundwater pollution.8, 10 Alternatively, other methods such as drying and incineration are also employed to stabilize sludge and decrease its volume.11-14 The amount of sludge is dramatically reduced, but the residue requires further disposal, and landfilling is still necessary. Ash from incineration is another type of hazardous waste and needs to be stabilized using concrete. These strategies are not sustainable, and more environmentally conscious approaches are essential for sludge treatment.8 Recently, energy recovery and sludge utilization have become a new trend in sludge treatment, which includes the sludge carbonization approach.15-17 Sludge carbonization is a process where sludge is heated up under anaerobic conditions and transformed into odorless carbon-containing particles.18 In the late 1990s, sludge carbonization techniques were developed by Japanese companies under the regulations of the Japan Sewage Works Association for emission reduction.19 By using carbonized sludge as an alternative fuel, emissions were reduced effectively by 23.5–35.8% compared with direct incineration.19-20 Hence, sludge carbonization has been industrialized, and carbonized sludge is now mainly used as an alternative fuel for power generation in Japan.21
3
Compared to raw sludge, carbonized sludge is a better platform to derive sludge based materials since the composition of carbonized sludge is less complicated. Meanwhile, different modification methods can be chosen for carbonized sludge based on its carbon content as well as the target application. Carbonized sludge can be activated to produce activated carbonized sludge with high porosity and surface area.18 Low-cost, tunable properties, sustainability, and the possibility for mass production make carbonized sludge and activated carbonized sludge promising candidates in fields like air purification and water treatment.22 The combined sewage system in China, where municipal wastewater is collected with storm wastewater, introduces additional challenges for sludge treatment. The first large scale industrial sludge carbonization system in China was set up by CECEP Both Environment Eng. & Tech. Co., Ltd in 2015, with a capacity of 60 ton dry sludge/day. The carbonized was successfully applied in soil amelioration.23 However, the specific surface area of carbonized sludge produced was found to be too low to be directly efficiently used as adsorbent material. Compared with the sludge used in some previous works24, the sludge produced in the plant had a lower organic content (30% ~ 40%)25 and reflects the general situation in China’s wastewater treatment plants. Low organic content in sludge brought in challenges in preparation of activated carbon from sludge. As installed capacity for sludge carbonization increases, there is a need to investigate routes for large scale sludge carbonization-activation systems in order to guide the industry level sludge carbonization/activation in China and worldwide. The aim is to advance the sustainability of the wastewater treatment sector providing a safe disposal method for a waste via the conversion into valuable material. We report in this work a general strategy for large-scale production of activated carbonized sludge from carbonized sludge and apply them in liquid-phase adsorption. The major objectives
4
of this study are (i) to investigate the effects of carbonization temperature on the thermal properties of carbonized sludge and find the optimal carbonization temperature for sequential activation, (ii) to clarify the influence of activation temperature on the specific surface area and pore structures of activated carbonized sludge, and (iii) to evaluate the performance of both carbonized sludge and activated carbonized sludge in liquid-phase Rhodamine B (RhB) removal and to study the interaction between their adsorption differences and porosity. To achieve these objectives, different carbonization temperatures were used to prepare carbonized sludge. For each carbonization temperature, the corresponding carbonized sludge was activated at a different activation temperature to obtain activated samples. Liquid-phase adsorption of RhB was then studied for all samples to assess their performance and collect data for adsorption isotherm and adsorption kinetics. In addition, different conditions such as pH and ionic strength were tested to determine their effects on adsorption. This work presents an approach for massive, sustainable, and low-cost sludge treatment with a focus on application of its products. EXPERIMENTAL METHODS Materials. Sludge was obtained from Ezhou Qingyuan sewage treatment plant (Hubei, China). Ezhou Qingyuan water treatment plant receives storm and domestic wastewater (~95%) and a smaller fraction of pre-treated wastewater from a nearby iron processing plant (5%). The facility includes a 60 t/d sludge carbonization unit on-site. Prominent properties of the raw sludge and the dry sludge used in this work and corresponding wastewater treated by the plant were summarized in Table 1. Table 1. Prominent properties of the wastewater, raw sludge and the dry sludge.25
5
COD
pH Wastewater
BOD5 Total P NH3-N Total Cr Cr(VI) Total Cd
(mg/L) (mg/L) (mg/L) (mg/L) (mg/L) (mg/L) (mg/L)
7.42
52.6
15.9
0.546
6.999
0.012
H2O Organic Ash NH3-N COD SBET (wt%) (wt%) (wt%) (mg/L) (mg/L) (m2/g)
0.009
0.001
pH
Raw sludge
83
5.2
11.8
N/A
N/A
N/A
N/A
Raw sludge_1_HS26
82.4
6.6
11
N/A
N/A
N/A
N/A
Raw sludge_2_TXL26
83
6.8
11.2
N/A
N/A
N/A
N/A
Dry sludge
<1
30.3
68.9
27.17
254.48
9.13
6.96
Cd
Hg
Pb
Cr
As
Ni
Zn
Cu
(mg/kg) (mg/kg) (mg/kg) (mg/kg) (mg/kg) (mg/kg) (mg/kg) (mg/kg) Dry sludge
1.68
0.915
41.6
105
23.9
68.2
731
232
Standard*
20
15
1000
1000
75
200
4000
1500
(*CJ/T 248-2007: The disposal of sludge from municipal wastewater treatment plant – The quality of sludge used for afforestation in gardens or forests; CJ/T: Voluntary Standards issued by the Ministry of Housing and Urban-Rural Development of China.) The physicochemical properties of the sludge used in this work were similar to those from Huangshi WWTP (at Huangshi) and Tangxun Lake WWTP (at Wuhan) in the Hubei Province, also listed for comparison in Table 1 as Raw sludge 1 HS and 2 TXL, and thus, it can be considered representative of large WWTPs in China. Potassium carbonate (K2CO3), hydrochloric acid (HCl), Rhodamine B, hydrochloric acid (HCl), sodium hydroxide (NaOH), potassium nitrate (KNO3), potassium chloride (KCl), potassium sulfate (K2SO4), and potassium dihydrogen phosphate (KH2PO4) were purchased from Sinopharm Chemical Reagent Co,.Ltd. All chemical reagents were of analytical grade and
6
directly used as received without further purification. Ultrapure water (18.2 MΩ·cm) was obtained in the laboratory and used for the preparation of all solutions. Preparation of carbonized sludge and activated carbonized sludge. The synthesis of carbonized sludge and activated carbonized sludge is shown in Figure 1. The process started at the water treatment plant, with the operations limited by the red line. Firstly, sludge from a secondary settling tank was sent to a storage tank and then pumped into a decanter via a sludge feed pump. Flocculants were introduced at the end of the sludge feed pipeline. Most of the water content in the sludge was removed in the decanter, and further treated at 100 ℃ for 12 h to obtain absolute dry sludge, which was used in the synthesis of the carbonized sludge and activated carbonized sludge together with the other chemical reagents.27-28
Figure 1. Preparation procedure of carbonized sludge and activated carbonized sludge; wastewater treatment plant processes are marked in red; carbonization in blue and activation process in green.
7
The dry sludge obtained in the plant was taken to the laboratory to continue with the carbonization process, marked in blue in Figure 1. In a typical experiment, 12 g of absolute dry sludge was added to a quartz crucible without pretreatment, which was then placed in the middle of a tube oven. The oven was heated to 500 ℃ at a constant rate of 10 ℃/min and kept at the final temperature for 1 h.22 Nitrogen was used to provide an inert atmosphere at a flowrate of 60 mL/min. At the end of the carbonization process, the carbonized sludge was collected once the oven was cooled down to room temperature and stored for further use. In order to obtain the activated material, a 2 g amount of the as-synthesized carbonized sludge was mixed and ground with 4 g of K2CO3 as activation agent in a mortar, as schematically depicted in Figure 1, delineated in green. The mixture was added to a quartz crucible in an oven, heated to 600 ℃ at a constant rate of 10 ℃/min and kept at the final temperature for 1 h. Nitrogen at 60 mL/min was again used as a protecting gas during the thermal treatment. After the oven cooled down, the activated carbonized sludge was collected and washed 3 times with 1 mol/L HCl and ultrapure water, followed by drying at 80 ℃ overnight. A series of control experiments was done to investigate the carbonization temperatures and activation temperatures. The products were labeled as C_X for carbonized sludge samples and C_X_Y for activated carbonized sludge samples, where X and Y represent the corresponding carbonization temperature and activation temperature, respectively (e.g., C_500 is carbonized at 500 ℃, while C_500_600 is carbonized at 500 ℃ and activated at 600 ℃). Material characterization. The thermal properties of carbonized and activated carbonized sludge were studied by a thermogravimetric analysis (TGA) instrument (ZCT-1, JYGK, heating rate = 20 ℃/min). The specific surface area and pore volume of all samples were measured by
8
isothermal nitrogen adsorption-desorption analysis (V-sorb 2800P, GAPP). The morphology of the activated carbonized sludge was studied by Scanning Electron Microscopy (SEM) (JSMIT300, JEOL). Surface functional groups of the activated carbonized sludge were determined by Fourier transform infrared spectroscopy (FTIR) analysis (Nicolet 6700, Thermo Scientific), while surface charge was investigated by zeta potential measurements using a zeta potential analyzer (Zetasizer Nano ZS, Malvern). Adsorption performance study. Batch adsorption tests. To evaluate the adsorption performance of the prepared sludges, RhB dye was chosen to represent organic pollutants encountered in dyeing industry in an adsorption test. In a typical experiment, 25 mg of treated (carbonized/activated carbonized) sludge was added to a 100 mL flask containing 25 mL 100 mg/L RhB solution. The suspension was left to equilibrate for 40 min under continuous stirring (600 rpm) at room temperature (25 ℃). After the test, the treated sludge was separated from the solution by centrifugation at 6000 rpm for 10 min, and the concentration of the RhB solution was measured by UV-vis absorbance, using a spectrophotometer (UV-2800, Unico). The removal rate was measured by Eqn. 1: h r
where
and
o
1
(Eqn. 1)
represent the initial RhB concentration and that at time t, respectively.29
Adsorption isotherm studies. Adsorption isotherms were generated for the condition of highest % RhB removal (as determined by the batch adsorption tests), at room temperature (25 ℃) by adding different amounts of treated sludge (2.5, 5, 7.5, 10, 12.5, 15, 20, 25, 30, and 40 mg) to several 100-mL flasks each containing 25 mL of 100 mg/L RhB solution. The pH of the solution
9
was 7 and the ionic strength fixed at 300 mmol/L by addition of KNO3. Separation, RhB concentration determination and percentage removal calculations were performed as described above. Adsorption kinetics. The kinetics of the adsorption process was studied for the treated sludge of highest % RhB removal. In this experiment, 200 mg of the treated sludge were added to 200 mL of a 200 mg/L RhB solution. The pH of the solution was 7 and the ionic strength was controlled by KNO3 (300 mmol/L). The adsorption test was conducted for 40 min under continuous stirring (600 rpm) at room temperature (25 ℃), with samples of the solution taken every 8 min to determine their concentrations by UV-vis spectrophotometry. The stability of the adsorbent was determined by performing the kinetics tests for three consecutive cycles, both with and without regeneration. When regeneration was incorporated, treated sludge was separated after each adsorption cycle and heated to 300 ℃ for 30 min under nitrogen. Effects of pH, ionic strength and anions. The effect of pH on adsorption was studied for the sample with the highest % RhB removal adjusting the pH of the RhB solution between 4 and 10 by addition of 0.1 M HCl or NaOH. Variable concentrations of inorganic salts were added to the solution to investigate the influences of ionic strength on removal capacity.
RESULTS AND DISCUSSION Thermal properties of sludge and carbonized sludge. The sludge carbonization process was analyzed by TGA, as shown in Figure 2. Dry sludge was heated from 30 to 800 ℃ with a programed temperature rise in an anaerobic atmosphere with nitrogen as a protecting gas.
10
Figure 2. Thermogravimetric analysis of sludge carbonization process under inert atmosphere. The curve can be divided into a low temperature region (region I) and a high temperature region (region II). The volatilization of moisture and some compounds of small organic molecules took place in region I with temperature below 300 ℃, and the corresponding weight loss was 16%. The volatilization and decomposition of higher molecular weight organic molecules, such as tar oil, began once the temperature reached 300 ℃,30 and resulted in a weight loss of 22%. As a result, the brown sludge was carbonized and turned black. The weight loss increased an additional 2% when the temperature was raised from 700 to 800 ℃, indicating that further carbonization occurred at those temperatures. The carbonization temperature for sludge processing in industry generally varies from 400 to 600 ℃; the product obtained at 400 ℃ being
11
used for alternative fuel because of the relatively high organic content.30 To study the possibility of transforming carbonized sludge into an activated carbon adsorbent, the highest final temperature in this study was set to 700 ℃. Carbonized sludge samples were produced with final temperatures of 500, 600, and 700 ℃ (C_500, C_600, and C_700). The carbonized sludge contains fixed carbon and inorganic compounds. The fixed carbon content is expected to vary, since the components of the product are temperature dependent. As sludge fixed carbon content is key for its application as absorbent, it was investigated by TGA with an air supply. Based on the TGA (Figure 2), the ratio of fixed carbon to inorganic inert compounds was determined to be around 1:3. The weight percentage of fixed carbon in the carbonized sludge was evaluated as 25% for sludge carbonized at 500 ℃. The carbon content in the carbonized sludge was significantly lower than that in carbon-based materials derived from other waste precursors31-32. The results are shown in Figure 3.
Figure 3. TGA of carbonized sludge with air supply. (a) TGA curves and (b) residuals carbon content.
12
As shown in Figure 3 (a), the TGA curves for samples C_500, C_600, and C_700 had similar shapes with a sharp weight loss between 300 and 550 ℃. However, with increasing carbonization temperature, the weight loss interval became narrower, and its boundaries shifted from 300 and 550 ℃ to 350 and 500 ℃. The shift in the lower boundary to the high temperature region could be attributed to the lower amount of organic residuals in the samples produced at higher carbonization temperature. Considering the constant heating rate during the analysis, the lower carbon content required less time for complete combustion and therefore shifted the upper boundary to the low temperature region. The residual carbon content decreased with carbonization temperatures, and it was determined to be 23.8%, 18.4%, and 17.3% for C_500, C_600, and C_700, respectively as shown Figure 3 (b). The results demonstrated that an increase in the carbonization temperature and residence time lowered the carbon content, indicating that more carbonization occurred. The carbon content for C_700 was only 1% lower than that for C_600, which suggests that the process was almost completed at 700 ℃. The decrease in carbon content is consistent with the hypothesis explaining the shift in the upper boundary of TGA curve to the low temperature region. Surface area and morphology. To study the influence of carbonization temperature and activation temperature on the absorption ability of samples, the surface area and pore size distribution were investigated by N2 adsorption tests at 77 K. Samples tested included carbonized sludge before and after activation with K2CO3 at temperatures of 600 ℃, 700 ℃,and 800 ℃. The BET surface are (SBET), micropore volume (Vmicro), and meso pore volume (Vmeso) are shown in Table 2.
13
Table 2. Textural parameters derived from the N2 adsorption data at 77 K; BET surface are (SBET), micropore volume (Vmicro), and meso pore volume (Vmeso). Sample Code SBET
Vmicro Vmeso
Vtotal
(m2/g) (cm3/g) (cm3/g) (cm3/g) C_500
24
0.01
0.12
0.13
C_500_600
119
0.05
0.23
0.28
C_500_700
642
0.25
0.55
0.8
C_500_800
390
0.15
0.68
0.83
C_600
37
0.02
0.09
0.11
C_600_600
145
0.06
0.27
0.33
C_600_700
389
0.15
0.52
0.67
C_600_800
375
0.12
0.7
0.82
C_700
45
0.02
0.16
0.18
C_700_600
60
0.04
0.19
0.23
C_700_700
271
0.11
0.4
0.51
C_700_800
249
0.09
0.56
0.65
For the different series of carbonized sludge, higher carbonization temperature resulted in higher surface area of the carbonized sludge.33-34 The surface areas of carbonized sludge were 24 m2/g, 37 m2/g, and 45 m2/g for C_500, C_600, and C_700, respectively. The surface area was 119 m2/g for C_500_600 and 642 m2/g for C_500_700. When the activation temperature was raised from 600 to 700 ℃, the corresponding surface area increased remarkably. However, a further increase in temperature from 700 to 800 ℃ resulted in a decrease in surface area, which can be explained by samples losing part of their microporous structures at high temperature. Similar results were observed for both the C_600 and C_700 series, where 700 ℃ was again the
14
activation temperature corresponding to the largest surface area. However, the surface area for activation products processed at 700 ℃ decreased as the carbonization temperature increased from 500 to 700 ℃. Since lower residual carbon content was found for carbonized sludge produced at higher temperature, less carbon was available to react with the activation agent to develop additional porosity and resulted in lower specific surface areas. The volume of micropores exhibited the same trend as the surface area, and the product activated at 700 ℃ had the largest Vmicro for each series.35 This could be explained by the reaction between carbon and K2CO3 being slow, thus developing less microporosity at the relatively low temperature of 600 ℃. On the other hand, a rapid reaction at 800 ℃ resulted in the transformation of micropores into mesopores, reducing the microporal volume and the surface area. The volume of mesopores and total pore volumes monotonically increased with the activation temperature for each series. To investigate the relation between pore structure and adsorption capability, the micropore size distribution and adsorption-desorption isotherms were analyzed. The pore size distribution of the C_500 series is presented in Figure 4 (a): narrow peaks with high intensity were found for each sample between 0.5 and 0.7 nm. At higher carbonization temperatures, the peaks became lower and wider, indicating the loss of well-defined micropore structures. Adsorption-desorption isotherms are shown in the inset graphs: all isotherms were of Type IV according to the IUPAC classification. A sharp increase in the low relative pressure region revealed the existence of micropores; hence, the activated sample could be classified as a microporous material hybrid with mesopores, which is consistent with the textural parameters in Table 2. Similar results were observed for both the C_600 and C_700 series in Figures 4 (b) and (c).
15
Figure 4. Micropore size distributions: (a) C_500 series, (b) C_600 series, and (c) C_700 series. Inset: corresponding nitrogen adsorption and desorption isotherms at 77 K. The morphology of activated carbonized sludge C_500_800 was characterized by SEM and corresponding images at different magnifications are shown in Figure 5.
16
Figure 5. SEM images of C_500_800; scale bars (a) 1µm and (b) 10 µm. The carbonized sludge was composed of porous particles with average size of approximately 50 µm and exhibited high surface roughness with features in the nanometer scale. An irregular surface morphology is favorable for adsorption, as it results in a large fraction of potential readily available sites to the adsorbate molecule, as opposed to those present in the inner porosity that require diffusion of the contaminant within the pore tortuosity to reach them. The inner pores of the carbonized sludge could be accessed through the pore openings observed in the surface, as the one shown in Figure 5 (a) which was approximately 2 m in diameter. Overall, the SEM images of the C_500_800 exhibited well-developed surface pore structures, suggesting an internal sponge-like structure. Moreover, the relatively large particle size (hundreds of micrometers) introduced significant advantages in adsorbent separation. FT-IR analysis. FT-IR analysis was carried out for C_500 series to further investigate the effects of activation temperature on surface functional groups; the results are presented in Figure 6.
17
Figure 6. FT-IR spectra for C_500 series. Five major peaks and absorption bands were found at 3415 cm-1, 1618 cm-1, 1033 cm-1, 780 cm-1 and 471 cm-1 for carbonized sludge C_500. The strong absorption band at 3441 cm-1 can be attributed to the stretching vibration mode of the O-H bond in surface hydroxyl groups or adsorbed water molecules. The peak at 1618 cm-1 can be attributed to the stretching vibration mode of C=O bond or C=C bond. Additionally, the bending vibration of H2O adsorbed on the adsorbent surface may also contribute to the absorption here. A very intensive absorption was observed at 1033 cm-1, indicating the presence of Si-O-Si or Si-O-C bond.36 The doublet at 797 cm-1 and 778 cm-1 corresponded to the Si-O symmetric stretching mode and was the characteristic peak of -quartz.36 The sharp peak at 471 cm-1 was attribute to the asymmetric bending of Si-O-Si bond.36 A small peak found at 2225 cm-1 corresponded to the stretching
18
vibration mode of C≡N or C≡N bond. Another small peak was found at 694 cm-1 and was attributed to the symmetric bending vibration mode of Si-O bond within SiO4 group. Activation with K2CO3 at 600 ℃ introduced a major peak at 2080 cm-1 to the spectrum of C_500_600, that increased with activation temperature. This peak may indicate the existence of cumulated double bonds or carboxyl metal complex (C=O--M); similar results were reported by others using K2CO3 for chemical activation of sludge.37 Another peak at 1397 cm-1 was also observed for C_500_600, which corresponded to the bending vibration mode of C-H bond. When activation temperature was 700 ℃, the doublet at 797 cm-1 and 778 cm-1 was replaced by a weak peak at 775 cm-1. Such transformation might indicate the loss of -quartz in the adsorbent at elevated temperature. Meanwhile, no peak was found around 2200 cm-1 for C_500_700. Similar spectrum was obtained for C_500_800 except for the occurrence of a new peak at 596 cm-1. The Si-O bond was found in all sample spectra and demonstrated the formation of silicate, which was in good accordance with the low organic content in sludge determined from TGA. The lack of abundance of surface groups was also related to the low organic content in the raw sludge. Carboxyl metal complex was found for all activated samples and loss of -quartz in the samples occurred at activation temperature higher than 700 ℃. No major change in surface functional groups was observed when activation temperature was further raised to 800 ℃. Hence, activation and its temperature only slightly altered the composition of the treated sludge and only had minor effects on their surface groups. Batch adsorption studies. The absorption capacity of the carbonized sludge towards RhB was investigated. Figure 7 shows the percentage removal of RhB achieved by each treated sludge sample.
19
Figure 7. RhB removal ratio (%) for (a) C_500 series, (b) C_600 series, and (c) C_700 series. As shown in Figure 7 (a), the RhB removal of the activated sample of the C_500 series was 14.59% for C_500, 60.38% for C_500_600, 92.73% for C_500_700, and 96.85% for C_500_800. Comparison of C_500 and C_500_600 performance showed that activation enhanced the RhB removal ability by more than 40% since more surface area was available for adsorption. When the activation temperature was increased from 600 to 700 ℃, the RhB removal ratio was increased an additional 30%. Activation at 800 ℃ led to the highest RhB removal of 96.85%. Figure 7 (b) shows that the RhB removal for C_600, C_600_600, C_600_700, and C_600_800 was 11.45%, 60.87%, 95.78%, and 78.80%, respectively. Similar to the C_500 series, when the activation temperature was increased from 600 ℃ to 700 ℃, the RhB removal
20
increased by 35%. However, further increases in activation temperature negatively affected adsorption performance. The C_700 series followed exactly the same trend as the C_600 series, as illustrated in Figure 7 (c). The RhB removal for each sample was 14.23%, 59.58%, 82.21%, and 71.34%. Overall, the RhB removal for C_500, C_600, and C_700 samples was much lower than that obtained after activation, which underscores the key role of this process in the reuse of waste sludge as an adsorbent material. Adsorbents carbonized at different temperatures showed similar removal performance (about 60%) when activated at 600 ℃. When 700 ℃ was used for activation, the adsorbent carbonized at 600 ℃ gave the best RhB removal, followed by those carbonized at 500 and 700 ℃. With activation at 800 ℃, an increase in performance was observed for only the adsorbent carbonized at 500 ℃, while a decrease in RhB removal was found for adsorbents processed at the other two temperatures. Generally, the removal performance increased with the surface area. However, C_500_800 showed slightly higher RhB removal capacities than C_500_700, which had much larger surface area. Due to the complex nature of sludge, we conjectured that other factors in addition to surface area and surface functional groups also contribute to the adsorption process for C_500_700 and C_500_800, such as mass transfer in the pore structure of the adsorbent. Adsorption isotherm studies. Since C_500_800 had the best removal performance, it was chosen for the adsorption isotherm and kinetics studies. Figure 8 (a) presents the relationship between the mass of C_500_800 and the removal of RhB. The removal increased with the adsorbent mass and reached its maximum value of 96.85% at 25 mg. The adsorption behavior was analyzed by fitting the experimental data to two widely used models: the Freundlich and Langmuir isotherm models.
21
Figure 8. (a) Effect of mass of adsorbent C_500_800 on RhB removal ratio. Linear fitting curve for experimental data with (b) Freundlich model and (c) Langmuir Model. (d) Adsorption isotherms of RhB on C_500_800; [RhB]=100 mg/L, pH=7, T=25℃, [KNO3]=300 mM. The Freundlich isotherm model is expressed with the following equation:
(Eqn. 2) where
and
represent the mass of RhB adsorbed per mass of adsorbent (mg/g) and the
concentration of RhB in solution (mg/L) under equilibrium conditions.38 The Freundlich isotherm constant
is an indicator for adsorption capacity (mg/g) of the adsorbent, and 1/n is
22
the adsorption intensity. By taking the logarithm of both sides of the equation, the linear form of the Freundlich isotherm model is represented as: (Eqn. 3)
As shown in Figure 7 (b), is 0.9012. The values for
is plotted against
. The coefficient of determination R2
and 1/n are 67.27 mg/g and 0.2264, respectively. The Freundlich
isotherm model did not provide a very good fit for the experimental data points.38 The Langmuir isotherm model is described as follows: (Eqn. 4)
The Langmuir isotherm constant
(L/mg) is related to the energy of adsorption, and
the maximum monolayer adsorption capacity (mg/g).39 To determine
and
is
, the Langmuir
isotherm model is linearized to the following form: (Eqn. 5)
is plotted against
in Figure 7 (c), and the corresponding R2 is 0.9998. Compared with the
Freundlich model, the Langmuir model gave a much better description of the adsorption behavior of the adsorbent.39 The values for
and
are 170 mg/g and 0.3486 L/mg,
respectively. The experimental isotherm and those predicted by the Freundlich and Langmuir models are presented in Figure 8 (d). A low adsorbent input limits the RhB removal, while a high adsorbent doses prevent full and efficient use of its capacity. When 25 mg of adsorbent (1 g/L) was used to
23
treat a 25 mL 100 mg/L RhB solution,
was 97 mg/g, which is up to 60% of the maximum
capacity as determined by the Langmuir model. Meanwhile,
for the RhB solution was only 3
mg/L, indicating a very successful 97% removal. Adsorption kinetics.
Figure 9. Kinetic adsorption data for RhB by C_500_800: (a) without adsorbent regeneration after each cycle and (c) with regeneration. Data fitting with pseudo-second-order kinetic model for (a) and (c) are presented in (b) and (d). [RhB]=200 mg/L, pH=7, T=25℃, [KNO3]=300 mM. The kinetics of RhB removal by C_500_800 was analyzed for better understanding of its adsorption behavior over three cycles of adsorption.
24
A pseudo-second-order kinetic model was used for data fitting, and the rate equation is presented in the following differential form: (Eqn. 6)
where
(mg/g) is the RhB loading on C_500_800 at any time t, and
capacity at equilibrium conditions.40 The term force for the adsorption, and
(mg/g) is the adsorption
(mg/g) could be interpreted as the driving
(g·mg-1·min-1) is the rate constant at a given temperature T (K).
This differential equation is linearized by integrating from t = 0 to t = t: (Eqn. 7)
is plotted against t in Figures 9 (b) and (d) to determine
and
fits the data very well for each cycle. The initial adsorption rate from Eqn. 5 by setting
.40 It is clear that the model (mg·g-1·min-1) is calculated
equal to zero and substituting the values of
and
:
(Eqn. 8) In a 3-cycle test without adsorbent regeneration, the concentrations of RhB solutions after each cycle were 11.84 mg/L, 44.07 mg/L, and 75.28 mg/L, as shown in Figure 9 (a). The performance decreases remarkably after each cycle and was only 27% of the best performance under the same conditions after the 3 cycles. However, when regeneration was applied after each cycle, as shown in Figure 9 (c), the corresponding concentrations of RhB were 13.43 mg/L, 14.73 mg/L, and 27.45 mg/L. Clearly, the adsorption capacity was almost complete restored after the first cycles and decreases only 15% after the second cycle, indicating that the adsorbent could be reused.
25
The parameters predicted by the pseudo-second-order model are presented in Table 3. The equilibrium adsorption capacity qe was 94.49 mg/g, 62.22 mg/g, and 33.56 mg/g for each cycle without adsorbent regeneration and 93.26 mg/g, 91.87 mg/g, and 80.22 mg/g with regeneration. In both cases, qe for the first cycle resulted in good agreement with the value determined from the adsorption isotherm under the same conditions, thus justifying the pseudo-second-order model for the removal system. The rate constant kT and initial adsorption rate r0 decreased rapidly from the first cycle to the third cycle in the absence of adsorbent regeneration. In contrast, when regeneration is performed, kT and r0 were almost constant for the first two cycles and decreased slightly for the third cycle. Table 3. Parameters of pseudo-second-order model for RhB removal using adsorbent C_500_800. Cycle 1 Without regeneration 2 3 1 With regeneration 2 3
qe
kT
r0
94.4875 0.0034
30.1593
64.2152 0.0025
10.2282
33.5626 0.002
2.2573
93.26
27.061
0.0031
91.8657 0.003
25.2652
80.2253 0.0027
17.427
The adsorption behavior of porous materials is affected by both surface adsorption and diffusion in pores. Intraparticle diffusion can be described by the following equation41: (Eqn. 9)
26
Where
is the rate constant (mg·g-1·min-0.5) for intraparticle diffusion. The instantaneous
adsorbent loading
is plotted against
in Figure 10 for adsorption both with and without
adsorbent regeneration.
Figure 10. Intraparticle diffusion analysis for adsorption (a) without adsorbent regeneration and (b) with adsorbent regeneration. An initial steep segment with larger slope followed by a gradual linear segment with smaller slope was observed for all curves in Figure 10. The initial steep segment represented the outer surface adsorption, while intraparticle diffusion resulted in the subsequent gradual linear segment.42 The rate constant
was determined by calculating the slope for the gradual linear
segment. From Figure 10 (a), the rate constant for each cycle was 5.080, 3.687 and 2.308 mg·g1
·min-0.5 when adsorbent regeneration was not performed. By regenerating the adsorbent after
each cycle, the rate constant was 5.428, 5.305 and 5.164 mg·g-1·min-0.5, as shown in Figure 10 (b). Hence, the adsorbent regeneration was also effective in removing the adsorbate from the internal porosity, preventing the quick build-up of the mass transfer resistance after each cycle. In a pure intraparticle diffusion mechanism, qt would increase linearly with
, producing a line
27
without y-intercept.43 Therefore, the adsorption behavior for C_500_800 was governed by both surface adsorption and diffusion in pores. Effects of pH. The effects of pH on the RhB removal were studied, and the results are presented in Figure 11 (a). The pH ranged from 4 to 10 with intervals of 1, and the RhB removal for each condition was 97.30%, 97.80%, 98.42%, 98.27%, 96.93%, 97.16%, and 96.43%, demonstrating a negligible effect of solution pH on adsorption capacity. Zeta potential of C_500_800 was measured at five different pH values, namely 3, 5, 7, 9, and 12, to have a deeper insight into the pH effects on adsorption (Figure 11 (b)). The zeta potential of C_500_800 at each experiment pH were -3.97 mV, -28.33 mV, -31.5 mV, -34.33 mV, and -28.63 mV. No significant changes in zeta potential were observed when pH was higher than 5, but it showed an abrupt increase at pH=3, suggesting the proximity of the isoelectric point for C_500_800 at a slightly lower pH. The results are compatible with a surface with a minimal density of ionizable surface groups in the range of pHs used in the adsorption experiments, which may be partially responsible for the minor effects of pH on adsorption capacity. Therefore, zeta potential measurements were in good accordance with those from adsorption tests at different pH values.
Figure 11. (a) Effects of pH on adsorption and (b) Zeta potential measurements of C_500_800.
28
Effects of ionic strength and anions. The effects of ionic strength and different anions are shown in Figure 12 (a) and (b) . Similarly to the pH dependence experiments, no significant differences were observed under all conditions tested. The concentration of KNO3 was varied between 50 mmol/L and 500 mmol/L, but the RhB removal remained practically constant ranging from 97.98 % to 99.24%. Anions
,
,
, and
were tested at both 150
mmol/L and 300 mmol/L, showing a similar outcome.
Figure 12. Effects (a) ionic strength and (b) different anions on RhB removal. In order to quantify the ionic strength effect, the influence factor K was introduced: ∑
(Eqn. 10)
where y and x represent the dependent variable and independent variable. Experimental conditions for the adsorption isotherm and adsorption kinetics (pH=7, c(KNO3)=300 mmol/L) were used as references, from which % deviation in y and x were calculated to determine K. Generally, the larger the K value, the more responsive the adsorbent is to external influences. The respective K values calculated for pH and ionic strength were 0.014 and 0.0019, which suggest that the performance was unaffected by the water matrix investigated. Therefore,
29
C_500_800 is suitable for various complex systems and could be considered as a promising candidate for industrial waste water treatment. CONCLUSION Carbonized sludge and activated carbonized sludge were synthesized, characterized and studied as an adsorbent under different conditions. The specific surface area of activated carbonized sludge increased with the activation temperature, reached its maximum value at 700 ℃, and decreased with further increase in temperature, while the total pore volume increased with increasing temperature. The specific surface area for C_500_700 and C_500_800 were 642 m2/g and 390 m2/g, and their corresponding total pore volume 0.8 cm3/g and 0.83 cm3/g. The liquid-phase adsorption performance of carbonized sludge and activated carbonized sludge were evaluated by RhB removal. Compared with carbonized sludge, activation was found to be significant for liquid-phase adsorption since RhB removal increased remarkably by more than 400% for all activated carbonized sludge samples. C_500_800 gave the highest removal of 96.85%. A Langmuir-type adsorption isotherm was observed for C_500_800, and
and
were 170 mg/g and 0.3486 L/mg. Kinetics studies showed that the RhB adsorption by C_500_800 followed a pseudo-secondorder model both with and without sample regeneration. The reusability of the activated carbonized sludge was demonstrated over three cycles of adsorption and regeneration. The adsorption performance of C_500_800 was stable under a wide range of pH and ionic strength, which was supported by the surface characterization showing a lack of pH-sensitive ionizable groups under environmentally relevant pH levels. This makes the material highly suitable for various complex treatment systems in industrial wastewater, where performance is expected to
30
be maintained for a wide range of water matrixes. In summary, sludge carbonization and activation is a promising technology for mass sludge treatment due to its low cost and sustainability.
AUTHOR INFORMATION Corresponding Author *Phone 573-884-6777. Email: [email protected] Author Contributions The manuscript was prepared through contributions of all authors. All authors have approved the final version of the manuscript. ACKNOWLEDGMENT This work is supported by CECEP Both Environment Eng. & Tech. Co., Ltd, as a part of the carbonized sludge application project.
31
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