Replacing Potassium Hydroxide with Carbide Lime Waste in Preparing Sludge-Based Activated Carbon for Methylene Blue Removal from Aqueous Solutions

Abstract

Domestic wastewater treatment plants produce large amounts of waste sludge. Sludge can be used to produce activated carbon using potassium hydroxide (KOH) as an activating agent. However, KOH is expensive (relative to the cost of waste carbide lime), making the conversion of waste into valuable products unsustainable. This study explored the utilization of a solid waste by-product, carbide lime waste, as a replacement for KOH to produce sludge-based activated carbon (SBAC). The effects of activation conditions on the characteristics of SBAC were investigated and its performance for methylene blue (MB) removal from a solution was assessed. Post-production analyses using scanning electron microscopy and Fourier-transform infrared spectroscopy indicated that the SBAC produced had a porous surface rich in hydroxyl, aromatic, and alkyl functional groups. Among the tested cases of SBAC prepared using carbide lime, the highest removal of MB (240 mg/g) was achieved for the SBAC prepared at 700 °C with a 1:1 impregnation ratio when activated for 60 min and post-treated with 5M hydrochloric acid. The equilibrium adsorption of MB on SBAC was nonlinear. A strong correlation was found between the pore volume and adsorption capacity of the SBAC produced. The findings of this study suggest that the use of carbide lime waste for SBAC production is a viable alternative to an analytical-grade KOH activator.

1. Introduction

The enormous volumes of sewage sludge produced, along with the associated contaminants that can be released into the environment, present a major challenge in its processing [1,2,3,4]. Approximately 13.0 million dry tons of sewage sludge is produced annually in the European Union [5], while the annual production reached 5.7 million dry tons in 2013 in China [6] and nearly 12.7 million dry tons in 2018 in the United States [7]. Worldwide, the sewage sludge production reached 45 million dry tons in 2017 [8].

Increasing sewage sludge production has compelled waste management authorities to develop sustainable management methods. Conventional sewage sludge disposal methods include land application, landfilling, and composting. One of the concerns regarding these methods is their associated costs. For example, the cost of converting sewage sludge into biosolids ranges from USD 300 to 800 per dry ton, contributing to an average of 40% of the operational cost of wastewater treatment plants [9,10]. Another concern is the high contribution of these disposal methods to greenhouse gas emissions, which are estimated in tons CO₂-eq/year, to be 2176–6026 for landfilling, 2102–5923 for composting, and 1825–5823 for land application [11]. With the evolution of strict legislation for sewage sludge management in certain countries, more sustainable solutions to manage sewage sludge at low costs are needed. One option that considers the circular economy is the transformation of sludge into a high-value-added product, such as activated carbon (AC); this conversion can stabilize trapped contaminants, preventing their release into the environment [12]. Sludge-based activated carbon (SBAC) could be used as an economical alternative to commercial activated carbon (CAC) for removing a broad spectrum of contaminants from water [13,14,15]. Although SBAC is generally of lower quality than CAC [16], the massive amount of sewage sludge produced provides an incentive to improve its performance so that it can be adopted for environmental remediation rather than being treated as an environmental burden.

The production of SBAC from sewage sludge requires carbonization and either physical or chemical activation. Compared to physical activation, chemical activation typically produces SBAC adsorbents with larger surface areas [17,18], but it involves the use of an activating agent, such as sulfuric acid (H₂SO₄), zinc chloride (ZnCl₂), or potassium hydroxide (KOH) [19]. Among the activating agents, KOH has been widely used (Table S1) and has been found to produce more porous structures [19] and enhance the formation of OH functional groups on the carbon surface [20]. Nonetheless, the surface area of the SBAC produced is influenced by the characteristics of the sewage sludge [21], impregnation ratio (i.e., the weight ratio of the activating agent to the sludge) [18,22], activation time, and activation temperature [22]. Although a higher impregnation ratio generally leads to a larger surface area, using large amounts of analytical-grade activating agent does not favor sustainable production. From an economic perspective, the production cost of SBAC is approximately USD 1.2/kg [23]. This is much cheaper than the production cost of CAC (>USD 3.0/kg) [24] but more expensive than those of other adsorbents derived from dried biomass and natural materials (less than USD 0.2/kg [25]). A significant portion of the SBAC production cost is associated with the use of activating agents, which can exceed 55% of the total production cost [23]. Thus, low-cost activating agents are required to lower the overall production costs of SBAC.

One option is to replace the analytical-grade activator with a waste-derived material. Carbide lime waste generated during acetylene production is a promising candidate for this purpose. It consists predominantly of calcium hydroxide (>90% [26]). Because of its elevated pH, carbide lime waste is usually disposed of in designated landfills. To the best of our knowledge, no study has used carbide lime waste in the chemical activation of sewage sludge, nor has it been used to prepare any AC material. Analytical-grade hydrated lime, however, has been used to produce AC from other carbonaceous materials. For example, Takemura et al. [27] reported that the use of analytical-grade hydrated lime in the carbonization of waste phenol resin resulted in a carbonized material with a surface area nearly five times larger than that obtained without the use of the activating agent. AlOthman et al. [28] prepared AC from palm, paper, and plastic waste using analytical-grade hydrated lime for activation. They found that the adsorption capacity of the prepared AC towards methylene blue (MB) increased with an increase in the impregnation ratio. Two other studies investigated the use of analytical-grade hydrated lime as an auxiliary activating agent with ZnCl₂ in preparing SBAC to remove dyes [29] and some pharmaceuticals [30]. Alternative approaches have been suggested to eradicate dyes from water sources. Among these methods is the utilization of ZnFe₂O₄-layered double hydroxide or MgAl-layered double hydroxide/carbon fiber, both recognized as potent adsorbents for eliminating Cango Red from water [31,32]. Others used bamboo-derived biochar for the removal of MB from aqueous solution [33].

MB has been selected in this study as a model organic contaminant due to its wide applications in various industries including paper, cotton, wool, silk, and pharmaceuticals. Exposure to MB has been associated with several health concerns, including eye irritations, mental disorders, digestive complications, and respiratory issues [34].

The potential use of carbide lime waste to prepare SBAC for water purification conforms to circular economy principles and the United Nations Sustainable Development Goals (SDG 6: Clean Water and Sanitation and SDG 11: Sustainable Cities and Communities). Specifically, combining sewage sludge waste with carbide lime waste for SBAC production could produce a value-added product, reducing the cost and environmental impact of current disposal methods. Utilizing these waste materials could reduce the production cost of AC, possibly enhancing its feasibility of usage compared to those of other adsorbents. However, several research questions regarding the use of carbide lime waste to produce SBAC must be addressed. The questions relevant to the current study are the following: (1) Is SBAC produced using carbide lime waste (SBAC-CL) as effective as that produced using KOH (SBAC-KOH) in removing MB from the solution? (2) How do the preparation conditions affect the characteristics and adsorption behavior of SBAC-CL? Accordingly, this study aimed to investigate the use of carbide lime for sewage sludge activation and evaluate the properties and performance of the thus-produced SBAC in removing MB from aqueous solutions. Another objective was to evaluate the performance of SBAC-CL and compare it with those of SBAC-KOH and CAC.

2. Materials and Methods

2.1. Materials

Undigested sewage sludge samples were collected from the Sharjah Municipality Wastewater Treatment Plant in the UAE. Sludge samples were collected from various locations in the drying bed and combined into a composite sample to represent the bulk volume at the site. Characteristics of the collected sewage sludge are listed in Table S2. Carbide lime waste slurry was collected in airtight plastic containers from the Emirates Gas Company (Dubai, United Arab Emirates), where it is generated as a waste product of acetylene gas production. This waste material mainly consists of calcium hydroxide (90.5%), iron oxide (1.85%), magnesium oxide (2.83%), and silicon oxide (1.54%) [26]. The slurry was decanted in the laboratory, and settled residues were dried at 105 °C for 24 h. Dried carbide lime was crushed to pass through a 100 µm sieve and stored in a dark and dry area until use. KOH and hydrochloric acid (HCl) were used as the analytical reagents. KOH was purchased from Millipore Sigma (Darmstadt, Germany), and HCl was obtained from SDFC (Chennai, India). MB was obtained from Sigma-Aldrich (Taufkirchen, Germany). High-purity (99.999%) nitrogen was purchased from the Sharjah Oxygen Company (Sharjah, United Arab Emirates). All solutions were prepared using deionized (DI) water produced using a Millipore Synergy UV Milli-Q system. CAC (granular NORIT^®charcoal) was obtained from Sigma-Aldrich (Germany). Sodium dihydrogen phosphate was obtained from Merck (Darmstadt, Germany), and disodium hydrogen phosphate was obtained from Honeywell Fluka (Seelze, Germany).

2.2. Preparation of Adsorbents

Several sewage-sludge-based adsorbents were prepared in this study (

Table 1. Preparation conditions of the tested adsorbents.

Adsorbent	Activating Agent	Impregnation Ratio	Activation Time (min)	Activation Temperature (°C)
CS	None	NA	60	700
SBAC-CL1	Carbide lime	1:4	60	700
SBAC-CL2	Carbide lime	1:1	30	700
SBAC-CL3	Carbide lime	1:1	60	700
SBAC-CL4	Carbide lime	1:1	60	800
SBAC-KOH1	KOH	1:4	60	700
SBAC-KOH2	KOH	1:1	30	700
SBAC-KOH3	KOH	1:1	60	700
SBAC-KOH4	KOH	1:1	60	800

). To prepare the SBAC adsorbents, the collected sewage sludge composite samples were first dried in an oven at 105 °C for 24 h to achieve complete dryness. The dry material was ground in a grinder, and the portion passing through a 425 µm mesh was used. To chemically activate the sludge, either KOH or carbide lime was used. The activation parameters (impregnation ratio, activation temperature, and activation time) were within the ranges reported by other investigators (Table S3). For a given SBAC mixture, 100 g of dried sludge powder was mixed with a certain amount of an alkali agent (carbide lime or KOH), depending on the impregnation ratio. After adding 500 mL of DI water, continuous mixing for 12 h at 70 °C was carried out in a covered beaker. The mixture was then poured into an aluminum tray and dried at 105 °C for 24 h. Chemically activated dried sludge was crushed and sieved through the 425 µm mesh. Ten grams of sludge powder from each SBAC material prepared was transferred into an alumina ceramic crucible boat of 1.2 cm depth, 1.7 cm width, and 9.7 cm length. The boat was transferred to a furnace tube (GSL-1500X-50-UL, MTI Corporation, Richmond, CA, USA) under high-purity N₂ gas. Nitrogen gas was run for 5 min at a flow rate of 100 mL/min to deoxygenate the tube before pyrolysis. During pyrolysis, the furnace temperature was set at 700 °C or 800 °C for 1 h at 11.31 °C/min or 13.3 °C/min, respectively. The materials were activated for 30 or 60 min (Table 1). Then, the furnace was cooled to 30 °C, at which point the flow of N₂ gas was switched off. A carbonized sludge (CS) adsorbent was also prepared from the sewage sludge but without chemical activation. For this adsorbent, 10 g of ground sewage sludge was placed in the furnace tube with N₂ gas; the material was activated at 700 °C for 60 min (Table 1).

After pyrolysis, the prepared adsorbents (CS and SBAC) were treated with 5 M HCl prepared from 37% concentrated HCl. A solution with a solid-to-acid ratio of 1:2 was placed in a glass beaker and mixed using a magnetic stirrer for 30 min. The material was transferred to 50 mL tubes for centrifugation (8000 rpm) to separate the adsorbent from the solution. The adsorbent was then washed with DI water and transferred to a beaker. The acid-washing cycle was repeated two more times. The adsorbent samples were then dried at 105 °C for 24 h, cooled at 21–23 °C, and crushed to pass through a 425 µm sieve.

2.3. Characterization

Semi-quantitative X-ray fluorescence (XRF) analysis was performed to ascertain the approximate percentage of minerals in the produced adsorbents. Each adsorbent (0.3–0.5 g) was subjected to XRF analysis using a Shimadzu 7000EDX instrument (Kyoto, Japan). Fourier transform infrared spectroscopy (FT-IR) was used to determine the surface functional groups of the produced materials. FTIR was performed for the molecule’s functional group identification. FTIR is particularly sensitive to the functional groups present in organic compounds. It excels at detecting molecular vibrations involving changes in dipole moments (e.g., O-H, C=O, N-H bonds). We mixed 1 mg of each adsorbent with 300 mg of potassium bromide (KBr) before placing the mixture in a spectrometer (Nexus Thermo Nicolet 670, Waltham, MA, USA). The analysis was performed over the 400–4000 cm⁻¹ range at a resolution of 1 cm⁻¹. Scanning electron microscopy (SEM) was used to study the surface morphologies of the prepared adsorbents. After coating with a gold layer to ensure electron conductivity, the samples were analyzed using a JEOL JSM-6390 microscope (Tokyo, Japan) in the high-vacuum mode. The microstructural properties of the adsorbents were characterized by N₂ adsorption–desorption isotherms at 77 K using an Autosorb IQ3 (Quantchrome, Boynton Beach, FL, USA) automated gas sorption analyzer. Before the test, 0.05–0.08 g of each adsorbent was degassed under a vacuum at 250 °C for 24 h to remove the moisture in the sample. The Brunauer–Emmett–Teller (BET) equation was used to calculate the specific surface area, and the pore volume was calculated from the amount of adsorbed N₂ at P/P_o = 0.9. The pore size distribution was determined using the Barrett–Joymer–Hanlenda model.

2.4. Adsorption Studies

Adsorption experiments included rate and equilibrium studies. MB was purchased from Sigma-Aldrich with an analytical purity > 99%. A 1000 mg/L MB stock solution was prepared in a pH 7 buffer. The buffer solution was prepared by dissolving 1.20 g of sodium dihydrogen phosphate and 0.885 g of disodium hydrogen phosphate in one liter of DI water. The desired MB concentration used in the adsorption experiments was obtained by diluting the buffer solution.

The adsorption rate studies were conducted over a 96 h period with initial MB concentrations of 60 and 400 mg/L. Glass vials (50 mL) with Teflon caps were used. The adsorbents (0.04 g SBAC and CAC, and 0.1 g CS) were transferred into vials, and MB solution was added to each vial. The reaction vials were wrapped in aluminum foil and rotated end-over-end using a rotator (Parvalux, Poole, UK) at 52 rpm for the desired time (1–96 h). Aqueous samples were filtered using 0.45 µm polytetrafluorothylene filters (Sartorius, Göttingen, Germany) and analyzed using a HACH model DR6000 (Loveland, CO, USA) spectrophotometer to determine the absorbance at 664 nm. The MB concentration was determined using a previously constructed calibration curve with a limit of linearity (R² > 0.99) below 20 mg/L. Samples with concentrations exceeding 20 mg/L were diluted with DI water to lower their concentrations to within the linear range. The determined concentrations were then multiplied by the dilution factor to obtain the actual concentrations.

Based on the results of the rate experiments, adsorption isotherm experiments were conducted for a mixing time of 96 h using the same procedure described for the rate experiments. The initial MB concentration ranged from 20 to 100 mg/L when using CS and from 70 to 700 mg/L when using SBAC.

The concentration of MB on the adsorbent (q_t) was determined using

$${\mathrm{C}}_{\mathrm{o}}\mathrm{V}+{\mathrm{C}}_{\mathrm{t}}\mathrm{V}={\mathrm{q}}_{\mathrm{t}}\mathrm{M}$$

(1)

where C_o is the initial aqueous concentration of the adsorbate, C_t is the aqueous adsorbate concentration after a certain mixing time, V is the volume of the solution, and M is the adsorbent mass. At equilibrium, C_t is referred to as C_e, and q_t is referred to as q_e, which is the adsorbent uptake capacity. The removal efficiency (RE) was calculated as follows:

$$\mathrm{R}\mathrm{E}=\frac{{\mathrm{C}}_{\mathrm{o}}-{\mathrm{C}}_{\mathrm{e}}}{{\mathrm{C}}_{\mathrm{e}}}\times 100$$

(2)

The adsorption rate data were analyzed using the pseudo-first-order (PFO), pseudo-second-order (PSO), and intraparticle diffusion models [31,35], given by Equations (3)–(5), respectively.

$${\mathrm{q}}_{\mathrm{t}}={\mathrm{q}}_{\mathrm{e}}\left(1-{\mathrm{e}}^{-{\mathrm{k}}_1\mathrm{t}}\right)$$

(3)

$${\mathrm{q}}_{\mathrm{t}}=\frac{{\mathrm{k}}_2{\mathrm{q}}_{\mathrm{e}}^2\mathrm{t}}{1+{\mathrm{q}}_{\mathrm{e}}{\mathrm{k}}_2\mathrm{t}}$$

(4)

$${\mathrm{q}}_{\mathrm{t}}={\mathrm{k}}_{\mathrm{p}}{\mathrm{t}}^{0.5}+\mathrm{C}$$

(5)

where k₁ and k₂ are the PFO and PSO rate constants, respectively, k_p is the intraparticle diffusion rate constant, and C represents the intercept. The values of the PFO and PSO model parameters (q_e, k₁, and k₂) were determined through a regression analysis of the nonlinear and linearized forms of their respective equations, whereas the parameters k_p and C were determined using the linearized form of the intraparticle diffusion model. To perform nonlinear regression, the solver option in Excel was used.

The equilibrium adsorption data were fitted using the Langmuir and Freundlich adsorption isotherm models. These models are presented in Equations (6) and (7), respectively [36].

$$\frac{1}{{\mathrm{q}}_{\mathrm{e}}}=\frac{1}{{\mathrm{q}}_{\mathrm{m}\mathrm{a}\mathrm{x}}}+\left(\frac{1}{{\mathrm{q}}_{\mathrm{m}\mathrm{a}\mathrm{x}}{\mathrm{K}}_{\mathrm{L}}}\right)\frac{1}{{\mathrm{C}}_{\mathrm{e}}}$$

(6)

$${\mathrm{q}}_{\mathrm{e}}={\mathrm{K}}_{\mathrm{F}}{\mathrm{C}}_{\mathrm{e}}^{\mathrm{n}}$$

(7)

where K_L is the Langmuir adsorption constant, q_max is the maximum adsorption capacity, K_F is the Freundlich adsorption coefficient, and n is the Freundlich adsorption intensity.

3. Results

3.1. SBAC Characteristics

The elemental composition of CS and SBAC was analyzed via XRF (

Table 2. Elemental analysis of the adsorbents produced in weight percentages.

Adsorbent	Na	Mg	Si	K	Ca	S	Cl	Fe
CS	92.49	4.79	1.57	0.09	0.13	0.09	0.65	0.20
SBAC-CL1	92.23	4.76	1.42	0.02	0.24	0.38	0.87	0.08
SBAC-CL2	94.62	2.89	1.15	0.02	0.31	0.20	0.81	0.02
SBAC-CL3	93.82	3.37	1.40	0.02	0.31	0.27	0.79	0.03
SBAC-CL4	86.53	6.66	4.42	0.07	0.33	1.14	0.76	0.08
SBAC-KOH1	88.23	8.44	1.64	0.46	0.05	0.18	0.58	0.42
SBAC-KOH2	87.07	9.08	2.63	0.19	0.05	0.15	0.64	0.19
SBAC-KOH3	93.03	4.12	1.24	0.18	0.04	0.27	0.92	0.21
SBAC-KOH4	80.89	14.42	3.12	0.18	0.04	0.16	0.57	0.64

). In all the samples, Na had the highest percentage, followed by Mg and Si. The proportions of Mg and Na in the SBAC materials prepared were inversely related. In addition, K and Fe were higher in SBAC-KOH, Ca was higher in SBAC-CL, and the KOH-prepared samples had a higher proportion of K than CS. Increasing the activation time from 30 to 60 min did not affect the fraction of major cations (Na, Mg, and Si) in SBAC-CL, but it increased the Na and reduced the Mg and Si fractions in SBAC-KOH. However, increasing the activation temperature from 700 to 800 °C resulted in a decrease in the Na fraction and an increase in the Mg and Si fractions of SBAC-CL and SBAC-KOH. It should be noted that the presence of minerals on the surface of AC promotes contaminant adsorption [37,38].

Figure 1 shows the FTIR spectra of SBAC and CS. The SBAC samples had similar functional groups; however, the peak intensities differed between those activated with carbide lime and with KOH. The broad band at 3600–3200 cm⁻¹ was attributable to O–H stretching vibration caused by hydroxylic groups and chemisorbed water. Strong hydrogen bonds are responsible for the asymmetry of this band at lower wavenumbers [39,40]. This peak also superimposes the stretching vibration peaks of the amine group [41]. The bands at 3000–2800 cm⁻¹ can be assigned to saturated C−H stretching vibration [42], mainly the methyl and methylene groups. The bands in the region of 1600–1560 cm⁻¹ can be ascribed to the stretching vibrations of C=O moieties in conjugated systems or OH bending vibrations overlapped with aromatic ring stretching and double bond vibrations [39,40]. The broad band in the region of 1300–1000 cm⁻¹ can be attributed to C–O stretching or O–H bending of the alcoholic, phenolic, and carboxylic groups [43]. Within this region, the Si-O bond of silicates is at 1100 cm⁻¹ [36]. Small peaks at 800 cm⁻¹ can be assigned to cyclic compounds containing conjugated C=C and C=N [40]. Below 800 cm⁻¹, the adsorption can be attributed to the deformation vibrations of C-H groups at the edges of the aromatic planes [40]. SBAC-KOH materials are distinguished from SBAC-CL ones in terms of having an imine group at about 2090 cm⁻¹. The intensity of this group in SBAC-KOH materials appeared to increase with an increasing impregnation ratio and activation temperature. Notably, the adsorption peaks of SBAC-CL materials at 700 °C for 60 min had higher intensities than those of the corresponding peaks of the other SBAC-CL materials.

Figure 2 shows SEM images of CS and SBAC. The micrographs illustrate the surface roughness and porosity of the SBAC materials prepared. A comparison of the SBAC and CS images indicates more pores and fragmentation in the former. SBAC-CL likely had more irregular channels in the clusters (Figure 2c). In contrast, SBAC-KOH had less irregularity but more pores (Figure 2d). SBAC-CL appeared to contain more deposits on its surface.

The surface areas and pore volumes of the tested materials are given in

Table 3. Surface areas and pore volumes of adsorbents produced.

Adsorbent	SBET (m2/g)	Pore Volume (cm3/g)
CS	107.2	0.105
SBAC-CL1	238.8	0.213
SBAC-CL2	NA	NA
SBAC-CL3	330.8	0.278
SBAC-CL4	308.2	0.205
SBAC-KOH1	578.3	0.323
SBAC-KOH2	NA	NA
SBAC-KOH3	1068.8	0.518
SBAC-KOH4	742.2	0.446

. The CS exhibited the lowest porosity, with a surface area of 107 m²/g and a pore volume of 0.105 cm³/g; however, alkali activation significantly increased both. When carbide lime was employed for activation, the resulting SBAC exhibited a surface area ranging from 240 to 330 m²/g and a pore volume ranging from 0.2 to 0.28 cm³/g. When comparing the SBAC materials produced using KOH and carbide lime, the SBAC material created using a 1:1 impregnation ratio and activated at 700 °C displayed the largest surface area. However, increasing the activation temperature from 700 to 800 °C and reducing the impregnation ratio from 1:1 to 1:4 led to a decrease in both the surface area and the pore volume of both types of SBAC materials. At a low impregnation ratio, the amount of activating agent may not be sufficient to activate all sites. The increase in surface area with an increasing impregnation ratio is consistent with the findings of others [17,18,44].

The differences in surface area and pore volume between SBAC-KOH and SBAC-CL materials were attributed to the type of activator used. The activation process using a metal hydroxide can involve several reactions (see Equations (8)–(11) for KOH). The extent of these reactions, or their equivalence in the case of carbide lime, depends on the applied experimental conditions. For example, Liou et al. [44] indicated that the overall activation process using NaOH as the activating agent most likely occurred via a reaction equivalent to that of Equation (8). Otowa et al. [45] claimed that KOH activation occurred via Equation (9). Guo et al. [46], on the other hand, suggested that activation using KOH occurred by either Equation (9) or Equations (9) and (10), depending on the preparation method of the AC material. The larger surface area achieved for SBAC-KOH in this study could be due to the superiority of KOH as an activator over other metal hydroxides [46]. Another reason is that the use of KOH produces K₂CO₃, which could decompose to K₂O and participate in the activation process at 650–750 °C [46], whereas CaCO₃ formed using lime requires a temperature of 800 °C to completely decompose to CaO in the presence of N₂ [47]. The third reason could be metal carbonates remaining on the surface after activation. With KOH activation, any K₂CO₃ remaining on the surface or in the pores can be easily removed by acid washing because of the high solubility of the compound, whereas the removal of CaCO₃ in the case of carbide lime activation is more difficult because of its low solubility.

4KOH + 2C → 4K + 2CO + 2H₂O

(8)

6KOH + C → 2K + 3H₂ + 2K₂CO₃

(9)

2K₂CO₃ + C → 4K + 3CO₂

(10)

CO₂ + C → CO

(11)

The decrease in the surface area at higher temperatures for SBAC-KOH and SBAC-CL (Table 3) could be due to violent gasification reactions that could destroy some of the pores of the SBAC material [48]. At higher temperatures, contraction due to the aromatization of the sludge causes the narrowing or closing of the pore entrances, reducing the specific surface area [49]. In this study, the effect of damaging the carbonaceous material occurred when the activation temperature increased from 700 to 800 °C. This damage was associated with a decrease in the Na fraction of both materials (Table 2). Liou et al. [44] observed this damaging effect, but in their study, it occurred with the increase in the temperature from 900 to 1000 °C.

The N₂ adsorption–desorption isotherms and pore size distributions of the SBAC materials are shown in Figure 3. The N₂ adsorption–desorption isotherms show a sharp increase at low P/P_o (<0.09) and a hysteresis loop at P/P_o > 0.4, indicating the presence of micropores in the structure. In addition, the isotherms show a B-type hysteresis loop, suggesting the presence of parallel slit-shaped pores. The pore size with the highest frequency for the produced SBAC materials was 1.2 nm (Figure 3). Thus, the obtained SBAC materials mainly contained micropores (pore size < 2 nm); however, some mesopores (pore size 2–5 nm) were also present.

3.2. Adsorption of MB

3.2.1. Effects of Activation Conditions

Figure 4 shows the changes in the uptake capacity of the prepared adsorbents over time subject to the explored activation conditions. Activation using carbide lime or KOH at a 1:4 impregnation ratio improved the adsorption capacity of SBAC for MB compared to that of non-activated sludge (Figure 4a). Chemical activation is known to improve the adsorption capacity of SBAC owing to an increase in the specific surface area and the formation of functional groups on the surface [15,50]. A greater improvement in the uptake of MB was observed with SBAC-KOH than with SBAC-CL (Figure 4a). The uptake capacity of SBAC-KOH at a contact time of 96 h was approximately 120% higher than that of SBAC-CL. Despite the superiority of SBAC-KOH as an adsorbent, the use of carbide lime waste as an activating agent in the production of SBAC is more sustainable from an environmental, and probably an economic, perspective.

Increasing the impregnation ratio from 1:4 to 1:1 increased the uptake capacity of the SBAC materials for MB (Figure 4b); for SBAC-KOH, the increase was approximately 35% at 96 h. The impregnation ratio significantly affected the surface area and surface functional groups of the SBAC materials prepared. For example, SBAC-CL3 had a surface area of 330.8 m²/g (Table 3) and a higher intensity of hydroxyl, aromatic, and alkyl functional groups (Figure 1), whereas SBAC-CL1 had a surface area of 308 m²/g with a lower intensity of functional groups. When comparing the performance of SBAC materials activated with KOH and carbide lime, it was observed that the SBAC-KOH materials exhibited superior MB adsorption capabilities. This can be attributed to their significantly larger surface areas, ranging from 578 to 1068 m²/g, in contrast to those of SBAC-CL materials (239 to 308 m²/g). Although the FTIR results (Figure 1) showed that some SBAC-CL materials had a larger number of functional groups than their KOH-activated counterparts, the surface area had a more substantial influence on the adsorption of MB. This is consistent with the findings of Hadi et al. [51], who reported an increase in the adsorption capacity of SBAC for various contaminants with an increase in the surface area.

Figure 4c shows that SBAC-CL and SBAC-KOH materials prepared by activation for 60 min had a higher uptake capacity than their counterparts prepared by activation for 30 min. The increase in the uptake capacity at 96 h for SBAC-CL was 56% (from 160.9 to 252 mg/g), while that for SBAC-KOH exceeded 150% (from 116 to 297 mg/g). A higher uptake capacity with an increasing activation time was also reported by AlOthman et al. [28], who investigated MB removal by AC made of mixed solid waste. This enhancement can be attributed to the increased surface area of the prepared carbon at longer activation times [52]. Meanwhile, an increase in the activation time could increase the level of interaction between the raw sludge and the activating agent, resulting in a richer surface chemistry.

Figure 4d shows the adsorption results for SBAC-CL and SBAC-KOH materials prepared at different temperatures (700 and 800 °C). SBAC-CL produced at different temperatures showed similar MB adsorption for a contact time of less than 6 h. However, at a 96 h adsorption time, the adsorption capacity varied significantly, reaching 252 mg/g for SBAC prepared at 700 °C and 182 mg/g for SBAC prepared at 800 °C. SBAC-KOH showed a different trend (i.e., a higher adsorption capacity with a higher activation temperature). Previous studies have indicated that higher temperatures lead to better surface areas and pore structures, particularly at temperatures exceeding 600 °C [52,53]. In this study, SBAC materials prepared at 700 °C had larger BET surface areas and pore volumes than those prepared at 800 °C (Table 3). However, SBAC-KOH prepared at a higher temperature (800 °C) showed better adsorption. These results indicate that the BET surface area and pore size are not the sole indicators of the functionality of SBAC.

3.2.2. Adsorption Kinetics

The results presented in Figure 4 indicate that a mixing time of 96 h was close to equilibrium. In all cases, the reaction rate was characterized as initially rapid at early times, followed by a significantly slower rate to equilibrium. On average, two-thirds of the MB uptake by SBAC materials was reached within an hour. The rapid adsorption rate during early stages was due to the availability of a large number of adsorption sites; however, as they became occupied, the adsorption rate decreased. A slow approach to equilibrium could also indicate a mass-transfer limitation, possibly related to slow diffusion into the micropores and mesopores. The adsorption kinetics were modeled using PFO, PSO, and intraparticle diffusion models. Curve fitting of the nonlinear forms of the PFO and PSO models was performed to obtain the rate parameters (

Table 4. PFO, PSO, and intraparticle diffusion optimized model parameters ¹.

Adsorbent	qe-Exp	PFO (Nonlinear)			PSO (Nonlinear)			Intraparticle Diffusion			PSO (Linearized)
		qe	k1	R2	qe	k2	R2	kp	C	R2	qe	k2	R2
CS	20.7	17.9	0.17	0.913	19.5	0.014	0.959	1.5	7.2	0.909	21.6	0.0076	0.994
SBAC-CL1	99.2	83.2	0.44	0.921	90.8	0.006	0.972	6.0	45.4	0.850	101.0	0.0023	0.995
SBAC-CL2	160.9	159.6	1.73	0.995	162.3	0.025	0.998	2.9	141.3	0.545	161.3	0.0769	1.0
SBAC-CL3	252.0	209.5	0.93	0.908	224.0	0.005	0.953	16.4	144.5	0.944	256.4	0.0011	0.995
SBAC-CL4	182.1	161.0	1.31	0.965	166.2	0.013	0.978	6.3	123.7	0.876	185.2	0.0022	0.998
SBAC-KOH1	217.4	185.3	1.50	0.933	191.6	0.013	0.953	8.3	140.0	0.966	217.4	0.0016	0.996
SBAC-KOH2	115.9	105.6	2.02	0.978	107.3	0.047	0.984	2.63	90.8	0.939	116.2	0.0051	0.999
SBAC-KOH3	297.1	266.0	1.20	0.958	277.1	0.006	0.980	11.3	198.9	0.825	303.0	0.0017	0.999
SBAC-KOH4	318.2	307.6	1.71	0.994	312.2	0.014	0.995	5.6	272.3	0.546	322.6	0.0044	1.0

¹ q_e and C in mg/g. k_p in mg/g.h^0.5. k₁ in 1/h. k₂ in g/mg.h.

). The term q_e-Exp in Table 4 refers to the adsorbed concentration obtained independently from the experiments at t = 96 h. As shown in Table 4, the PFO and PSO models adequately explained the adsorption rate of MB onto the SBAC materials prepared. However, the PSO model generally outperformed the PFO model. The table also shows that k₁ for SBAC-CL was lower than that for SBAC-KOH under the same preparation conditions, suggesting a faster reaction with SBAC-KOH. The k₂ values of the PSO model also exhibited the same trend, but with smaller deviations between the compared values. The table also shows higher k₁ and k₂ values for SBAC materials than those of CS. The reaction kinetics were simulated using the linearized form of the PSO model. Linear fitting of the PSO model resulted in higher R² values than those obtained by nonlinear fitting (with all values close to 1.0).

The deviations between q_e-Exp and the fitted values of the rate models were generally minimal. The absolute deviations in q_e values ranged from 0.81% to 16.8% when the nonlinear fitting of the PFO model was used, and they fell within a narrower range (0.87–11.9%) with the nonlinear fitting of the PSO model. However, the absolute deviations in q_e decreased significantly and ranged between 0 and 2% when the linearized fitting of the PSO model was used. The predicted q_e values obtained by nonlinear fitting were lower than the experimental q_e. In contrast, the predicted q_e values were slightly higher than the experimental q_e values when linearized fitting was used. The latter indicates that the equilibrium may not have been reached by 96 h.

Fitting the intraparticle diffusion model generally resulted in a lower R² than those of the PFO and PSO models (Table 4). Nonetheless, the C values of the intraparticle model indicate that more than 50% of the total sorbed concentration, in most of the cases, instantaneously adheres to the adsorbent. The inadequacy of the intraparticle diffusion model to describe the rate data is possibly due to variations in the diffusion behavior over time where it is higher at early times but reduces as it approaches equilibrium due to the difficulty of accessing the interior adsorption sites at later times. The highest k_p value is associated with the SBAC-KOH3 and SBAC-CL3 materials while the lowest value is associated with the CS material. Figure S1 shows examples of the simulated rate behaviors of the PFO, PSO, and intraparticle diffusion models.

3.2.3. Adsorption Equilibrium

Figure 5a,b show an increase in the adsorption capacity with an increase in the initial concentration of MB until the capacity plateaus. Figure 5a shows that SBAC-CL with a 1:1 impregnation ratio and activated at 700 °C for 60 min (SBAC-CL3) reached an average adsorption capacity of 263.6 (±3.8) mg/g when the initial MB concentration was 500 mg/L and above. For the same activation conditions, SBAC-KOH3 reached an average capacity of 324.3 (±18) mg/g at an initial MB concentration of 500 mg/L and above (Figure 5b). The figure further shows that MB removal by the different SBAC materials before reaching the plateau capacity is nearly complete, as the percentage removal almost coincides with the line corresponding to a 100% RE.

Figure 6 shows the adsorption isotherms of different sludge-based adsorbents. The figure shows the superiority of SBAC-CL and SBAC-KOH materials over non-activated CS. Increasing the impregnation ratio from 1:4 to 1:1 for both SBAC-CL and SBAC-KOH significantly affected adsorption. Increasing the activation time from 30 to 60 min had a much more positive impact in the case of SBAC-KOH (SBAC-KOH3 versus SBAC-KOH2) relative to SBAC-CL (SBAC-CL3 versus SBAC-CL2). However, increasing the temperature from 700 to 800 °C caused a slight increase in the adsorption ability of SBAC-KOH (SBAC-KOH4 versus SBAC-KOH3) but a decrease in the adsorption ability of SBAC-CL (SBAC-CL4 versus SBAC-CL3). A comparison between the adsorption abilities of SBAC-CL and SBAC-KOH materials under the same activation conditions revealed the latter to be generally superior. An exception was the case with an activation time of 30 min, for which the carbide-lime-activated SBAC performed better, which was consistent with the findings of the rate experiments (Figure 4c). In all cases, the adsorption of MB on the SBAC materials was highly nonlinear.

The results of fitting the Freundlich and Langmuir models to the adsorption isotherm data are presented in

Table 5. Equilibrium parameters for methylene blue based on the Freundlich and Langmuir models.

Adsorbent	qmax-Exp (mg/g)	Freundlich Model			Langmuir Model
		KF (mg1-n/g. Ln)	n	R2	qmax (mg/g)	KL (L/mg)	R2
CS	20.4	13.5	0.12	0.94	19.0	7.28	0.96
SBAC-CL1	112.1	83.2	0.05	0.96	104.5	6.57	0.98
SBAC-CL2	182.6	126.6	0.07	0.93	172.0	2.64	0.98
SBAC-CL3	267.6	139.3	0.11	0.98	242.6	1.79	0.92
SBAC-CL4	182.1	121.1	0.09	0.97	173.7	2.09	0.98
SBAC-KOH1	217.4	153.5	0.08	0.91	216.7	6.03	0.99
SBAC-KOH2	115.9	54.2	0.15	0.98	123.2	0.09	0.99
SBAC-KOH3	345.3	219.6	0.07	0.99	322.8	0.20	0.98
SBAC-KOH4	318.2	212.6	0.10	0.98	292.9	14.27	0.94

. Based on the R² values, both models adequately fit the data, although a better fit was obtained when using the Langmuir model in more cases. The values of the Freundlich exponent suggest highly nonlinear adsorption behavior (n < 0.2 in all cases, indicating a high surface heterogeneity, whereas when n < 1, adsorption is considered favorable [54]). The values of q_max predicted by the Langmuir model were close to the maximum values obtained experimentally (q_max-Exp), with deviations ranging from 0.3 to 9.4%. Figure S2 shows examples of the equilibrium data simulated using the two models.

Although the Langmuir model fits the adsorption isotherm data better, it assumes monolayer coverage of the adsorbate. Given that the projected area of MB was about 1.3–1.35 µm² [55], only 4 × 10⁻⁷ mg of MB could sorb to form a single layer for 1 g of the SBAC material. However, the actual amount sorbed based on the q_max values was several orders of magnitude higher. Therefore, the Freundlich model was more appropriate because it allows for multiple adsorption layers [54]. However, its formulation does not allow for a maximum limit on the uptake capacity of the material (i.e., q_e always increases with an increase in C_o). Hence, both the Freundlich and Langmuir models were used to predict the equilibrium adsorption behavior of the produced SBAC materials; however, neither could explain the adsorption mechanism.

3.2.4. Comparison with Other Adsorbents

The removal of MB by SBAC materials and one of the CACs was compared. The CAC was obtained in a granular form and crushed to a powder (<425 µm) with a size similar to that of the SBAC materials. The characterization of the powdered activated carbon (PAC) revealed a surface area of 998.8 m²/g and a total pore volume of 0.447 cm³/g. Among the tested SBAC-CL materials, SBAC-CL3 with a 1:1 impregnation ratio and activated at 700 °C for 60 min showed the best results for MB removal (Figure 6a). Based on these observations, the performance of SBAC materials prepared under these conditions (SBAC-CL3 and SBAC-KOH3) was compared with that of PAC in removing MB. The uptake of MB by PAC at a low initial concentration (≤200 mg/L) was similar to those of SBAC-CL3 and SBAC-KOH3 (Figure 7). However, as the initial concentration increased beyond 200 mg/L, PAC and SBAC-KOH3 showed superiority over SBAC-CL3. Nonetheless, the results of this study suggest that using carbide lime waste to activate sewage sludge is a viable option, considering its positive environmental impact and expected lower cost, which warrants further exploration.

Table 6. Uptake capacities of different sludge-based adsorbents towards methylene blue.

Study	Raw Material	Activator	SBET (m2/g)	qmax (mg/g)
Calvo et al. [58]	Sewage sludge	H2SO4	390	20
Yu et al. [52]	Sewage sludge and herb residue	H3PO4	842	27
Ferrentino et al. [20]	Sewage sludge (KOH-based hydrochar)	None	31	147
Li et al. [59]	Paper mill sludge	Steam	135 1	159 2
Mahapatra et al. [60]	Food processing sludge	ZnCl2	NA	24
Li et al. [61]	Sewage sludge	ZnCl2	766	95
Ziyang et al. [62]	Sewage sludge	NaClO	423	68
Liu et al. [63]	Sewage sludge and CaSO4	None	14	132
Rozada et al. [64]	Sewage sludge	ZnCl2	472	137
Rozada et al. [64]	Sewage sludge	H2SO4	216	24.5
Hu et al. [56]	Sewage sludge	None	113.5	317
Current study	Sewage sludge	KOH	1068	322.8
Current study	Sewage sludge	Carbide lime	330	242.6

¹ Average of the reported range of 130–140 m²/g. ² Value for the isotherm at 30 °C.

shows the maximum adsorption capacity (q_max) for MB using SBAC or sludge-based hydrochar, as reported by others and found in this study. A direct comparison between the performance of the SBAC material in this study and the adsorbents reported in other studies may not be straightforward given the differences in the raw materials used, the process of adsorbent preparation, and the procedure of the adsorption equilibrium experiments. Nonetheless, the data indicate that the highest MB uptake (approximately 317 mg/g) was attained by the SBAC briquettes reported by Hu et al. [56] and SBAC-KOH in this study (approximately 322.8 mg/g). The latter exhibited the highest BET surface area among the materials listed in Table 6. The second highest MB uptake (242.6 mg/g) was achieved by SBAC-CL in this study. As is evident from the results, a larger surface area does not necessarily result in a higher uptake of MB. For example, sludge-based adsorbents augmented with KOH or CaSO₄, but not activated, showed the lowest surface areas coupled with a relatively high uptake capacity towards MB. In contrast, sludge activated with acids showed relatively large surface areas but the lowest uptake capacity. These differences could be attributed to the ions on the material surface; the KOH- and CaSO₄-modified sludges contain anions that enhance the uptake of cationic MB. In contrast, the surface of acid-activated SBAC is protonated, reducing cationic dye uptake because of the electrostatic repulsive forces between the dye and the adsorbent [57].

3.3. Further Discussion and Future Work

The adsorption of MB onto SBAC has been suggested to be positively influenced by the functional groups on the SBAC surface [20,34,51]. The FTIR spectra of the carbide lime SBAC materials (Figure 1a) show that these materials have higher intensities of functional groups than CS, which could be the reason for their improved adsorption uptake capacity. The intensity of the functional groups on the surface of SBAC-CL materials increased with an increasing impregnation ratio and activation time but decreased with an increasing activation temperature. The same trend was observed for the uptake capacity of SBAC-CL materials towards MB (Figure 6a). However, a comparison of the FTIR spectra (Figure 1b) and MB uptake (Figure 6b) of the SBAC-KOH materials showed different behaviors. First, the functional groups on SBAC-KOH1 had very low intensities compared to those on CS, but the material had a higher adsorption uptake capacity. Second, SBAC-KOH2 had a higher intensity of functional groups than the other SBAC-KOH materials; however, its uptake capacity was the lowest. Third, the SBAC-KOH materials had an imine group at 2090 cm⁻¹, and its peak intensity increased with the increase in the impregnation ratio and activation temperature. This group did not appear in SBAC-CL materials and could be the reason behind the higher adsorption uptake of SBAC-KOH materials relative to that of SBAC-CL materials. Fan et al. [37] suggested that functional groups containing nitrogen (imines, amines, and amides) are essential for MB removal. However, SBAC-KOH2 deviated from this trend, as it exhibited a lower uptake capacity than SBAC-CL2 (Figure 4c). For this material, a peak at 2900 cm⁻¹ was observed that corresponded to methyl and methylene groups. The presence of these groups might have suppressed the uptake of MB by SBAC-KOH2.

Although the intensities of the identified functional groups of the SBAC-CL materials correlated well with their adsorption capacities, such a trend was not evident for the SBAC-KOH materials, and the above-mentioned interpretations are rather speculative. Another factor that could have influenced the extent of adsorption of the SBAC material is related to the surface area and pore volume. SBAC materials contain micropores and mesopores, making them suitable for removing medium-sized molecules, such as MB (1.4 nm), from aqueous solutions [19]. The variations in the uptake capacity of the adsorbents used, including CS, SBAC, and PAC, with surface area and pore volume are shown in Figure 8a,b, respectively. The surface area explained approximately 93.8% of the variation in the maximum uptake capacity, whereas the pore volume explained 97.6% of the variation. The strong correlation between the uptake capacity and surface area or pore volume reflects the importance of surface chemistry. In particular, the presence of anionic surface functional groups such as aromatic rings, carboxylic groups, hydroxyl groups, and other groups plays an essential role in the adsorption of cationic MB. Santoso et al. [34] suggested that MB can interact with anionic groups on the SBAC surface through ionic bonds, hydrogen bridges, covalent bonds, and electron dispersion forces. Ferrentino et al. [20] indicated that alkali activation adds desirable hydroxyl groups to the adsorbent surface, thereby increasing MB adsorption. Yu et al. [52] suggested that the removal of MB by SBAC could be due to interaction with surface functional groups through either a complexation reaction or cation exchange. Adsorption of MB onto the SBAC material could thus be attributed to electrostatic interaction, hydrogen bonding, and hydrophobic partitioning [33,65]. The presence of carboxylic and carbonyl groups on the adsorbent surface promotes both electrostatic interaction and hydrogen bonding, while the presence of aromatic groups promotes hydrophobic partitioning [33,66].

One of the concerns about the use of waste material such as sewage sludge and carbide lime is the potential to leach metallic impurities in the treated water. To assess this, the SBAC materials produced were subjected to mixing for 96 h using the same experimental conditions employed in the batch adsorption study. The concentrations of cations in the solution after mixing were determined using a Varian ICP-OES (Thermoscientific, Bremen, Germany). The results of the leached cations are presented in

Table 7. Concentration of cations in solution (mg/L) after mixing with the SBAC adsorbents ¹.

Material	Na	K	Ca	Mg	Fe	Al	Zn	B	Mn	Cr	Cu
SBAC-CL1	1044	1.212	7.957	2.13	0.023	0.767	0.249	0.133	0.016	0.003	0.003
SBAC-CL2	1396	1.731	17.005	3.345	0.142	0.255	0.161	0.146	0.014	<LOD	0.003
SBAC-CL3	1235	1.109	3.504	3.08	0.062	0.374	0.065	0.159	0.005	0.002	0.004
SBAC-CL4	1077	1.437	8.8	3.577	0.037	0.389	0.058	0.076	0.008	0.003	<LOD
SBAC-KOH1	1093	0.708	1.792	0.303	0.084	0.667	0.033	0.277	0.003	0.007	<LOD
SBAC-KOH2	1192	11.847	1.209	2.49	1.206	<LOD	<LOD	0.118	0.005	0.015	0.011
SBAC-KOH3	1216	12.349	0.536	1.208	0.695	0.079	0.08	0.145	0.025	<LOD	0.003
SBAC-KOH4	1307	9.289	0.638	0.493	3.662	<LOD	0.027	0.127	0.163	0.005	0.003
Buffer	1155	0.161	0.463	0.029	0.039	0.014	0.002	0.065	0.001	0.006	<LOD

¹ LOD (in mg/L) is 0.0558 for Na, 0.0165 for K, 0.0007 for Ca, 0.0002 for Mg, 0.0027 for Fe, 0.0049 for Al, 0.001 for Zn, 0.0018 for B, 0.0004 for Mn, 0.0012 for Cr, and 0.0022 for Cu.

. Sodium is very high in all the solutions exposed to the SBAC materials because it is a main component of the buffer solution used. Potassium is relatively high in SBAC-KOH materials due to the use of a KOH activator, while calcium is relatively high in solutions exposed to SBAC-CL materials. Several heavy metals such as Pb, Cd, Ni, Co, and Ag were found to be below their limit of detection (LOD). Other metals including Al, Zn, Mn, Cr, and Cu were found at trace levels, as shown in Table 7. This suggests that the SBAC materials produced do not pose an environmental concern.

Another issue is related to the potential of the SBAC-CL materials produced to support regeneration. To address this, a regeneration experiment of the adsorbent was carried out, as described by Li et al. [67]. SBAC-CL3 was selected for the reusability test because it gave optimal adsorption among the tested SBAC-CL materials. The experiment involved successive adsorption–desorption tests using triplicate bottles. In the adsorption test, 0.04 g of SBAC-CL3 was added to each bottle and mixed for 1 h with 40 mL of 300 mg/L MB in the buffer solution (Section 2.4). Following that, the bottles were centrifuged at 8000 rpm for 10 min, and the MB aqueous concentrations were measured, while the remainder of the solution was decanted. The desorption test was performed for 1 h using a HCl solution of pH 2. After desorption, the bottles were centrifuged, and the solutions were decanted to start the first cycle of adsorbent regeneration. The reusability experiment involved seven regeneration cycles. The regeneration rate (R) of the SBAC adsorbent was determined as follows:

$$\mathrm{R}=\frac{{\mathrm{q}}_{\mathrm{i}}}{{\mathrm{q}}_{\mathrm{o}}}$$

(12)

where q_o is the adsorption capacity by the fresh SBAC-CL3 (mg/g) and q_i is the adsorption capacity of the material after the i^th cycle of regeneration (mg/g). The results of the regeneration test are shown in Figure 9. The regeneration rate dropped to 90% after seven regeneration cycles, while the adsorption capacity of SBAC-CL3 with a 1 h mixing time dropped from an initial value (q_o) of 215 mg/g to 195 mg/g. These results suggest that the SBAC materials made with carbide lime waste are sustainable.

The results of this study support the potential use of SBAC prepared using carbide lime waste for the removal of MB from solutions. However, additional in-depth research is needed before mainstreaming SBAC for wastewater treatment. Pilot testing should be performed to evaluate the performance of the SBAC materials using actual wastewater samples, to determine the effects of the solution matrix and environmental conditions on the effectiveness of the treatment method. Also, there is a need to conduct a lifecycle assessment of SBAC materials considering their technical effectiveness, regeneration potential, production cost, and environmental aspects.

The findings of this study can be considered a stepping stone towards future investigations aimed at exploring the use of SBAC activated with carbide lime waste for the purification of water or air polluted with other organic or inorganic pollutants. This adsorbent can be used to investigate the removal of emerging organic compounds, heavy metals, and gaseous pollutants from waste streams. Efforts in this direction will promote the reuse of globally increasing amounts of waste materials (sludge and carbide lime waste) for environmental applications and will contribute to the circular economy.

4. Conclusions

In this study, we investigated the use of carbide lime waste to produce SBAC from domestic sewage sludge. MB removal by SBAC materials prepared using carbide lime waste was significantly affected by changes in the activation time, impregnation ratio, and activation temperature. The SBAC-CL materials were high-quality adsorbents with characteristics similar to those of SBAC-KOH materials. The improvement in MB uptake by SBAC-CL could be explained by the presence of anionic functional groups on the SBAC surface. SBAC-CL had a porous surface with irregular channels, which enhanced its adsorption capacity. However, it had a smaller surface area and lower pore volume than SBAC-KOH, while they were much larger and higher than those of CS. The uptake capacity of the produced SBAC materials for MB was strongly correlated with the surface area and pore volume. The results also indicated that SBAC-CL at a 1:1 impregnation ratio, 700 °C activation temperature, and 60 min activation time had an MB uptake capacity that exceeded 240 mg/g. The adsorption of MB by the SBAC materials prepared could be explained by PFO or PSO models; however, the latter performed better. Equilibrium adsorption modeling suggested a highly nonlinear adsorption behavior of MB on the SBAC materials, which the Freundlich and Langmuir models adequately described. Although SBAC-KOH generally showed superior MB uptake compared to SBAC-CL, replacement of the analytical-grade activator with carbide lime waste has the potential to reduce production costs, recycle waste products, and lower the carbon footprint associated with SBAC production. Additional investigations using carbide lime waste as an activator in the production of SBAC are suggested before mainstreaming SBAC for wastewater treatment.