In this study, readily available inexpensive water treatment sludge (WTS) was used to prepare adsorbent for the removal of Congo red (CR) and tetracycline (TC) from aqueous solutions. The structural characteristics and adsorption properties of WTS biochar were characterised via scanning electron microscope, energy dispersive X-ray spectroscopy, Brunauer-Emmett-Teller and Fourier Transform infrared spectroscopy. In batch experiments, the adsorption factors, kinetics, isothermal curves and thermodynamics of the adsorption properties were investigated. The optimum preparation condition of WTS biochar was 400 °C for 4 h under O2-limited pyrolysis, which exhibited increased specific surface area and pore structures. The best adsorption was observed when the pH of the CR and TC solutions was 7 and 4, respectively. The adsorption process followed the pseudo-second-order model, indicating that the main control step was the chemical adsorption process. Isotherm data were best described by the Langmuir model, and the maximum adsorption capacities for CR and TC were 116.4 and 58.5 mg·g−1, respectively. Thermodynamic parameters revealed that the adsorption process was spontaneous and endothermic. According to the analysis, the adsorption mechanism of CR could be attributed to electrostatic attraction, π–π conjugation and hydrogen bonding, whereas that of TC was potentially associated with cation exchange, complex precipitation, π–π conjugation and hydrogen bonding.

  • Synthetic dyes and antibiotics have been widely used and are difficult to biodegrade.

  • In this study, water treatment sludge was prepared, and its adsorption properties of the dye and antibiotics were investigated.

  • It is a promising adsorbent with stable and efficient adsorption capacity, low cost, environmental protection and safety.

Synthetic dyes have been widely used in textile, leather manufacturing and papermaking and other fields. In addition to bringing economic value, dyes have the characteristics of complex chemical structure, high stability and high toxicity. Moreover, most dyes are difficult to biodegrade, and their metabolic intermediates possess mutagenesis and carcinogenesis potential (Sun et al. 2011). Wastewater containing dyes hardly meets emission requirements after biological treatment, which causes acute and chronic problems to the ecological system and become one of the most hard-to-treat kinds of wastewater (Guo et al. 2012). Azo dye Congo red (CR), which has a high loss rate in the process of production and use, is a representative pollutant in printing and dyeing wastewater (Hua et al. 2018; Patel 2018).

Recently, as a type of emerging trace pollutant, antibiotics have been frequently detected in natural water, which has attracted widespread concern (Sui et al. 2012). Tetracycline (TC) has become one of the most widely used antibiotics globally because of its remarkable efficacy and low price. It is difficult for TC to be completely metabolised by human and animal bodies; therefore, it is discharged into water in large quantities (Martins et al. 2015). Existing wastewater treatment systems cannot effectively remove TC (Miao et al. 2004). It is often detected in surface water, soil and vegetables, posing a severe threat to human health (Schmitt et al. 2006; Akiyama & Savin 2010). At present, the primary methods for removing dyes and antibiotics from aqueous solutions include adsorption (Liu et al. 2010; Peng et al. 2012; Markandeya et al. 2021), advanced oxidation (Santos et al. 2015; Yan & Wang 2016), biodegradation and membrane separation (Amorim et al. 2014; Hussein & Scholz 2018). Among these technologies, adsorption is considered an effective and inexpensive method to remove pollutants from wastewater without producing more toxic or difficult-to-degrade pollutants than was begun with. In addition, factors, such as low cost, easy availability, high yield, adsorption performance and availability of adsorbents, must be considered. Hence, it is essential to seek an optimal adsorbent that meets the above requirements.

Water treatment sludge (WTS) is a by-product of coagulation–sedimentation–filtration and other treatment processes in water supply plants. The main components are primarily comprise amorphous masses of iron and aluminium hydroxides. It also contains sediments carried in raw water, humic substances and dissolved colloids (Babatunde & Zhao 2007; Abo-El-Enein et al. 2017). Owing to urbanisation and population growth, the output of WTS is increasing yearly, and the conventional landfill means are unsuitable. Therefore, resource utilisation will become an inevitable trend. WTS is rich in iron and aluminium, has a large specific surface area, a big internal pore size and a high yield and is inexpensive and simple to obtain, which are significant advantages of an adsorbent. Because of its efficient adsorption performance to water pollutants, new applications of WTS in water purification are constantly being tried. In recent years, reuse of WTS as low-cost adsorbent for phosphorus immobilisation represents a large number of investigations (Ahmad et al. 2016; Hidalgo et al. 2017). It has been reported that oxytetracycline is adsorbed by WTS under the optimal condition of 20 g·L−1, and more than 95% oxytetracycline is adsorbed within 2 h (Punamiya et al. 2013). Makris carried out the adsorption and desorption of ClO4 by WTS (Makris et al. 2006). To expand the scope of reuse, WTS has been tested for adsorption of a number of heavy metals, which include Cd, Cu, Pb, Hg (Shen et al. 2019). Nevertheless, there are few reports about using WTS as the source material for dye and antibiotic adsorption. The adsorption properties and mechanism by which WTS adsorbs different types of contaminants have not been clearly elucidated.

In this study, WTS was prepared by an O2-limited heated process, and its adsorption properties for dye and antibiotics were investigated. CR and TC were selected as typical anion dye and antibiotic, respectively. The effects of pH, biochar dosage, initial concentration, temperature and adsorption time on the adsorption effect were investigated, and the physicochemical properties of WTS biochar before and after adsorption of CR and TC were characterised. On this basis, the adsorption mechanism was analysed by the adsorption kinetic, isotherm model and thermodynamic model to provide a theoretical basis and a new approach for the resource utilisation of WTS.

Raw materials and biochar preparation

Congo red and tetracycline were purchased from Tianjin Kermel Chemical Reagent Co., Ltd. The KNO3, NaOH, HCl and other reagents used in the experiment were all analytical reagents and purchased from Sinopharm Chemical Reagent Co., Ltd. Ultrapure water was used in all experiments.

Instruments: muffle furnace (SX2-8-10, Shanghai Institute of Light); acidity meter (PH-3C, Shanghai ray magnetic); UV–vis spectrophotometer (UV − 1800PC, Japan HITACHI); electronic analytical balance (JT2003, Shanghai lichen); gas bath thermostat oscillator (THZ-82, Jintan Yineng experimental instrument factory); Fourier Infrared Spectrometer (Nicolet 670, Thermo Nicolet Corporation); field emission scanning electron microscope (PV77-47170 ME, USA EDAX); specific surface area analyser (ASAP 2460, America Micromeritics).

Raw water treatment sludge (RS) was obtained from a drinking water treatment plant in Chengdu, which used PAC and PAM as flocculants. After being dried in an oven at 105 °C for 24 h, the sludge was ground with a mortar and sifted through a 60-mesh sieve to remove large particles.

An appropriate amount of RS was placed into a crucible. The crucible was then put in a muffle furnace and heated under the O2-limited condition at each of 300, 400, 500, 600 and 700 °C, with the heating rate set at 10 °C·min−1 and holding times were 1, 2, 4 and 6 h. After pyrolysis finished, the biochars were milled with a mortar and pestle to penetrate a 100-mesh sieve and collected to keep in a desiccator for characterisation and sorption experiment. According to different temperature and time conditions, the biochars obtained were named WS300−2, WS400−2, WS500−2, WS600−2, WS700−2, WS400−1, WS400−4 and WS400−6.

Biochar characterisation

The scanning electron microscopy-energy dispersive X-ray spectroscopy (SEM-EDS) provided morphological information and the main chemical composition of the biochar surface. Fourier transform infrared spectroscopy (FTIR) analysis was employed to verify the presence of surface functional groups on the biochar and raw sludge. The spectra ranged from 4,000 to 400 cm−1. The Brunauer–Emmett–Teller (BET) method was used to determine the specific surface area, total pore volume and average pore size of WTS samples.

The zero point charge (pHZPC) of the adsorbent is a key factor affecting the adsorption capacity. In general, pH > pHZPC is favourable for the adsorption of cations, whereas pH < pHZPC is favourable for the adsorption of anions. The zero point charge (pHZPC) of WS400−4 was determined by the following method (Liu & Zhang 2011; Liu et al. 2011):

  • (1)

    A 50-mL sodium chloride (NaCl, 0.01 M) solution was placed into a 100-mL conical flask with a stopper. Afterwards, 0.01 M HCl and NaOH were used to adjust the successive initial solution pH values from 2 to 12, and 0.25 g of the WS400−4 sample was added to the conical flask.

  • (2)

    The conical flask was filled with N2 to eliminate the effect of carbon dioxide on the pH change, and then the mixture was placed in an oscillator at 303 K for 48 h reaction.

  • (3)

    The difference between the final pH and the initial pH, denoted as ΔpH, was plotted against the initial pH. The solution pH at which the curve crossed the line of ΔpH = pH (final) − pH (initial) = 0 was taken as the pHZPC of the sample (Zhu et al. 2014).

Adsorption experiments

0.2 g of RS, WS300−2, WS400−2, WS500−2, WS600−2, WS700−2, WS400−1, WS400−4 and WS400−6 were weighed into a 250 mL conical flask filled with 100 mL of 100 mg·L−1 CR and TC solutions, respectively. The initial pH was unadjusted. Next, all adsorption experiments were performed via a batch technique in a thermostat oscillator at a shaking rate of 150 rpm and 30 °C for 12 h to adsorption equilibrium. After the adsorption, the mixture was filtered by a 0.45-μm inorganic filter membrane and the remaining concentration of CR and TC were determined using a UV–vis spectrophotometer. The equilibrium adsorption capacity (qe, mg/g) and removal rate (R, %) were calculated using Equations (1)–(3) respectively. All batch experiments were performed in triplicate, and the average data were presented.
formula
(1)
formula
(2)
formula
(3)
where C0 (mg·L−1) is the initial concentration of CR and TC; Ce (mg·L−1) is the equilibrium concentration of CR and TC; Ct is the concentration at contact time of t (min); qe (mg·g−1) means that the amount of CR and TC adsorption at equilibrium; qt (mg·g−1) means the amount of CR and TC adsorbed by biochar at contact time of t (min); M (g) is biochar mass; V (L) means the volume solution; R(%) is the removal ratio.

Effect of different factors on adsorption of CR and TC: (1) Initial pH effect: 0.1 M HCl or NaOH was added to the 100 mL of 100 mg·L−1 solutions to adjust the pH to 3–9, then 0.2 g of WS400−4 was weighed into each solution of different pH. In subsequent experiments, the pH of the CR and TC solutions was adjusted to 7 and 4, respectively. (2) Effect of biochar dosage: different dosages of WS400−4 were added to the 100 mg·L−1 CR and TC solutions with liquid–solid ratios ranging from 0.2 to 4 g·L−1. (3) Effect of initial adsorption concentration: 0.2 g of WS400−4 was added to the CR and TC solutions at concentrations of 60–200 mg·L−1. (4) Effect of temperature: 0.2 g of WS400−4 was added to solutions which were placed in a thermostatic oscillator and oscillated for 12 h at 10, 20, 30, 40 and 50 °C to adsorption equilibrium.

Adsorption kinetic analysis

WS400−4 (2 g·L−1) was added to different initial concentrations of CR (100 and 200 mg·L−1) and TC (120 and 180 mg·L−1); then, the adsorption reaction was performed at 30 °Cand 150 rpm. The samples were measured at 0, 5, 10, 15, 20, 30, 40, 60, 120, 180, 240, 360, 480, 600 and 720 min. To understand the adsorption process, three adsorption kinetic models were analysed.

Pseudo-first-order kinetic model
formula
(4)
Pseudo-second-order kinetic model
formula
(5)
Particle diffusion model
formula
(6)
where qt (mg·g−1) and qe (mg·g−1) are the amounts of CR adsorbed by biochar at contact time t and equilibrium, respectively; k1 (min−1), k2 (g·(mg·min)−1), and k3 (mg·(g·min0.5)−1) are the rate constants of pseudo-first-order, pseudo-second-order, and intra-particle diffusion models, respectively; t (min) is the adsorption time; A is a constant related to the thickness of the boundary layer thickness.

Adsorption isotherms and thermodynamic models

According to the parameters observed from the experiments described above, 0.2 g of WS400-4 was added to different initial concentrations of CR (60–200 mg·L−1) and TC (20–200 mg·L−1). The experiment results were analysed with Langmuir and Freundlich models (Equations (7) and (8), respectively) to evaluate the adsorption (Li et al. 2005; Ünlü & Ersoz 2006). Thermodynamic parameters, such as ΔGθ, ΔHθ and ΔSθ, for CR and TC adsorption, were calculated using Equations (9) and (10).
formula
(7)
formula
(8)
where KL (L·mg−1) and KF (mg·g−1) are the constants in Langmuir and Freundlich models, respectively; qm (mg·g−1) is the theoretical maximum of adsorption amount; Ce (mg·L−1) is the equilibrium concentration; 1/n is the heterogeneity factor indicating how favourable the adsorption process is.
formula
(9)
formula
(10)

Plot lnqe/Ce with respect to qe, and then extend to the intercept to obtain the value of lnKd; where R (8.314 J·mol−1·K−1) is the ideal gas constant, T(K) is Kelvin temperature, ΔGθ (kJ·mol−1) is the change in the Gibbs free energy, ΔHθ(kJ·mol−1) is the change in enthalpy, ΔSθ (kJ·mol−1·K−1) is the change in entropy.

Determination of optimal preparation conditions

The adsorption capacity and removal rate of CR and TC by biochar prepared under different conditions are shown in Figure 1. With an increase in preparation temperature from 300 °C to 400 °C, the removal rates of CR and TC of biochar increased slightly. When the temperature increased from 400 °C to 700 °C, the adsorption capacity of CR decreased to 4.05 mg·g−1 and the removal rate was only 8.10%, whereas the adsorption capacity of TC decreases to 16.56 mg·g−1 and the removal rate was 33.14%. It may be that high temperature leads to structure collapse, pore blockage, specific surface area reduction, and adsorption capacity decline (Liu et al. 2007). The fixed reaction temperature was 400 °C, and the removal of pollutants increased first and then decreased with the extension of pyrolysis time. When the reaction temperature was 400 °C and the reaction time was 4 h, the biochar had an optimal adsorption effect on CR and TC. Therefore, 400 °C and 4 h were selected as the best reaction conditions for the preparation of biochar.

Figure 1

Effect of different biochars on CR and TC adsorption.

Figure 1

Effect of different biochars on CR and TC adsorption.

Close modal

Characterisation of adsorbents

Figure 2 presents the SEM micrographs of RS and WS400−4. The surface morphology of RS is a closely connected lamellar structure. After heating treatment, the surface of the lamellar structure became rough. This may be due to the decomposition of organic matter in the WTS by heating, which changed the morphology, increased the pore size and loosened the structure (Nguyen et al. 2019; Wu et al. 2020). Moreover, the stable Si-Al lattice structure was damaged, the surface was rougher and larger pores were formed inside, which was also reflected in the element composition in Figure 3. Compared with RS, the carbon content in WS400−4 decreased significantly, the effective pore size increased, and the contents of inorganic matter, Fe, Al and Ca increased. There was no significant change in the elemental composition of RS and WS400−4, and the main elemental composition was C, O, Na, Al, Si, Ca, Fe, etc., accounting for more than 85% of the total proportion, which agrees with the composition of WTS (Yang et al. 2015). The pore structure analysis results of biochar are shown in Table 1. Under the optimum preparation conditions (temperature 400 °C and reaction time 4 h), the BET surface area increased slightly from 29.39 m2/g to 34.22 m2/g. The increase in the specific surface area increased the active site for CR and TC adsorption. The total pore volume increased from 0.091 cm3/g to 0.127 cm3/g, and the average pore diameter increased by 1.35 times from 7.77 nm to 10.5 nm. It shows that thermal modification of WTS can effectively improve the pore diffusion adsorption capacity. Zero charge point (pHzpc) refers to the pH value of the solution when the net charge on the surface of carbon material is zero. According to pHzpc, we can not only provide a theoretical basis for the adsorption of carbon material, but also predict the adsorption capacity of carbon material under different pH values. Figure 4 depicts the pHzpc of the biochar of WS400−4. When the pH value of the solution is 8.9, the numbers of positive and negative charges of biochar are equal. When the pH value of the solution is less than pHzpc, the surface of biochar is positively charged and can adsorb anions in the solution. Therefore, the higher the pHzpc value is, the more likely the adsorbent is to be positively charged in a wider pH range, which is more conducive to adsorption through electrostatic action.

Table 1

BET analysis of biochars

BiocharsSurface area/(m2·g−1)Total pore volume/(cm3·g−1)Average pore diameter/(nm)
RS 29.39 0.091 7.77 
WS400−4 34.22 0.127 10.5 
BiocharsSurface area/(m2·g−1)Total pore volume/(cm3·g−1)Average pore diameter/(nm)
RS 29.39 0.091 7.77 
WS400−4 34.22 0.127 10.5 
Figure 2

SEM figure (a) (b) RS(raw water treatment sludge); (c) (d) WS400−4.

Figure 2

SEM figure (a) (b) RS(raw water treatment sludge); (c) (d) WS400−4.

Close modal
Figure 3

EDS figure (a) RS; (b) WS400−4.

Figure 3

EDS figure (a) RS; (b) WS400−4.

Close modal
Figure 4

The pHzpc determination of WS400−4.

Figure 4

The pHzpc determination of WS400−4.

Close modal

Effect of solution pH

The pH value of the solution further affects the adsorption process by affecting the adsorbent and the adsorb surface charge and form. The effects of different initial solution pH values on the adsorption of CR and TC are shown in Figure 5. With the increase of pH value from 5 to 7, the CR removal rate of WS400−4 increased from 83.36% to 90.57%. Under acidic conditions, CR ionises and becomes negative, and the surface of WS400−4 is positively charged at this time, due to the increased adsorption of CR by electrostatic attraction. As the solution pH value increases to 9 and tends to be alkaline, the solution pH value is greater than the pHzpc of WS400−4, making the WS400−4 surface negatively charged, thereby repelling CR and significantly reducing its adsorption of CR。

Figure 5

Effect of pH on adsorption.

Figure 5

Effect of pH on adsorption.

Close modal

TC has three dissociation equilibrium constants, pKa1 = 3.3, pKa2 = 7.68, pKa3 = 9.69 (Cao et al. 2018). The morphology and electrification of TC differ for different pH values. The main ionic species of TC were TC+ at pH < 3.3, TC± at 3.3 < pH < 7.68, TC at 7.68 < pH < 9.69 and TC2− at pH > 9.69 (Yan et al. 2020). When the pH of the adsorption system is 3–4, WS400−4 is positively charged on the surface, TC exists in the form of cation at this time, and there is electrostatic repulsion between the two. However, in this study, the removal rate was relatively high and showed an increasing trend, indicating that there may be other forces such as coordination complexation or ion exchange affecting the adsorption process. WS400-4 has high ash content and is rich in mineral cations such as K+, Na+, Fe3+, Mg2+ and Ca2+. It has been reported that cationic TC can be combined with the surface coordination sites of biochar and cation exchange reaction occurs with coordinated cations (Saravanan et al. 2016; Qu et al. 2019), resulting in a high removal rate of TC. When the pH value is 5–8, TC mainly exists in the form of cations and neutral molecules, and the electrostatic repulsion between them inhibits the adsorption of WS400−4 on TC. It has been reported that iron oxide has strong coordination and a complex effect on TC (Zhu et al. 2014). The removal rate of TC decreased significantly with pH increase, probably because, under acidic conditions with low pH value, cations such as iron, calcium and magnesium in biochar are dissolved and the coordination reaction between iron ion and TC occurs, resulting in adsorption or precipitation after polymerisation with TC (Tongaree et al. 1999; Arias et al. 2007; Liang et al. 2019).

As the initial pH of the solution increases from 5 to 7, the ionisation property of TC gradually changes from electropositive to electroneutral, the electrostatic repulsion with adsorbent decreases gradually, and the adsorption capacity increases slightly. However, the adsorption capacity decreases significantly when the pH is higher than 4, indicating that cation exchange and complex precipitation contribute more to adsorption than electrostatic action. At pH = 9, WS400−4 is negatively charged on the surface and repels TC in the form of anionic ions, thereby reducing the removal rate considerably. In summary, the optimal pH conditions for CR and TC adsorption are 7 and 4, respectively. We infer that the adsorption process of CR was electrostatic attraction, whereas the adsorption process of TC was cation exchange, complex precipitation.

Effect of adsorbent dosage

According to Figure 6(a), with an increase in adsorbent dosage, more active adsorption sites were provided, the removal rate of CR increased from 19.18% to 99.39%, and TC rose from 9.92% to 96.72%. In addition, because the total concentrations of pollutants were the same, the increase of adsorbent will reduce the driving force of each adsorption activity point (Wang et al. 2014), thereby making the adsorption amount present a downward trend. The adsorption capacity of CR decreased from 95.90 mg·g−1 to 24.85 mg·g−1, and that of TC decreased from 49.58 mg·g−1 to 24.18 mg·g−1. Considering the economy and removal efficiency, the dosage is 2 g·L−1, and the adsorption capacity of CR and TC at adsorption equilibrium is 45.28 and 28.61 mg·g−1, respectively.

Figure 6

Effects of different variables on adsorption efficiency (a) biochar dosage, (b) CR and TC concentration, (c) temperature, and (d) contact time.

Figure 6

Effects of different variables on adsorption efficiency (a) biochar dosage, (b) CR and TC concentration, (c) temperature, and (d) contact time.

Close modal

Effect of initial concentration and temperature

The effects of initial concentration and temperature are shown in Figure 6(b) and 6(c). The equilibrium adsorption capacity of CR and TC increased with an increase in the initial concentration of CR and TC, and the increase in ambient temperature also promoted the adsorption of CR and TC. When the temperature increased from 10 °C to 50 °C, the equilibrium adsorption capacity of CR increased from 30.40 mg·g−1 to 48.04 mg·g−1, and the removal rate increased from 60.81% to 96.08%. The constant adsorption capacity of TC increased from 18.24 mg·g−1 to 32.86 mg·g−1. Based on these results, when the temperature increased, the diffusion rate of CR and TC increased, thereby increasing the amounts of CR and TC adsorbed and removal rates.

Effect of contact time

The effects of different contact times on the adsorption process are shown in Figure 6(d). The adsorption of CR and TC by WS400−4 shows the characteristics of fast adsorption and slow equilibrium. In the first 4 h, the adsorption capacity increases rapidly and the growth rate tends to be slow from 4 to 12 h. The dye and drug molecules are first rapidly adsorbed on the active sites on the adsorbent surface, and then the system advances to the equilibrium state of adsorption and desorption. Therefore, 12 h was selected as the adsorption equilibrium time point.

Adsorption kinetics

The adsorption kinetic curves of CR and TC onto WS400−4 are presented in Figure 7, and the parameters are shown in Table 2. From the fitting results that the pseudo-second-order model (R2 > 0.99) was superior to the pseudo-first-order model, the theoretical equilibrium adsorption quantity qe.cal calculated was close to the experimental equilibrium adsorption quantity qe.exp, and the relative error was less than 3.89%. Therefore, the pseudo-second-order model can more accurately describe the adsorption process of CR and TC by WS400−4. The pseudo-second-order equation included adsorption processes such as liquid film diffusion, surface adsorption and intra-particle diffusion, and electron sharing or transferring (Fan et al. 2016). The results showed that the main control step of CR and TC adsorption by WS400−4 was the chemical adsorption process, and there was electron transfer between adsorbent and pollutant (Vadivelan & Kumar 2005; Simonin 2016). It can be speculated that the adsorption of CR and TC by WS400−4 was not a single mechanism of action.

Table 2

Adsorption kinetics parameters

Pseudo-first-order kinetic model
Pseudo-second-order kinetics model
Intra-particle diffusion model
k1qe,calR2k2qe,calR2A1k3,1R2A2k3,2R2
CR 100 mg/L 0.0019 11.339 0.8804 0.0027 44.504 0.9999 27.089 1.3128 0.9642 38.701 0.2347 0.9418 
200 mg/L 0.0031 24.422 0.9813 0.0011 69.252 0.9987 32.434 3.1181 0.9289 52.845 0.6466 0.9208 
TC 120 mg/L 0.0025 22.925 0.9721 0.0007 33.898 0.9926 1.4946 2.4654 0.9886 15.6531 0.6996 0.9155 
180 mg/L 0.0021 27.881 0.9674 0.0006 45.455 0.9935 4.9613 3.1854 0.9983 22.8119 0.8934 0.9525 
Pseudo-first-order kinetic model
Pseudo-second-order kinetics model
Intra-particle diffusion model
k1qe,calR2k2qe,calR2A1k3,1R2A2k3,2R2
CR 100 mg/L 0.0019 11.339 0.8804 0.0027 44.504 0.9999 27.089 1.3128 0.9642 38.701 0.2347 0.9418 
200 mg/L 0.0031 24.422 0.9813 0.0011 69.252 0.9987 32.434 3.1181 0.9289 52.845 0.6466 0.9208 
TC 120 mg/L 0.0025 22.925 0.9721 0.0007 33.898 0.9926 1.4946 2.4654 0.9886 15.6531 0.6996 0.9155 
180 mg/L 0.0021 27.881 0.9674 0.0006 45.455 0.9935 4.9613 3.1854 0.9983 22.8119 0.8934 0.9525 
Figure 7

Adsorption kinetic curves of CR and TC onto biochars.

Figure 7

Adsorption kinetic curves of CR and TC onto biochars.

Close modal

To further explore the steps of adsorption speed control, the data of CR and TC were fitted by the intra-particle diffusion model. The graph was divided into two parts, indicating that the whole diffusion process was divided into two steps. The first part was the liquid film diffusion process, and k3,1 was larger, indicating that the boundary diffusion process was faster. The second stage was the internal diffusion process, and k3,2 was smaller, indicating that the internal diffusion process was the rate − limiting step. According to this model, if the line does not pass the origin, it means that intra-particle diffusion is not the only controlling step, and other processes control the reaction rate. According to the relevant results, the line never crossed the origin, indicating that the adsorption rate of CR and TC may be controlled by liquid film diffusion, intra-particle diffusion and surface adsorption (Wei et al. 2019).

Adsorption isotherms and thermodynamics

The adsorption isotherm is the curve of the relationship between the concentration of adsorbate in the adsorbent and the solution when the adsorption equilibrium is reached at a specific temperature (Zhang et al. 2018). Adsorption temperatures were set at 303.15, 313.15 and 323.15 K. Langmuir and Freundlich isothermal model equations were fitted for experimental adsorption data; the results are shown in Figure 8 and Table 3. Both isothermal models can well describe the entire adsorption process at different temperatures. The equilibrium adsorption capacity increased with ambient temperature. The R2 of the Langmuir model was greater than 0.98 and closer to 1 at different temperatures, whereas the R2 of the Freundlich model was between 0.87 and 0.98. The Langmuir model agreed more with the adsorption process of CR and TC by WS400−4. Hence, the distribution of active sites on the adsorbent was relatively uniform, and the adsorption was uniform monolayer adsorption (Wang et al. 2019). At 323.15 K, the maximum adsorption capacity of CR and TC reached 116.41 and 58.58 mg·g−1, respectively. The negative values for ΔGθ at different temperatures indicated that the adsorption of CR and TC occurred in a spontaneous process. In addition, as the temperature increased, a more significant decrease was observed in the ΔGθ values, suggesting that higher temperature can be conducive to the CR and TC adsorption (Niu et al. 2021). The positive value of ΔHθ indicated that the adsorption process was endothermic, and increasing the temperature was beneficial to the adsorption process. The value of ΔSθ was positive, indicating that the entropy of the adsorption process increased, and the adsorption system occurred in the direction of increasing degrees of freedom at the solid-liquid interface (Maszkowska et al. 2014).

Table 3

Isotherm equation and thermodynamic parameters

T/KLangmuir
Freundlich
ΔGθΔHθΔSθ
qmKLR21/nKFR2kJ·mol−1kJ·mol−1J·mol−1·K−1
CR 303.15 69.013 0.1998 0.9951 0.1398 35.119 0.9147 −3.945   
313.15 71.942 0.1905 0.9916 0.1419 35.991 0.8756 −4.177 37.731 136.31 
323.15 116.410 0.1784 0.9845 0.3023 34.687 0.9638 −6.730   
TC 303.15 50.659 0.0409 0.9918 0.5599 3.716 0.9899 −1.851   
313.15 53.248 0.0468 0.9969 0.5212 4.690 0.9789 −2.844 20.162 72.89 
323.15 58.582 0.0525 0.9943 0.5359 5.223 0.9813 −3.297   
T/KLangmuir
Freundlich
ΔGθΔHθΔSθ
qmKLR21/nKFR2kJ·mol−1kJ·mol−1J·mol−1·K−1
CR 303.15 69.013 0.1998 0.9951 0.1398 35.119 0.9147 −3.945   
313.15 71.942 0.1905 0.9916 0.1419 35.991 0.8756 −4.177 37.731 136.31 
323.15 116.410 0.1784 0.9845 0.3023 34.687 0.9638 −6.730   
TC 303.15 50.659 0.0409 0.9918 0.5599 3.716 0.9899 −1.851   
313.15 53.248 0.0468 0.9969 0.5212 4.690 0.9789 −2.844 20.162 72.89 
323.15 58.582 0.0525 0.9943 0.5359 5.223 0.9813 −3.297   
Figure 8

Isothermal adsorption curves of CR and TC onto biochars.

Figure 8

Isothermal adsorption curves of CR and TC onto biochars.

Close modal

Table 4 lists the adsorption effects of different adsorbents on CR and TC in water. By comparing the adsorption effects in the following table, the qmax values of WS400-4 in this study were 116.4 and 58.5 mg·g−1, respectively, which were higher than other biochars. Additionally, the WS400-4 sample do not require modification, the preparation process of WS400-4 is simple and efficient. Therefore, WS400-4 exhibits great potential to become an excellent CR and TC adsorbent.

Table 4

Comparison between various adsorbents used for CR and TC adsorption

Adsorbent raw materialContaminantqmax/mg·g−1Reference
Red mud CR 4.1 Toor & Jin (2004)  
Graphene oxide/chitosan fibres CR 294.0 Du et al. (2014)  
Chitosan hydrobeads CR 92.6 Chatterjee et al. (2007)  
WS400-4 CR 116.4 This study 
Kaolinite TC 47 Zhao et al. (2011)  
Vermiculite TC 36.8 Liu et al. (2017)  
Hydrothermal porous carbon TC 25.44 Zhu et al. (2014)  
WS400-4 TC 58.5 This study 
Adsorbent raw materialContaminantqmax/mg·g−1Reference
Red mud CR 4.1 Toor & Jin (2004)  
Graphene oxide/chitosan fibres CR 294.0 Du et al. (2014)  
Chitosan hydrobeads CR 92.6 Chatterjee et al. (2007)  
WS400-4 CR 116.4 This study 
Kaolinite TC 47 Zhao et al. (2011)  
Vermiculite TC 36.8 Liu et al. (2017)  
Hydrothermal porous carbon TC 25.44 Zhu et al. (2014)  
WS400-4 TC 58.5 This study 

FTIR analysis

The FTIR spectra of RS, WS400−4 and WS400−4 after the adsorption of CR and TC are shown in Figure 9. There was no evident increase or decrease in the types of functional groups before and after thermal modification. The peaks at 3,620, 3,440 and 1,633 cm−1 were the stretching and bending vibration peaks of O-H (Nibou et al. 2011). The peak at 1,440 cm−1 was assigned to the C = C double bond of aromatic organic matter. The intensity of these peaks weakened in WS400−4, indicating the moisture and organic matter content in WTS after pyrolysis were significantly reduced. The Si-O stretching vibration peak near 1,002.8 cm−1 in the intermediate frequency region, the characteristic absorption peak of Si-OH at 798 cm−1, the symmetrical stretching and asymmetric bending vibration peak of Si-O-Si at 746.3 cm−1 and 580 cm−1, and Al-O stretching vibration peak near 476.3 cm−1 basically did not change, indicating that the silicon oxide in WTS still maintained its stable structure after pyrolysis (Mezni et al. 2011; Yang et al. 2014). Some peak positions and heights shifted in the samples containing CR and TC compared with the FTIR spectra of the WS400−4 alone. After adsorption, the band at 3,620 cm−1 in the WS400−4 spectrum was shifted to 3,617 and 3,615 cm−1, the band at 3,430 cm−1 spectrum was shifted to 3,440 and 3,434 cm−1, the band at 1,633 cm−1 was shifted to 1,631 cm−1,and the band at 798 cm−1 was shifted to 796 cm−1 and 790 cm−1, suggesting that hydrogen bonds might form through the interactions of the hydroxyl groups on WS400−4 and −COOH, –NH2 and N − containing heterocyclic ring in the adsorbate (Akar et al. 2015). In addition, the delocalised π bond between TC heterocyclic ring and benzene ring has strong electron-absorbing property, as a π electron acceptor. There may be a conjugation effect between TC and WS400−4 aromatised organic compounds (He et al. 2018). Therefore, after adsorption of TC onto WS400−4, the characteristic peak of the C = C double bond at 1,440 cm−1 on the benzene ring significantly weakened and exhibited a blueshift, indicating that there was a certain π-π stacking effect (Xiao et al. 2015). This phenomenon can also be observed in FTIR spectra after CR adsorption. In general, the efficient adsorption mechanism of WS400−4 was mainly hydrogen bonding, accompanied by π-π stacking.

Figure 9

FTIR spectra of RS and WS400−4 before and after adsorption.

Figure 9

FTIR spectra of RS and WS400−4 before and after adsorption.

Close modal

WS400−4 was produced from WTS through O2-limited pyrolysis, and it displayed high specific surface area and well-developed porous structure. This low-cost adsorbent exhibited superior adsorption performance for both CR and TC. The adsorption process was more accurately described by the pseudo-second-order and Langmuir models. The entire process was jointly controlled by liquid film diffusion, intra-particle diffusion and surface adsorption. CR and TC adsorption onto WS400−4 were spontaneous, endothermic and randomness increase processes. According to the analysis, the adsorption process of CR was dominated by electrostatic attraction, π-π conjugation and hydrogen bonding, whereas the adsorption process of TC was dominated by cation exchange, complex precipitation, π-π conjugation and hydrogen bonding. WTS is a promising adsorbent with stable and efficient adsorption capacity, low cost, environmental protection and safety.

All relevant data are included in the paper or its Supplementary Information.

Abo-El-Enein
S. A.
,
Shebl
A.
&
Abo El-Dahab
S. A.
2017
Drinking water treatment sludge as an efficient adsorbent for heavy metals removal
.
Applied Clay Science
146
,
343
349
.
Ahmad
T.
,
Ahmad
K.
&
Alam
M.
2016
Sustainable management of water treatment sludge through 3‘R’ concept
.
Journal of Cleaner Production
124
,
1
13
.
Akar
S. T.
,
Yilmazer
D.
,
Celik
S.
,
Balk
Y. Y.
&
Akar
T.
2015
Effective biodecolorization potential of surface modified lignocellulosic industrial waste biomass
.
Chemical Engineering Journal
259
,
286
292
.
Amorim
C. L.
,
Maia
A. S.
,
Mesquita
R. B. R.
,
Rangel
A. O. S. S.
,
Loosdrecht
M. C. M.
,
Tiritan
M. E.
&
Castro
P. M. L.
2014
Performance of aerobic granular sludge in a sequencing batch bioreactor exposed to ofloxacin, norfloxacin and ciprofloxacin
.
Water Research
50
,
101
113
.
Arias
M.
,
García-Falcón
M. S.
,
García-Río
L.
,
Mejuto
J. C.
,
Rial-Otero
R.
&
Simal-Gándara
J.
2007
Binding constants of oxytetracycline to animal feed divalent cations
.
Journal of Food Engineering
78
(
1
),
69
73
.
Babatunde
A. O.
&
Zhao
Y. Q.
2007
Constructive approaches toward water treatment works sludge management: an international review of beneficial reuses
.
Critical Reviews in Environmental Science & Technology
37
(
2
),
129
164
.
Chatterjee
S.
,
Chatterjee
S.
,
Chatterjee
B. P.
&
Guhaet
A. K.
2007
Adsorptive removal of Congo red, a carcinogenic textile dye by chitosan hydrobeads: binding mechanism, equilibrium andkinetics
.
Colloids and Surfaces A: Physicochemical and Engineering Aspects
299
,
146
152
.
Guo
J.
,
Chen
S.
,
Liu
L.
,
Li
B.
,
Yang
P.
,
Zhang
L.
&
Feng
Y.
2012
Adsorption of dye from wastewater using chitosan-CTAB modified bentonites
.
Journal of Colloid & Interface Science
382
(
1
),
61
66
.
He
L.
,
Liu
F.
,
Zhao
M.
,
Qi
Z.
,
Sun
X.
,
Muhammad
Z. A.
,
Sun
X.
,
Li
Y.
,
Hao
J.
&
Wang
S.
2018
Electronic-property dependent interactions between tetracycline and graphene nanomaterials in aqueous solution
.
Journal of Environmental Science
66
(
4
),
286
294
.
Hidalgo
A. M.
,
Murcia
M. D.
,
Gomez
M.
,
Gomez
E.
,
Garcia-Izquierdo
C.
&
Solano
C.
2017
Possible uses for sludge from drinking water treatment plants
.
Journal of Environmental Engineering
143
(
3
),
7
16
.
Hussein
A.
&
Scholz
M.
2018
Treatment of artificial wastewater containing two azo textile dyes by vertical-flow constructed wetlands
.
Environmental Science & Pollution Research
25
(
7
),
6870
6889
.
Li
Y. H.
,
Di
Z.
,
Ding
J.
,
Wu
D.
,
Luan
Z.
&
Zhu
Y.
2005
Adsorption thermodynamic, kinetic and desorption studies of Pb2+ on carbon nanotubes
.
Water Research
39
(
4
),
605
609
.
Liang
J.
,
Fang
Y.
,
Luo
Y.
,
Zeng
G.
,
Deng
J.
,
Tan
X.
,
Tang
N.
,
Li
X.
,
He
X.
,
Feng
C.
&
Ye
S.
2019
Magnetic nanoferromanganese oxides modified biochar derived from pine sawdust for adsorption of tetracycline hydrochloride
.
Environmental Science & Pollution Research
26
(
6
),
5892
5903
.
Liu
C. J.
,
Li
Y. Z.
,
Luan
Z. K.
,
Chen
Z. Y.
,
Zhang
Z.
&
Jia
Z. P.
2007
Adsorption removal of phosphate from aqueous solution by active red mud
.
Journal of Environmental Science
19
(
10
),
1166
1170
.
Liu
R.
,
Liu
H.
,
Zhao
X.
,
Qu
J.
&
Zhang
R.
2010
Treatment of dye wastewater with permanganate oxidation and in situ formed manganese dioxides adsorption: cation blue as model pollutant
.
Journal of Hazardous Materials
176
(
1–3
),
926
931
.
Liu
W. J.
,
Zeng
F. X.
,
Jiang
H.
&
Zhang
X. S.
2011
Preparation of high adsorption capacity bio-chars from waste biomass
.
Bioresource Technology
102
(
17
),
8247
8252
.
Liu
S.
,
Wu
P.
,
Yu
L.
,
Li
L.
,
Gong
B.
,
Zhu
N.
,
Dang
Z.
&
Yang
C.
2017
Preparation and characterization of organo-vermiculite based on phosphatidylcholine and adsorption of two typical antibiotics
.
Applied Clay Science
137
,
160
167
.
Markandeya
T.
,
Shukla
S. P.
,
Srivastav
A. L.
&
Reddy
B. R. K. K.
2021
Removal of disperse orange and disperse blue dyes present in textile mill effluent using zeolite synthesized from cenospheres
.
Water Science and Technology
84
,
445
457
.
Martins
A. C.
,
Pezoti
O.
,
Cazetta
A. L.
,
Bedin
K. C.
,
Yamazaki
D. A. S.
,
Bandoch
G. F. G.
,
Asefa
T.
,
Visentainer
J. V.
&
Almeida
V. C.
2015
Removal of tetracycline by NaOH-activated carbon produced from macadamia nut shells:kinetic and equilibrium studies
.
Chemical Engineering Journal
260
,
291
299
.
Maszkowska
J.
,
Wagil
M.
,
Mioduszewska
K.
,
Kumirska
J.
,
Stepnowski
P.
&
Bialk-Bielińska
A.
2014
Thermodynamic studies for adsorption of ionizable pharmaceuticals onto soil
.
Chemosphere
111
,
568
574
.
Mezni
M.
,
Hamzaoui
A.
,
Hamdi
N.
&
Srasra
E.
2011
Synthesis of zeolites from the low-grade Tunisian natural illite by two different methods
.
Applied Surface Science
52
,
209
218
.
Miao
X. S.
,
Bishay
F.
,
Chen
M.
&
Metcalfe
C. D.
2004
Occurrence of antimicrobials in the final effluents of wastewater treatment plants in Canada
.
Environment Science & Technology
38
,
3533
3541
.
Nguyen
V. T.
,
Nguyen
T. B.
,
Chen
C. W.
,
Hung
C. M.
,
Vo
T. D.
,
Chang
J. H.
&
Dong
C. D.
2019
Influence of pyrolysis temperature on polycyclic aromatic hydrocarbons production and tetracycline adsorption behavior of biochar derived from spent coffee ground
.
Bioresource Technology
284
,
197
203
.
Nibou
D.
,
Khemaissia
S.
,
Amokrane
S.
,
Barkat
M.
,
Chegrouche
S.
&
Mellah
A.
2011
Removal of UO22+ onto synthetic NaA zeolite. characterization, equilibriumand kinetic studies
.
Chemical Engineering Journal
172
,
296
305
.
Niu
S.
,
Xie
X.
,
Wang
Z.
,
Zheng
L.
,
Gao
F.
&
Miao
Y.
2021
Enhanced removal performance for Congo red by coal-series kaolin with acid treatment
.
Environmental Technology
42
(
10
),
1472
1481
.
Patel
H.
2018
Charcoal as an adsorbent for textile wastewater treatment
.
Separation Science & Technology
53
(
17
),
2797
2812
.
Peng
H.
,
Pan
B.
,
Wu
M.
,
Liu
Y.
,
Zhang
D.
&
Xing
B.
2012
Adsorption of ofloxacin and norfloxacin on carbon nanotubes: hydrophobicity-and structure-controlled process
.
Journal of Hazardous Materials
233
,
89
96
.
Qu
Z.
,
Wu
Y.
,
Zhu
S.
,
Yu
Y.
,
Huo
M.
,
Zhang
L.
,
Yang
J.
,
Bian
D.
&
Wang
Y.
2019
Green synthesis of magnetic adsorbent using groundwater treatment sludge for tetracycline adsorption
.
Engineering
5
(
5
),
880
887
.
Santos
L. V.
,
Lucilaine
V.
,
Meireles
A. M.
&
Lange
L. C.
2015
Degradation of antibiotics norfloxacin by Fenton,UV and UV/H2O2
.
Journal of Environmental Management
154
,
8
12
.
Saravanan
R.
,
Sacari
E.
,
Gracia
F.
,
Khan
M. M.
,
Mosquera
E.
&
Gupta
V. K.
2016
Conducting PANI stimulated ZnO system for visible light photocatalytic degradation of coloured dyes
.
Journal of Molecular Liquids
221
,
1029
1033
.
Schmitt
H.
,
Stoob
K.
,
Hamscher
G.
,
Smit
E.
&
Seinen
W.
2006
Tetracyclines and tetracycline resistance in agricultural soils: microcosm and field studies
.
Microbial Ecology
51
,
267
276
.
Shen
C.
,
Zhao
Y. Q.
,
Li
W.
,
Yang
Y.
,
Liu
R.
&
Morgen
D.
2019
Global profile of heavy metals and semimetals adsorption using drinking water treatment residual
.
Chemical Engineering Journal
372
,
1019
1027
.
Tongaree
S.
,
Flanagan
D. R.
&
Poust
R. I.
1999
The interaction between oxytetracycline and divalent metal ions in aqueous and mixed solvent systems
.
Pharmaceutical Development & Technology
4
(
4
),
581
591
.
Toor
M.
&
Jin
B.
2004
Adsorption characteristics, isotherm, kinetics, and diffusion of modified natural bentonite forremoving diazo dye
.
Chemical Engineering Journal
99
,
79
88
.
Vadivelan
V.
&
Kumar
K. V.
2005
Equilibrium, kinetics, mechanism, and process design for the sorption of methylene blue onto rice husk
.
Journal of Colloid & Interface Science
286
(
1
),
90
100
.
Wang
Y.
,
Dai
X.
,
Zhan
Y.
,
Ding
X.
,
Wang
M.
&
Wang
X.
2019
In situ growth of ZIF-8 nanoparticles on chitosan to form the hybrid nanocomposites for high-efficiency removal of Congo Red
.
International Journal of Biological Macromolecules
137
,
77
86
.
Wu
J.
,
Yang
J. W.
,
Feng
P.
,
Huang
G. H.
,
Xu
C. H.
&
Lin
B. F.
2020
High-efficiency removal of dyes from wastewater by fully recycling litchi peel biochar
.
Chemosphere
246
,
125734
.
Yan
F. C.
&
Wang
X.
2016
Treatment of dye wastewater using hydrothermally prepared nano-TiO2, under natural light
.
Journal of Inorganic & Organometallic Polymers & Materials
26
(
1
),
142
146
.
Yan
L.
,
Liu
Y.
,
Zhang
Y.
,
Liu
S.
,
Wang
C.
,
Chen
W.
,
Liu
C.
,
Chen
Z.
&
Zhang
Y.
2020
ZnCl2 modified biochar derived from aerobic granular sludge for developed microporosity and enhanced adsorption to tetracycline
.
Bioresource Technology
297
,
122381
.
Yang
L.
,
Wei
J.
,
Zhang
Y.
,
Wang
J.
&
Wang
D.
2014
Reuse of acid coagulant-recovered drinking waterworks sludge residual to remove phosphorus from wastewater
.
Applied Surface Science
305
,
337
346
.
Yang
L.
,
Wei
J.
,
Liu
Z.
,
Wang
J.
&
Wang
D.
2015
Material prepared from drinking waterworks sludge as adsorbent for ammonium removal from wastewater
.
Applied Surface Science
330
,
228
236
.
Zhang
J.
,
Yan
Z.
,
Jing
O.
,
Yang
H.
&
Chen
D.
2018
Highly dispersed sepiolitebased organic modified nanofibers for enhanced adsorption of Congo red
.
Applied Clay Science
157
,
76
85
.
Zhao
Y.
,
Geng
J.
,
Wang
X.
,
Gu
X.
&
Gao
S.
2011
Tetracycline adsorption on kaolinite: pH, metal cations and humic acid effects
.
Ecotoxicology
20
,
1141
1147
.
Zhu
X.
,
Liu
Y.
,
Qian
F.
,
Zhou
C.
,
Zhang
S.
&
Chen
J.
2014
Preparation of magnetic porous carbon from waste hydrochar by simultaneous activation and magnetization for tetracycline removal
.
Bioresource Technology
154
(
2
),
209
214
.
This is an Open Access article distributed under the terms of the Creative Commons Attribution Licence (CC BY 4.0), which permits copying, adaptation and redistribution, provided the original work is properly cited (http://creativecommons.org/licenses/by/4.0/).