Preparation of a Novel Activated Carbon from Cassava Sludge for the High-Efficiency Adsorption of Hexavalent Chromium in Potable Water: Adsorption Performance and Mechanism Insight

Abstract

Particularly, because of the leakage risk of metal elements from sludge carbon, little attention has been focused on using sludge activated carbon as an adsorbent for the removal of Cr (VI) from contaminated water sources. Herein, a novel sludge carbon derived from dewatered cassava sludge was synthesized by pyrolysis using ZnCl₂ as an activator at the optimal conditions. The prepared sludge activated carbon possessed a large BET surface (509.03 m²/g), demonstrating an efficient removal for Cr (VI). Although the time to reach equilibrium was extended by increasing the initial Cr (VI) concentration, the adsorption process was completed within 3 h. The kinetics of adsorption agreed with the Elovich model. The whole adsorption rate was controlled by both film and intra-particle diffusion. The Cr (VI) removal efficiency increased with elevating temperature, and the adsorption equilibrium process followed the Freundlich isotherm model. The adsorption occurred spontaneously with endothermic nature. The removal mechanism of Cr (VI) on the prepared sludge activated carbon depended highly on solution pH, involving pore filling, electrostatic attraction, reduction, and ion exchange. The trace leakage of metal elements after use was confirmed. Therefore, the prepared sludge activated carbon was considered to be a highly potential adsorbent for Cr (VI) removal from contaminated raw water.

1. Introduction

Hexavalent chromium (Cr (VI)) has been widely applied in petroleum, oil refining, electroplating, tanning, metallurgy, fertilizer, and other industrial processes [1,2]. Over the past two decades in many developing countries, raw water sources have been frequently contaminated by wastewater and solid waste containing Cr (VI) due to abnormal discharge [3]. The total Cr (VI) in potable water has been limited within the range of 0.05 mg/L by WHO due to its carcinogenicity, teratogenicity, and mutagenicity [4]. For raw water contaminated by Cr (VI) due to sudden pollution accidents, traditional drinking water treatment processes, comprising coagulation, sedimentation, filtration, and disinfection, are poorly effective for Cr (VI) removal. For example, local potable water treatment plants were closed for nearly ten days due to water pollution caused by Cr (VI) in Qujing city, China [3]. Hence, it is essential to search for a high-efficiency treatment technology to remove Cr (VI) from contaminated raw water sources.

For eliminating Cr (VI) from industrial wastewater, many researchers have reported different technologies, such as bioremediation [5], reduction-coagulation (chemical or electrochemical) [6,7], membrane separation [8], and ion exchange [9]. However, few studies were concerned with the removal of Cr (VI) from raw water sources contaminated by sudden water pollution accidents. In drinking water treatment plants, divalent iron salt, such as ferrous sulfate (FeSO₄) or ferrous chloride (FeCl₂), is often used emergently to remove Cr (VI) from contaminated raw water sources due to sudden water pollution accidents [10]. Firstly, ferrous ion reduces Cr (VI) into Cr (III) at pH 2–3 [11]. Then, Cr (III) is precipitated into chromic hydroxide (Cr (OH)₃) at alkaline conditions and is removed by traditional coagulation, sedimentation, and filtration processes. Although reduction-precipitation with ferrous ions can effectively remove Cr (VI), it is difficult and expensive for a large-scale drinking water treatment plant to adjust the pH of raw water at a wide pH range from strong acidity to alkalinity. Furthermore, during the removal of Cr (VI) with ferrous ion, sludge containing chromium is generated and needs to be dealt with [12]. Activated carbon is also considered to be more favorable to cope with sudden Cr (VI) pollution because of its high specific surface area (BET) and pore volume [13]. At present, generally used commercial activated carbon is usually made of wood or coal [14,15]. This usually makes commercial activated carbon more expensive. Therefore, the development of low-price, green, and environment-friendly activated carbon using easily available materials that are carbon-rich has been a widely-held concern.

Sludge, a by-product generated from sewage treatment, can be used instead of coal or wood to produce activated carbon because of its rich organic matter [16]. This not only can lower the price of the prepared activated carbon but also can realize sludge resource utilization. At present, most sludge activated carbons are produced using urban sludge generated from urban sewage treatment plants [17]. The produced sludge activated carbons above have been applied as adsorbents removing Cr (VI) from industrial wastewater and have demonstrated a high removal efficacy for Cr (VI) [18,19]. However, few sludge activated carbons produced using urban sludge have been used in drinking water treatment because of the risk that heavy metal ions may be released from the sludge activated carbons into the drinking water [1]. Particularly, little attention has been focused on removing Cr (VI) from raw water sources contaminated by sudden water pollution accidents using sludge activated carbon as adsorbent. Therefore, we must seek out a type of sludge containing few heavy metal ions to produce activated carbon as a prospective adsorbent for the removal of Cr (VI) from contaminated raw water sources.

Cassava is widely grown worldwide and is used as raw material to produce fuel ethanol by fermentation [18]. Wastewater containing large amounts of organic matter is created in the process of producing fuel ethanol. Then, dewatered cassava sludge (DCS) is generated from wastewater treatment processes. As an organic industrial sludge, DCS has the advantages of being easy available, low-cost, and carbon-rich, and is a prospective raw material for the preparation of activated carbon. Furthermore, few heavy metals are involved in DCS. This makes it possible that activated carbon produced using DCS as a raw material can be applied to drinking water treatment to avoid the leakage of heavy metal ions. However, as far as we know, there have been few research reports on the preparation of activated carbon from DCS and the application of Cr (VI) removal in contaminated water sources.

Accordingly, we produced a new activated carbon (ACDCS), in which dewatered cassava sludge (DCS) and zinc chloride were used as raw materials and activators, respectively, and optimized the prepared process parameters in this study. Feasibility was investigated for the prepared ACDCS being used to remove Cr (VI) from a contaminated water source; the equilibrium, kinetics, and thermodynamics of Cr (VI) adsorbed on ACDCS were given; and the adsorption mechanisms of Cr (VI) on ACDCS were discussed.

2. Materials and Methods

2.1. Materials and Reagents

DCS collected from an alcohol plant in Taizhou, China, was first dried for 24 h at 378 K. All reagents comprising potassium dichromate (K₂Cr₂O₇), sodium hydroxide (NaOH), hydrochloric acid (HCl), and Zinc chloride (ZnCl₂) were obtained from Sinopharm Chemical Reagent Co., Ltd., Shanghai, China. Only ZnCl₂ was of analytical reagent grade, and the other reagents were all in guaranteed grade.

2.2. Preparation Process of ACDCS and Parameters Optimization

The preparation of ACDCS was achieved by chemical activation using DCS as a raw material and zinc chloride (ZnCl₂) as an activator. The preparation process was as follows. A certain weight of DCS particles with different sizes (20–50, 50–100, 100–200, and >200 mesh) was immersed into a certain volume of ZnCl₂ solution with different concentrations (10%, 20%, 30%, 40%), subsequently activated at 333 K for a period (6, 12, 18, 24 h) in an Electro-thermostatic Water Cabinet (HWS24, Shanghai Yiheng Scientific Instrument Co., Ltd., Shanghai, China). Then, the activated DCS was pyrolyzed and carbonized in a muffle furnace (KSL-1200X, Hefei Kejing Material Technology Co., Ltd., Hefei, China) at the preset temperature (673, 773, 873, 973 K) for the desired retention time (30, 60, 90, 120 min). After cooling to room temperature, the obtained samples were fully pickled with 1.0 mol/L HCl solution, and then repeatedly rinsed with ultrapure water until turned neutral. After washing, the samples were dried, crushed, and ground into powder. The desired products were labeled as ACDCS, where (a) was DCS mesh, (b) was the activation time, (c) was the weight concentration of the ZnCl₂ solution, (d) was the carbonization temperature, and (e) was the carbonization time. For comparison, the unpretreated DCS, using Zinc chloride, produced sludge activated carbon according to a similar process as above, and the product was recorded as CDCS.

The orthogonal test method can optimize the process parameters of producing ACDCS by simplifying the multi-factor and multi-level experiment in this study. Based on literature and the preliminary experimental study, process parameters, including DCS particle size, activation time, weight concentration of ZnCl₂ solution, carbonization temperature, and carbonization time, play a key role in the performance of ACDCS_. Consequently, the parameters above were chosen as factors, each factor was set at four value levels, and an L₁₆(5⁴) orthogonal table was designed. Furthermore, the value of iodine adsorbed on the prepared ACDCS (named as iodine value in this study) as a response, was used as an index for evaluating the performances of the prepared ACDCS.

2.3. Characterizations of ACDCS

The iodine number of ACDCS was determined according to the national standard testing method for wood activated carbon—Determination of iodine number (GB/T 12496.8-2015). The BET surface area and pore characteristics of ACDCS were measured by nitrogen adsorption using a specific surface area and porosity analyzer (ASAP 2460, Micrometrology, Norcross, USA) at 77K. The surface morphologies of the prepared ACDCS were evaluated using SEM (JSM-6490LV, Akishima-shi, Japan) to evaluate the ACDCS elemental constitution. The functional groups on the surface of each sample were analyzed before and after modification in the frequency range of 4000–400 cm⁻¹ by Fourier transform infrared spectroscopy (Nicolet 6700, Nicolet, Madison, MI, USA). In addition, the zeta potentials of the ACDCS were analyzed by a Malvern Zetasizer Nano ZS90 (Malvern Instruments Ltd., Malvern, UK). X-ray photoelectron spectroscopy with the Al K-Alpha spectrometer (Axis Ultra DLD, Shimadzu-Kratos, Manchester, UK) was utilized to reveal the valence of elements and compositions of the samples’ surface. The metal contents in leachate solution from the DCS and ACDCS were detected by an inductively coupled plasma-atomic emission spectrometer (ICP-AES, ThermoFisher, Waltham, MA, USA).

2.4. Adsorption Experiment

A definite amount of K₂Cr₂O₇ (2.829 g) was dissolved in 1 L ultrapure water to prepare Cr (VI) stock solution (1000 mg/L). Cr (VI) solutions with the desired concentration used as model water samples in this study were obtained by diluting stock solution. A certain amount of ACDCS was added into a series of Erlenmeyer flasks containing 500 mL of Cr (VI) solution with the desired concentration. These flasks were sealed and shaken at 150 r/min for 180 min in a water bath shaker. After adsorption, the mixtures were filtered with 0.45 µm micro-membrane, and then the final concentration of Cr (VI) in each filtrate was determined using a UV-visible spectrophotometer at 540 nm. All the experiments were repeated three times.

The effects of adsorbent dosage (0.2–2.5 g), pH (2–13), contact time (0–180 min), and solution temperature (283–323 K) on Cr (VI) removal were investigated. Kinetics studies were carried out by varying initial Cr (VI) concentrations (1, 5, 10 mg/L), and a water sample was drawn at regular time intervals for the concentration of Cr (VI) measurements. Isotherm experiments were carried out at three temperatures (283, 293, 303 K) using varying initial Cr (VI) concentrations (1–15 mg/L). The removal percentage of Cr (VI), E (%), and the amount of Cr (VI) adsorbed on adsorbent at time t, q_t was calculated by using the following expressions:

$$\mathrm{R}(\%){=(\mathrm{C}}_0{-\mathrm{C}}_{\mathrm{t}}{)/\mathrm{C}}_0\times 100\%$$

(1)

$${\mathrm{q}}_{\mathrm{t}}{=(\mathrm{C}}_0{-\mathrm{C}}_{\mathrm{t}})\times \mathrm{V}/\mathrm{W}$$

(2)

where C₀ and C_t (mg/L) are the initial concentration and the final concentration of Cr (VI) at contact time t (min), respectively; R (%) is the removal efficiency of Cr (VI); q_t (mg/L) is the adsorption capacity and at time t (min); W (g) is the weight of ACDCS; and V (L) is the volume of Cr (VI) solution.

3. Results and Discussion

3.1. Optimization of Preparation Parameters of ACDCS

The preparation conditions of ACDCS were optimized by 16 groups of orthogonal experiments, and the results are shown in the supplementary material (demonstrated in Table S1). The iodine values of the prepared ACDCS varied from 274.52 mg/g to 557.10 mg/g at various preparation conditions. The results showed that the different preparation parameters had a considerable influence on the adsorption performances of the prepared ACDCS.

The R value (response at the average of the difference between the extreme values of Ki) can be applied to evaluate the significance of the influence of various factors on the performance of the prepared ACDCS, and the larger the R value, the more considerable the influence of the factor [20]. Based on the viewpoint above, the impact degree of the selected factors in this study was in the order DCS particle size > carbonization temperature > ZnCl₂ weight concentration > activation time > carbonization time. In addition, it was found that the R values of the DCS particle size and carbonization temperature were much higher than those of the other three factors. This indicated that sludge size and carbonization temperature were the key factors affecting the performances of the prepared ACDCS. DCS particle size had the greatest impact on the iodine value. This was because the contents of organic matter and ash varied remarkably for different-sized DCS sludge particles (demonstrated in Table S2). For raw materials used to produce activated carbon, a high content of organic matter and low ash content could significantly improve the adsorption properties of prepared activated carbon [21]. Accordingly, the organic matter content of the raw material was a key factor for producing high-quality activated carbon. Furthermore, carbonization temperature was another crucial parameter for influencing the properties of ACDCS. This was because the carbonization temperature affected the pyrolysis reaction process of sludge and ZnCl₂, thereby improving or decreasing the pore structure of the prepared ACDCS [22].

For a factor, the K_i value (sum response at the ith level) can be applied to recognize the optimal level, and the maximum value of K_i denotes the optimal level [23]. Based on the theory above, the optimal preparation conditions for ACDCS in this study are as follows: DCS particle size 20–50 mesh, ZnCl₂ weight concentration 20%, activation time 12 h, carbonization temperature 773 K, and carbonization time 60 min. The iodine number adsorbed on the prepared ACDCS at the optimal process parameters added to 557.10 mg/g, in contrast to other adsorbents [22,24,25]. The prepared ACDCS at the optimal conditions was applied to the subsequent adsorption experiments and was recorded as ACDCS_optimal.

3.2. Effect of ACDCS_optimal Dosage

The effects of the dosage of DCS (dewatered cassava sludge), CDCS (carbonized dewatered cassava sludge without activation), and ACDCS_optimal on the removal of Cr (VI) were studied, and the results are presented in Figure 1A. Compared with DCS and CDCS, the removal efficacy of Cr (VI) adsorbed on ACDCS_optimal was much better at the same dosage. The SEM images of DCS, CDCS, and ACDCS_optimal can be found in the supplementary material (demonstrated in Figure S2). The surface of DCS was smooth without obvious pores. Although a few folds appeared on the surface of the CDCS, the surface was poor in the pore structure. However, the surface of the ACDCS_optimal was rich in pore structure, resulting in much larger specific surface areas (509.03 m²/g, shown in Table S2) compared with DCS and CDCS. This could be the reason for causing the better removal of the Cr (VI) adsorbed on ACDCS_optimal. Moreover, this indicates that activation with a proper activator (ZnCl₂) during the process of producing activated carbon can obviously elevate the performance of prepared activated carbon [26].

As shown in Figure 1A, the removal efficiency of Cr (VI) rapidly increased from 15.32% to 95.44% with an increasing dosage of ACDCS_optimal from 0.2 to 2.5 g/L. This was because the adsorption sites available increased with the increasing adsorbent dosage. Nevertheless, the amount of Cr (VI) adsorbed on ACDCS_optimal declined from 7.66 to 3.82 mg/g with the increasing adsorbent dosage. This was attributed to the overlapping of active sites at high dosage, causing a decrease in effective contact surface [27]. In addition, at a high dosage, the excessive active sites were not utilized effectively for a certain Cr (VI). This may be another reason causing the decrease in the amount of Cr (VI) adsorbed on ACDCS_optimal with an increase in the adsorbent dosage. Accordingly, the prepared ACDCS_optimal at the optimal level was applied in subsequent tests, and the dosage of ACDCS_optimal was 1.0 g/L in this study, unless otherwise stated.

3.3. Effect of pH

The pH value of the solution, affecting the surface properties of the adsorbent and the chemical form of the adsorbate, is an important parameter of the adsorption process [28]. Figure 1B shows the relationship between the solution’s pH and the removal efficiency of Cr (VI) adsorbed on ACDCS_optimal. The maximum amount of Cr (VI) adsorbed on ACDCS_optimal was found to be 9.84 mg/g, and the removal efficiency of Cr (VI) was 98.43% at pH 2.0. The amount of Cr (VI) adsorbed dwindled from 9.84 mg/g to 0.15 mg/g with the increase in pH from 2.0–10.0. This result suggests the adsorption removal of Cr (VI) is a highly pH-dependent process.

Depending on the pH of the solution and the Cr (VI) concentration, Cr (VI) exists in different forms (H₂CrO₄, HCrO₄⁻, CrO₄²⁻, Cr₂O₇²⁻) (shown in Figure 1C) in aqueous solution [29]. Moreover, the surface charge of ACDCS_optimal (given in Figure 1B) was affected by aqueous solution pH as well [28]. The pH at the point of zero charges (pH_pzc) for ACDCS_optimal was found to be 4.8, indicating that the surface of ACDCS_optimal was charged positively at pH < 4.8 and, conversely, the surface of ACDCS_optimal showed a negative charge at pH > 4.8. At pH = 2, Cr (VI) mainly exists in the anion form of HCrO₄^-. The surface of ACDCS_optimal was loaded with a lot of positive charges due to the protonation, and the strong electrostatic attraction between the positively charged surface of ACDCS_optimal and the negatively charged anion HCrO₄⁻ achieved the highest removal of Cr (VI) at pH = 2.0. The electrostatic attraction decreased gradually due to the fade of the protonation with the increase in pH from 2 to 4.8, leading to a decline in the removal of Cr (VI). After pH > 4.8, the surface of the ACDCS_optimal was charged negatively, and the electrostatic attraction between ACDCS_optimal and HCrO₄⁻ vanished. The reduction of Cr (VI) to Cr (III), and the ion exchange of Cr (III) with adsorption sites at acidic conditions, might play key roles in the removal of Cr (VI) [30]. With an increase in pH from 4.8 to 7.0, the reduction and ion exchange were weakened gradually, resulting in a decline in the removal of Cr (VI). After pH > 7.0, the van der Waals forces probably contributed to ACDCS_optimal adsorption of Cr (VI) removal because of the vanishing of the reduction and electrostatic attraction [31]. CrO₄²⁻ ions were confirmed as the dominant species of Cr (VI) in an aqueous solution at pH > 7.0 (seen in Figure 1C). In comparison with HCrO₄⁻, the stronger repulsion forces between the surface of ACDCS_optimal and CrO₄²⁻ ions decreased the removal of Cr (VI) with the increase in pH from 7.0 to 10.0 [32]. Furthermore, the competition between the excess OH^– and CrO₄²⁻ for the adsorption sites might be another reason causing the decline in the Cr (VI) removal with the increase in pH from 7.0 to 10.0 [31].

3.4. Adsorption Kinetics

3.4.1. Effect of Initial Cr (VI) Concentration and Adsorption Time

Adsorbate concentration and adsorption time are significant parameters for developing cost-effective adsorption procedures [33]. Figure 2A shows the relationship of the amount of Cr (VI) adsorbed on ACDCS_optimal and adsorption time at different initial Cr (VI) concentrations.

It can be easily observed that the adsorption process of Cr (VI) on ACDCS_optimal included three stages: rapid adsorption (0–20 min), slow adsorption (20–150 min), and reaching adsorption equilibrium (180 min). In the initial stage of adsorption, the rapid removal of Cr(VI) was because there were more vacancies on the ACDCS_optimal surface; and, because the concentration gradient between the bulk solution and the ACDCS_optimal surface was larger, the driving force was higher. [34]. However, with the continuous decrease of the density of vacant active centers, the adsorption rate gradually slowed down [35]. At nearly 180 min, the amount of Cr (VI) adsorbed showed no obvious change, indicating that equilibrium was achieved [33].

Moreover, It was also observed that as the initial concentration of Cr(VI) increased from 1.0 mg/L to 10 mg/L, the amount of Cr (VI) adsorbed increased. The active adsorption sites, including the external surface sites and the internal pore sites, were fixed at the ACDCS_optimal dosage of 1 g/L in this study. At a low Cr (VI) concentration (for example, 1.0 mg/L), the active sites of ACDCS_optimal were sufficient for the contents of Cr (VI), and all the Cr (VI) ions were almost adsorbed by the external surface sites. As the initial Cr (VI) concentration gradually increased, the external surface active sites of ACDCS_optimal were completely occupied, and the higher Cr (VI) concentration gradient promoted the continued diffusion of the Cr (VI) ions into the internal pore sites [27]. The sufficient utilization of active sites of ACDCS_optimal might be the reason causing the increase in the amount of Cr (VI) adsorbed with the increase in the initial concentration of Cr (VI) solution [33]. Moreover, Figure 2A demonstrates that the initial concentration of Cr (VI) solution was directly proportional to the time required to reach equilibrium. Compared with the internal pore adsorption at the higher Cr (VI) concentration, the external surface adsorption was much faster at the low Cr (VI) concentration. This could be because as the initial concentration of Cr(VI) increased, it needed to take longer to reach equilibrium.

3.4.2. Kinetics Models

Adsorption kinetics is of major significance for the study of reaction pathways and mechanisms [36]. In order to investigate the adsorption kinetics of Cr (VI) on ACDCS_optimal, the pseudo-first-order model (PFO), pseudo-second-order model (PSO), and Elovich model were used employed to describe the adsorption process of Cr (VI) on ACDCS_optimal in this work [37], and they were expressed by Equations (3)–(5):

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

(3)

$${\mathrm{q}}_{\mathrm{t}}={\mathrm{k}}_2{\mathrm{q}}_{\mathrm{e}}^2\mathrm{t}/\left(1+{\mathrm{k}}_2{\mathrm{q}}_{\mathrm{e}}\mathrm{t}\right)$$

(4)

$${\mathrm{q}}_{\mathrm{t}}={\mathsf{\beta }}^{-1}\ln \left(1+\mathsf{\alpha }\mathsf{\beta }\mathrm{t}\right)$$

(5)

where q_t (mg/g) is the adsorption capacity at time t; k₁ and k₂ (mg/(g min)) are the equilibrium rate constants of the PFO model and the PSO model, respectively; q_e represents the equilibrium adsorption capacity; α is the original adsorption rate constant (mg/(g min)); and β is concerned with the extent of surface coverage and activation energy for chemisorption (mg/g).

The relevant parameters of the three kinetic models with the correlation coefficients at diverse initial solution concentrations of Cr (VI) are presented in

Table 1. The kinetic parameters at different initial Cr (VI) concentrations.

Models and Parameters	1 mg/L	5 mg/L	10mg/L
qeexp	0.99	4.13	5.66
Pseudo-first-order model
qecal (mg/g)	0.93	3.59	5.11
k1 (mg/(g·min))	0.3189	0.1049	0.1609
R2	0.4939	0.6580	0.6855
SSE	0.02216	1.39233	1.37441
Pseudo-second-order model
qecal (mg/g)	0.97	3.95	5.49
k2 (mg/(g·min))	0.6849	0.0384	0.0469
R2	0.8761	0.8901	0.9282
SSE	0.00542	0.44748	0.29537
Elovich model
α (mg/(g·min))	8072.6012	3.9535	39.0889
β (g/mg)	17.084	1.7504	1.6394
R2	0.9844	0.9960	0.9823
SSE	6.83 × 10−4	0.01627	0.1229
Weber-Morris model
kid1 (mg/(g·min0.5))	0.0431	0.3424	0.5247
C1 (mg/g)	0.6846	1.3215	2.2228
R2	0.8553	0.9841	0.9322
kid2 (mg/(g·min0.5))	0.0170	0.1845	0.1920
C2 (mg/g)	0.8042	2.0196	3.5770
R2	0.9225	0.9162	0.9548
Kid3 (mg/(g·min0.5))	0.0076	0.1011	0.1057
C3 (mg/g)	0.8871	2.7024	4.2572
R2	0.9963	0.9263	0.9808
Boyd model
R2	0.9833	0.9624	0.9809

and Figure 2B. The results showed that the correlation coefficients (R²) of the Elovich model were larger than those of the PFO model and PSO model, and the SEE values of the Elovich model were the smallest among the three models. This indicated that the Elovich was the best representation for adsorption of Cr (VI) on ACDCS_optimal. Further, adsorption of Cr (VI) on ACDCS_optimal might be chemisorption by transferring electrons [37].

3.4.3. Diffusion Mechanism

Although the adsorption process of Cr (VI) on ACDCS_optimal was fitted well by the Elovich kinetic model, it still could not confirm the diffusion mechanism and the rate-controlling step in the adsorption process, such as film diffusion and intra-particle diffusion. To identify the step of actual rate-controlling in the adsorption process for Cr (VI), the Weber-Morris model was used [38].

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

(6)

where q_t (mg/g) is the amount of Cr (VI) adsorbed at time t (min), k_id is the intra-particle diffusion rate (mg/(g min^0.5)), and C_i represents the diffusion effects of the boundary layer. Based on the Weber-Morris model, the fitting result of q_t to t^0.5 is linear, and the line passes through the origin if the intra-particle diffusion process is a rate-determining step. Otherwise, the adsorption process could be affected by more steps [39].

The fitted plots of q_t against t^0.5 at different initial Cr (VI) concentrations are shown in Figure 2C, and the parameters of k_id and C_i were calculated according to the model and are listed in Table 1. Obviously, the adsorption data were fitted very well by the three straight lines, suggesting that three stages occurred during Cr (VI) adsorption on ACDCS_optimal [40]. Firstly, the linear section had a higher slope, indicating that Cr (VI) was transferred to the external surface of ACDCS_optimal from an aqueous solution [38]. Then, the slope of the straight section decreased gradually, suggesting that Cr (VI) diffused from the external into the interior pore structure of ACDCS_optimal [27]. Finally, the slope of the straight section approached zero, indicating that the adsorption equilibrium was reached [40]. In addition, the values of C_i at all the stages were greater than zero, indicating that the process was not only intra-particle diffusion but also some other steps that might control the rate of the kinetic process [38]. Further, the Boyd model (given in Equation (7)) was used to check whether the film diffusion was involved in the adsorption process, and the results are represented in Figure 2D and Table 1, respectively.

$$\mathrm{Bt}=-0.4977-\ln \left(1-\frac{{\mathrm{q}}_{\mathrm{t}}}{{\mathrm{q}}_{\mathrm{e}}}\right)$$

(7)

where Bt is the time constant; q_t (mg/g) is the amount of Cr (VI) adsorbed at time t (min); and q_e is the equilibrium adsorption capacity. Based on the Boyd model, the fitting result of B to t⁻¹ is not linear, or the linear does not pass through the origin of the film diffusion process, which is a rate-determining step [39].

It was found that the plots fitted by the Boyd model demonstrated good linearity with a high R² value (0.9833, 0.9624, 0.9809), and the fitting lines were the extended model and did not pass through the origin at different initial Cr (VI) concentrations. Those results confirmed that film diffusion occurred during the adsorption process of Cr (VI) on ACDCS_optimal [39]. To sum up, the adsorption process of Cr (VI) on ACDCS_optimal was mainly affected and controlled by film diffusion and intra-particle diffusion. Firstly, Cr (VI) ions diffused from the bulk solution to the external surface of ACDCS_optimal, and this process was controlled by film diffusion. Then, the next process was dominated by intra-particle diffusion, in which Cr (VI) ions entered the internal pores to load into the internal active adsorption sites until adsorption equilibrium.

3.5. Adsorption Equilibrium

3.5.1. Effect of Temperature

In a solid-liquid adsorption system, solution temperature plays a key role [36]. The equilibrium tests of Cr (VI) adsorbed on ACDCS_optimal were investigated at 283, 293, and 303 K, and the results are shown in Figure 3A. It was observed that the equilibrium adsorption capacity of Cr (VI) on ACDCS_optimal increased slightly with the increase in solution temperature from 283 to 303 K, indicating that the increasing temperature of the solution was conducive to the adsorption removal of Cr (VI). This was because the increase of temperature promoted the thermal movement of Cr (VI) ions in the solution and facilitated the diffusion of Cr (VI) ions into the internal active adsorption sites of ACDCS_optimal [27].

3.5.2. Adsorption Isotherm Models

The adsorption isotherm showed a specific relationship between its accumulation on the surface of the adsorbent and the concentration of adsorbate [39]. Three isotherm models were used to fit the equilibrium experimental data of Cr (VI) on ACDCS_optimal. These isotherm models included the Langmuir model, Freundlich model, and Temkin, and they were expressed by Equations (8)–(10) [41]. The fitted curves and model parameters with R² are presented in Figure 3B–D and

Table 2. Fitting results of adsorption isotherm model of Cr (VI).

Models and Parameters	283 K	293 K	303 K
qeexp	6.76	7.07	8.01
Langmuir model
qm (mg/g)	6.6368	7.0219	7.4593
KL (L/mg)	0.9441	0.9524	1.9104
RL	0.5143–0.0659	0.5121–0.06542	0.3436–0.03372
R2	0.88	0.89	0.90
Freundlich model
KF ((mg/g)/(mg/L)1/n)	3.0958	3.3290	4.3829
1/n	0.3587	0.3515	0.3044
R2	0.98	0.99	0.99
Temkin model
AT (L/g)	27.6142	40.1145	100.0136
BT (J/mol)	2.3527	2.436	2.5191
R2	0.91	0.90	0.92
Dubinin-Radushkevich model
qmax (mg/g)	6.210	6.628	7.678
Kdr (mol2/J2)	5.0555 × 10−8	4.5033 × 10−8	2.5372 × 10−8
E (kJ/mol)	3.14	3.33	4.44
R2	0.87	0.87	0.93

, respectively.

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

(8)

$${\mathrm{q}}_{\mathrm{e}}={\mathrm{K}}_{\mathrm{F}}{\mathrm{C}}_{\mathrm{e}}^{\frac{1}{\mathrm{n}}}$$

(9)

$${\mathrm{q}}_{\mathrm{e}}={\mathrm{B}}_{\mathrm{T}}\ln \left({\mathrm{A}}_{\mathrm{T}}{\mathrm{C}}_{\mathrm{e}}\right)$$

(10)

where q_e is the equilibrium adsorption capacity; q_m is the maximum single-layer adsorption capacity of the adsorbent; K_L is the Langmuir constant; Ce is the equilibrium concentration of Cr (VI) solution; K_F is the Freundlich adsorption constant; n is the adsorption strength; B_T is the heat of adsorption; and A_T is the equilibrium constant.

It was evident that the Freundlich isotherm model showed a better fitting to the equilibrium adsorption data in terms of the coefficient of determination (R² > 0.98). The Freundlich isotherm model is suitable for highly heterogeneous surfaces, and the lack of a saturated platform for Freundlich isotherm may indicate a multilayer adsorption mechanism [39]. Based on this viewpoint, the adsorption process of Cr (VI) on ACDCS_optimal was a multilayer adsorption process over a heterogeneous surface. Moreover, the parameter n was very important for the Freundlich isotherm model. The adsorption type is considered favorable when n is more than 1.0, and the adsorption process occurs very easily when n⁻¹ is less than 0.5. The fitted results confirmed the adsorption process of Cr (VI) on ACDCS_optimal was very easy to carry out.

3.6. Energy Changes during the Adsorption Process

3.6.1. Adsorption Thermodynamics

Adsorption thermodynamics can reflect the energy changes during the process of target pollutants adsorbed on the adsorbent, and can be used to analyze and explain the adsorption mechanism from the perspective of thermodynamics [25]. The thermodynamic parameters of the adsorption of Cr (VI) onto ACDCS_optimal, including the Gibbs free energy change (ΔG°, kJ/mol), changes of adsorption heat (ΔH°, kJ/mol), and changes of adsorption entropy (ΔS°, J/mol/K), were determined by the van’t Hoff approach using Equations (11)–(13) [29]:

$${\mathrm{K}}_{\mathrm{D}}=\frac{{\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{t}}}{{\mathrm{C}}_{\mathrm{t}}}\times \frac{\mathrm{V}}{\mathrm{M}}$$

(11)

$$\mathrm{l}\mathrm{n}{\mathrm{K}}_{\mathrm{D}}=\frac{\mathsf{\Delta }{\mathrm{S}}^{°}}{\mathrm{R}}-\frac{\mathsf{\Delta }{\mathrm{H}}^{°}}{\mathrm{RT}}$$

(12)

$$\mathsf{\Delta }{\mathrm{G}}^{°}=\mathsf{\Delta }{\mathrm{H}}^{°}-\mathrm{T}\mathsf{\Delta }{\mathrm{S}}^{°}$$

(13)

where C₀ and C_t (mg/L) are the initial concentration and the final concentration of Cr (VI) at contact time t (min), respectively; R (8.314 J/(mol·K)) is the universal gas constant; T (K) is the absolute temperature; W (g) is the weight of ACDCS; V (L) is the volume of Cr (VI) solution; and K_D (L/g) is the distribution coefficient of Cr (VI) in ACDCS_optimal and solution.

The plot of ln (K_D) against 1/T is shown in Figure S1. The parameters of ΔH° and ΔS° were determined through the slope and intercept value, respectively. The calculated values of ΔH°, ΔS°, and ΔG° for the adsorption of Cr (VI) on ACDCS_optimal are given in

Table 3. Adsorption thermodynamic parameters of Cr (VI).

T (K)	∆G° (kJ/mol)	∆S° (J/mol/K)	∆H° (kJ/mol)
283	−10.43	36.88	10.13
293	−10.80
303	−11.16
313	−11.53
323	−11.90

.

The negative ΔG° indicates that the process of Cr (VI) adsorbed on ACDCS_optimal was spontaneous. The positive values of ΔS° indicate that the adsorption process is entropy-driven rather than enthalpy. Furthermore, the positive ΔH° confirms the endothermic nature, and that the increasing temperature favored the adsorption process. Moreover, the values of ΔG° indicate the physical or chemical nature of adsorption [25]. The values of ΔG° were <−10 kJ/mol in this study, implying that physisorption was one of the mechanisms for the removal of Cr (VI) adsorbed on ACDCS_optimal.

3.6.2. Average Free Energy

The calculation of average free energy is a method based on the fitting analysis of the Dubinin-Radushkevich (D-R) isotherm model to determine the reaction type [42]. The isotherm model and the average free energy calculation formula were as follows:

$${\mathrm{q}}_{\mathrm{e}}={\mathrm{q}}_{\max }\exp (-{\mathrm{K}}_{\mathrm{dr}}{\mathsf{\epsilon }}^2)=\mathrm{RT} \ln (1+1/{\mathrm{C}}_{\mathrm{e}})$$

(14)

$$\mathrm{E}={\left(2{\mathrm{K}}_{\mathrm{dr}}\right)}^{-0.5}$$

(15)

where q_max is the maximum adsorption capacity of Cr (VI); q_e is the equilibrium adsorption capacity; R (8.314 J/(mol·K)) is the universal gas constant; T (K) is the absolute temperature; Ce is the equilibrium concentration of Cr (VI) solution; K_dr (mol²/J²) is a model parameter related to the average adsorption energy; and E is the average free energy.

The fitted results and model parameters with R² are presented in Table 2. It was evident that the Dubinin-Radushkevich isotherm model showed a good fitting to the equilibrium data in terms of coefficients of determination (R² > 0.87). The average free energy (AFE) for the adsorption of Cr (VI) on ACDCS_optimal at 283 K, 293 K, and 303 K were 3.14 kJ/mol, 3.33 kJ/mol, and 4.44 kJ/mol, respectively. The values of AFE were less than 8 kJ/mol at a different temperature, indicating that the physisorption occurred during the process of Cr (VI) adsorbed on ACDCS_optimal [43].

3.6.3. Adsorption Activation Energy

The activation energy can be used to judge the difficulty of the adsorption process [31]. The Arrhenius equation was used to calculate the adsorption activation energy, expressed by Equations (13):

$$\mathrm{lnk}=\ln \mathrm{A}-{\mathrm{E}}_{\mathrm{a}}/\mathrm{RT}$$

(16)

where E_a is adsorption activation energy (kJ/mol); A is frequency factor; R is the gas constant (8.314 J/(mol·K)); T (K) is the absolute temperature; and k is the reaction rate constant.

The data of ln k versus 1/T were plotted and are shown in Figure S1. The correlation coefficient (R² = 0.97) showed that there was a good linear relationship between a value and 1/T. The weak van der Waals force is the cause of physical adsorption, and its activation energy is lower than 4.18 kJ/mol [31]. In this study, the value of E_a was 3.37 kJ/mol, indicating that the adsorption process of Cr (VI) on ACDCS_optimal was easy to carry out and that physical adsorption was involved.

3.7. Adsorption Mechanism

3.7.1. Pore Filling

The excellent adsorption performances of carbon material are based on its developed pore structure and high specific surface areas, resulting in the fact that pore filling is the most basic adsorption mechanism in the removal of pollutants [28]. The prepared fresh ACDCS_optimal in this study possessed rich pores, and the total pore volume added to 0.375 cm³/g (demonstrated in Table S2). Furthermore, Table S2 demonstrates that the pore sizes were mainly distributed in the range around the micropores and mesopores, leading to a high BET surface area (509.03 m²/g). This was favorable for the removal of Cr (VI) adsorbed on ACDCS_optimal. In addition, the surface pores were covered after adsorption, showing that Cr (VI) ions were adsorbed on the pores (demonstrated in Table S2 and Figure S2). The Cr (VI) ions could diffuse into the internal pores due to the smaller size, compared with the sizes of micropores and mesopores [28]. The decreased BET surface areas and total pore volume (demonstrated in Table S2) after adsorption indicate that the Cr (VI) ions were removed on the internal active sites. The results above confirm that pore filling was one of the mechanisms for the removal of Cr (VI) adsorbed on ACDCS_optimal.

3.7.2. Electrostatic Attraction

As discussed in Section 3.3, the ACDCS_optimal exhibited a positively charged surface resulting from the –OH₂⁺ functional groups because of the protonation at pH < 4.8. In addition, Cr (VI) mainly existed in the form of HCrO₄⁻ ion. The strong electrostatic attraction was formed between –OH₂⁺ and HCrO₄⁻, facilitating the removal of Cr (VI). The electrostatic attraction disappeared after pH > 4.8 due to the ACDCS_optimal surface having negative charges, although Cr (VI) still appeared in the anions (HCrO₄⁻ or CrO₄²⁻).

3.7.3. Adsorption-Combined Reduction

The Cr (VI) adsorbed on ACDCS_optimal might be reduced by the functional groups on the carbon surface at the acidic conditions (shown in Equation (14)). Compared with the fresh ACDCS_optimal, two new peaks at 576.51 eV and 585.88 eV, attributed to Cr2p_3/2 and Cr2p_1/2, respectively, were found in the XPS wide spectra of ACDCS_optimal after adsorption (demonstrated in Figure S3). This indicated chromium was adsorbed on ACDCS_optimal [27]. The two peaks at 579.17 eV and 587.60 eV were ascribed to Cr (VI), and the peaks at 585.88 eV and 576.51 eV were considered as Cr (III) in the high-resolution spectrum of Cr2p [32]. Those results confirmed that the Cr (VI) adsorbed on the ACDCS_optimal was partially reduced to Cr (III) by the rich functional groups on the surface of ACDCS_optimal, such as the –OH and C–H groups (shown in Equations (15) and (16)) [39].

In comparison, the atomic ratio of graphitic carbon after adsorption (C–C and/or C–H) was decreased to 73.88%, and the contents of the C=O and C–O bond increased to 7.58% and 18.53%, respectively [27]. This showed that some graphitic carbon was oxidized by Cr (VI) (shown in Equations (15) and (16)) [39]. The existence of chemical coordination between Cr (VI) and the C–O functional groups was due to a slight shift at the energy of 285.95 eV.

The FTIR spectra of ACDCS_optimal before and after adsorption was also used to analyze the functional groups involved in the redox reaction. The peak, at approximately 3428 cm⁻¹, was associated with the –OH functional groups, and the peak slightly shifted because of Cr (VI) adsorbed. Furthermore, the peaks at 2923 and 2857 cm⁻¹ slightly decreased because of –C–H stretching [25]. These results indicate that the –OH functional groups and graphitic carbon (–C–H) were oxidized because of the reduction of Cr (VI) to Cr (III) [44]. Further, the peak at 1617 cm⁻¹ shifted to 1565 cm⁻¹, which could be associated with stretching vibrations of –COO⁻ [35]. These results suggest that the carboxyl functional groups increased after Cr (VI) removal, and chromium was chemically immobilized on these oxygen-containing groups [39]. Hence, Cr (III), as a reductive product of Cr (VI), was removed by ion exchange with the –COOH functional groups as well pore filling. The characterizations of the above studies are presented in the form of Supplementary Materials (demonstrated in Figure S3).

$${\mathrm{HCrO}}_4^{-}+7{\mathrm{H}}^{+}+3{\mathrm{e}}^{-}\to {\mathrm{Cr}}^{3+}+4{\mathrm{H}}_2\mathrm{O}$$

(17)

−C−H + Cr(VI) + H₂O → −C−OH + Cr(III) + H⁺

(18)

−C−OH + Cr(VI) + H₂ → COOH + Cr(III) + H⁺

(19)

Based on the discussion above, the removal of Cr (VI) with ACDCS_optimal was dependent on the pH of the solution, and the mechanism was considered to comprise the following three sections, as illustrated in Figure S4: (i) the reduction reaction and ion exchange were involved in the removal of Cr (VI), as well pore filling and electrostatic attraction at pH < 4.8; (ii) pH = 6.5, with the electrostatic attraction disappearing due to deprotonation of the oxygen-containing groups, pore filling, reduction reaction, and ion exchange, contributed to the removal of Cr (VI) at 4.8 < pH < 6.5; (iii) pore filling played the main role for the removal of Cr (VI) on ACDCS_optimal at the alkaline conditions.

3.8. Comparison with Other Adsorbents for Cr (VI)

Although it is difficult to compare with various adsorbents due to different test conditions, the maximum adsorption capacity (q_max) is still used as a parameter for comparison. The results in the literature are listed in

Table 4. Comparison with various sludge activated carbons.

Adsorbents	Adsorption Time	qmax (mg/g)	Dosage (g/L)	pH	Reference
Paper mill sludge	3.0 h	23.18	3.5	4.0	[45]
Sludge derived biochar	16.0 h	16.18	1.0	5.0	[46]
Anaerobic granular sludge	2.5 h	13.19	1.5	4.5	[47]
Sewage sludge	24 h	11.56	4.0	2.0	[48]
ACDCSoptimal	3.0 h	9.84	1.0	2.0	This study
Peanut shell	16.0 h	8.31	2.5	2.0	[29]
Municipal sludge	8.0 h	7.0	1.25	2.0	[17]
Biochars	12.0 h	6.08	2.0	1.75	[49]
Distillery sludge	2.0 h	5.7	5.0	3.0	[50]
Oak wood	48.0 h	5.50	10	2.0	[51]

. Table 4 demonstrates that it was comparable for the prepared ACDCS_optimal in this study with the shorter adsorption time and the higher maximum adsorption capacity. In addition, the trace metal leakage (demonstrated in Table S3) indicated that the prepared ACDCS_optimal was a potential adsorbent for the removal of Cr (VI) from the contaminated raw water.

4. Conclusions

A novel sludge activated carbon was produced successfully by pyrolysis using dewatered cassava sludge as the raw material and ZnCl₂ as the activator. The removal performances and mechanisms of Cr (VI) on the prepared sludge activated carbon were systemically investigated. The optimal preparation parameters were as follows: DCS particle size 20–50 mesh, ZnCl₂ weight concentration 20%, activation time 12 h, carbonization temperature 773 K, and carbonization time 60 min. The specific surface areas and iodine value for the prepared sludge activated carbon (named as ACDCS_optimal) at the optimal process parameters added to 509.03 m²/g and 557.10 mg/g, respectively. The initial Cr (VI) concentration affected the time of reading equilibrium, but the adsorption could be fulfilled within 3 h. The kinetics could be fitted well by the Elovich model. The whole adsorption rate was controlled by both film and intra-particle diffusion. The adsorption equilibrium could be described well by the Freundlich isotherm model. The maximum equilibrium capacity was 8.01 mg/g. The adsorption was a spontaneous and endothermic process. Elevating temperature facilitated the removal of Cr (VI) adsorbed on ACDCS_optimal. The removal of Cr (VI) on ACDCS_optimal was dependent on the pH of the solution. The pore filling, electrostatic attraction, reduction, and ion exchange were involved in the removal mechanism of Cr (VI) adsorbed on ACDCS_optimal. The metal leakage contents were traced. Accordingly, the prepared ACDCS_optimal was a potential adsorbent for the removal of Cr (VI) from contaminated raw water.