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Environ Eng Res > Volume 24(4); 2019 > Article
Jang, Yoo, Park, and Kan: Engineered biochar from pine wood: Characterization and potential application for removal of sulfamethoxazole in water


The adsorption of sulfamethoxazole (SMX) onto a NaOH-activated pine wood-derived biochar was investigated via batch experiments and models. Surprisingly, the maximum adsorption capacity of activated biochar for SMX (397.29 mg/g) was superior than those of pristine biochars from various feedstock, but comparable to those of commercially available activated carbons. Elovich kinetic and Freundlich isotherm models revealed the best fitted ones for the adsorption of SMX onto the activated biochar indicating chemisorptive interaction occurred on surface of the activated biochar. In addition, the intraparticle diffusion limitation was thought to be the major barrier for the adsorption of SMX on the activated biochar. The main mechanisms for the activated biochar would include hydrophobic, π-π interactions and hydrogen bonding. This was consistent with the changes in physicochemical properties of the activated biochar (e.g., increase in sp2 and surface area, but decrease in the ratios of O/C and H/C).

1. Introduction

Various pharmaceutical compounds released to environments have attracted rising attention due to their adverse effects such as endocrine disruption and prevalence of antibiotic resistance genes [1]. According to recent literatures [2], the low metabolic rates of many pharmaceuticals in human and animals resulted in high environmental risks. Due to high cost-effectiveness, sulfamethoxazole (SMX) is one of the most widely used synthetic sulfonamides antibacterial agents to treat diseases and infections for human [3]. In addition, it is used for feed additives to promote growth rate and weigh gain of food animals [4]. As a consequence, it has been frequently detected in various environmental matrices including soil, sediment, river, surface and ground water [5, 6].
Several remediation technologies such as adsorption [7, 8], ozonation [9], photolysis [10], chemical [3] and electrochemical oxidation [11] are available for elimination of SMX from various water bodies. Adsorption is considered as one of the practical options for removal of SMX mainly due to its simple operation, no generation of harmful byproducts and cost-effectiveness [12]. Recently, adsorption of SMX on various materials has been investigated including clay minerals [13], graphene oxide [14], activated carbons (ACs) [15] and carbon nanotubes (CNTs) [1, 16]. Although some materials such as CNTs and graphene oxide have shown high adsorption capacities for SMX, high costs associated with manufacturing and disposal of these materials limited their practical application in fields [17].
As an alternative adsorbent, biochar (BC), which is a carbonaceous material (CM) produced by pyrolysis of biomass, has received enormous attention due to its low costs and adsorption of various contaminants [7]. Many BCs derived from paper mill sludge [18], bamboo [19], giant reed [8], herb-residue [20], crop straw [21], pine wood [22, 23] and anaerobically digested bagasse [24] have been applied to elimination of SMX. Unfortunately, these BCs showed the limited adsorption capacities for SMX (< 100 mg SMX/g BC). Therefore, possible activation and modification of raw BCs are suggested to enhance their adsorption capacity.
According to Thangalazhy-Gopakumar et al. [25], pine wood is one of the most abundant feedstock in the southern states in U.S. while making it an excellent precursor for BC. To date, a few studies have investigated the adsorption of SMX in water by the pine wood-derived BC (PBC) [22, 23]. However, these studies showed the limited adsorption capacities for SMX and have not investigated the detailed adsorption characteristics such as kinetic and isotherm modeling. The present study reports the detailed adsorption characteristics of pine wood-derived activated BC for SMX. This work also illustrated possible mechanisms for the adsorption of SMX onto the NaOH-activated PBC.

2. Materials and Methods

2.1. Reagents

All chemicals including SMX in the present study were purchased from Sigma-Aldrich Co. (Saint Louis, MO, USA) (Table 1). Based on the solubility of SMX in water at ambient conditions [26], 10–100 mg/L of SMX were used for the adsorption experiments in this study. Four commercial ACs used in the present study were obtained from Sigma-Aldrich Co. (Darco® G-60, Norit® GAC and Norit® GA1) and Calgon Carbon Corp. (Calgon F400).

2.2. Preparation of PBC

Pine wood (Pinus taeda) was used for production of PBC as reported in Jang et al. [27]. Briefly, the air-dried and sieved (20 mesh) pine wood was carbonized using the preheated pyrolysis furnace (MTI Corporation, Richmond, CA, USA) under 300°C for 15 min with 1 L/min of N2 flow. The pristine PBC was designated as “R-PBC”.

2.3. Activation of PBC

Activation of R-PBC was carried out as follows: (1) Soaking 3 g of R-PBC in 40 mL of 4 M NaOH solution for 2 h at room temperature, (2) Pyrolyzing the dried NaOH-treated R-PBC under 800°C for 2 h with a heating rate of 3°C/min while maintaining oxygen limited conditions (2 L/min of N2 flow), and (3) Washing the BC with 0.1 M HCl solution (200 mL) followed by a deionized (DI) water until the pH of BC reached 7.0. The washed and dried NaOH-activated PBC was designated as “A-PBC”.

2.4. Characterization of R- and A-PBCs

The physicochemical properties of the R- and A-PBCs were investigated in detail. Elemental composition analyzer (PerkinElmer 2400 Series II, MA, USA) and scanning electron microscope (SEM) (FEI Verios 460, Hillsboro, OR, USA) were used to evaluate the elemental compositions and the morphology, respectively. Thermogravimetric analysis (TA Instruments, New Castle, DE, USA) was used to evaluate the contents of fixed carbon, volatile compounds, and ash in R- and A-PBCs in accordance with ASTM standard D7582-15 [28]. Fourier transform infrared (FT-IR) spectroscopic analysis was conducted using a FT-IR spectrometer (Bruker Optik GmbH, Ettlingen, Germany). Brunauer-Emmett-Teller (BET) surface area and pore-size distribution were evaluated on the basis of the measurements relating to N2 adsorption with an apparatus at 77 K (Micromeritics Gemini VII 2390p, Norcross, GA, USA). X-ray photoelectron spectroscopy (XPS) analysis was performed using XPS/UPS -SPECS System (SPECS Surface Nano Analysis GmbH, Berlin, Germany) equipped with a monochromatic Mg Kα radiation to quantify carbon-oxygen functional groups. All spectra were acquired from a sample area of 1 mm × 1 mm. The resolution of PHOIBOS 150 photoelectron analyzer is lower than 1 eV. The pH of zero point charges (pHPZC) analysis was conducted as described previously [27]. Briefly, 0.01 M of 50 mL NaCl solution was used to maintain the ion strength and the initial pH of solution was adjusted ranged from 3.0 to 10.0. The final pH of solution was measured after 2 d incubation with 0.01 g of BC at 20°C and 150 rpm. The pHPZC was calculated by the ΔpH = 0.
Energy loss spectroscopy (EELS) was conducted to analyze sp2 content of the BCs (R- and A-PBCs). As described by Jang et al. [27], thin TEM samples of R- and A-PBCs were prepared by using UC7 Ultramicrotome (Leica Microsystems Inc. Buffalo Grove, IL, USA). Electron energy loss spectra for R- and A-PBCs were collected with the electron energy loss spectrometer (Gatan, Pleasanton, CA, USA) attached to probe aberration corrected scanning transmission electron Microscope (FEI Titan 80–300, Hillsboro, OR, USA). For further calculation, the carbon K-edge energy loss spectrum was deconvolved into three Gaussian spectra [29].
The sp2 content is defined as a relative value between the sample 1 s to π* transition peak area ratio and that of graphite as calculated by Eq. (1) [30]:
where the transition at 285.0 eV indicates the electronic transition from carbon 1 s orbital to C = C π* bonding orbital, the second transition at 292.0 eV indicates the electronic transition from carbon 1 s orbital to C − C σ* bonding orbital, and the third transition at 298.0 eV indicates the electronic transition from carbon 1 s orbital to C = C σ* bonding orbital [29, 31, 32].

2.5. Batch Adsorption Experiments

The effects of initial solution pH on adsorption of SMX was conducted at pH 1–10. Before mixing with the BCs, the solution pH was set to the desired pH by adding 0.1 M of HCl or NaOH solution. The foil covered glass bottles were agitated at 150 rpm and 20°C for 5 d. To quantitatively evaluate the contributions of individual SMX species to overall sorption at a tested pH value, the following empirical model was used (Eq. (2)) [33]:
where Kd (L/kg) is overall sorption coefficient and Kd, Kd0 and Kd+ are the sorption coefficients and α, α0 and α+ are the mass fraction for the SMX, SMX0 and SMX+, respectively.
The adsorption kinetic study was conducted by stirring 0.01 g of A-PBC and 0.1 L of SMX solution (20 and 100 mg/L) at pH 4 which was determined as the optimum pH from the above experiments. The aqueous samples were taken at a regular interval for 2 d. In addition, the adsorption experiments for isotherm study were conducted by vigorously stirring 0.1 L of SMX solution (10 to 100 mg/L, pH 4) and 0.01 g of A-PBC for 5 d to reach equilibrium. The SMX concentration in water was evaluated and used for development of an appropriate isotherm model.
During the batch adsorption experiments, the samples were centrifuged and then filtered using a 0.45 μm syringe filter. The concentration of SMX in filtrated solution was measured by a HPLC (LC-2030C model, SHIMADZU Corp., Torrance, CA, USA) with a C-18 column (Aeris PEPTIDE 3.6 μm XB-C18, Phenomenex Inc., Torrance, CA, USA) and a UV-vis detector operating at wave-length of 265 nm. More detailed information about analysis of SMX by a HPLC was described in previously [27].
Adsorption capacity at time t, Qt (mg/g) was calculated by Eq. (3):
where C0 and Ct are the SMX concentration of initial and at time t (mg/L), respectively, V is the volume of SMX solution (L) and M is the g of BC used in study.
The sum of squared error (SSE) was determined according to Eq. (4):
where Qe and Qc are the observed and calculated value of adsorption capacity (mg/g), respectively, and N is the number of measurements.

2.6. Adsorption Kinetic and Isotherm Models

The six adsorption kinetic models including pseudo-first order, pseudo-second order, Elovich, two-compartment first order, intra-particle diffusion and liquid film diffusion can be represented as follows (Eq. (5)(10)):
Pseudo-first order (PFO):Qt=Qe(1-exp(-K1t))
Pseudo-second order (PSO):Qt=K2Qe2t1+K2Qet
Elovich:Qt=(1n)ln ab+(1n)ln t,t0=1ab
Intra-particle diffusion:Qt=Kit+Ci
Liquid file diffusion:ln(1-F)=-Kfdt,F=QtQe
Two-compartment first order:QtQt==Ffast(1-exp-tKfast)+Fslow(1-exp-tKslow)
where Qe = adsorption capacity (mg/g) at equilibrium time, Qt = adsorption capacity (mg/g) at time t (min), K1 = rate constant of pseudo-first order, K2 = rate constant of pseudo-second order, a = rate constant of chemisorption, b = constant of the surface coverage, Ki = intra-particle diffusion rate constant (mg·min0.5/g), Ci = a constant (mg/g), Kfd = adsorption rate constant, Ffast = mass fraction of fast, Fslow = mass fraction of slow, Kfast = first order rate constant for transfer into fast (h−1), Kslow = first order rate constant for transfer into slow (h−1), h = hour.
In this study, three isotherm models including Langmuir, Freundlich and Temkin models were used to fit the experimental data (Eq. (11)(13)):
Langmuir:   Qe=QmKLCe1+KLCe,   RL=11+KLC0
Freundlich:   Qe=KfCe1/nf
Temkin:   Qe=RTbTln(kTCe)
Where Qe = adsorption capacity (mg/g) at equilibrium time, Qm = maximum adsorption capacity (mg/g), KL = Langmuir constant, Ce = equilibrium concentration (mg/L), RL = separation constant, Kf and nf = Freundlich constant, R = universal gas constant (8.314 J/mol), T = temperature in terms of Kelvin, bT = Temkin constant, KT = equilibrium bond constant related to the maximum energy of bond, t = time (min).

3. Results and Discussion

3.1. Effect of Activation on Physicochemical Properties of BC

The physicochemical properties of R-PBC and A-PBC are presented in Fig. 1. For the elemental compositions, the C content increased from 56.4 to 85.8% after activation, whereas the O and H contents decreased from 37.6 to 13.5% and from 5.9 to 0.5%, respectively. With decomposition of O and H under high temperature [27], significant structural changes occurred in A-PBC. As shown in Fig. 2, the ratio of O/C decreased from 0.67 to 0.16 while the ratio of H/C decreased from 0.10 to 5.8 × 10−3, respectively. The decrease in the ratios of O/C and H/C represented decarboxylation (i.e., loss of CO2) and demethylation (i.e., loss of CH3) [34], respectively, suggesting the formation of porous polyaromatic structures with rich carbon content after the NaOH activation of R-PBC. Interestingly, the O/C and H/C of A-PBC were lower than those of other BCs made from grasses, crop residues, manure and sewage sludge, but were close to that of powder activated carbon (Fig. 2). The low O/C and H/C of A-PBC indicated high hydrophobicity and aromaticity in A-PBC which would be closely related to adsorption characteristics of A-PBC for hydrophobic contaminants.
Fig. 1 showed the calculated sp2 content of R-PBC and A-PBC. Interestingly, A-PBC showed higher sp2 content (73.7%) than that of R-PBC (58.0%). Similary, Yoo et al. [30] reported the increase in sp2 content as a function of pyrolysis temperature during the pyrolysis of pine wood. Thus, the increase in sp2 content in A-PBC might be due to the increase in pyrolysis temperature (800°C) during NaOH activation of R-PBC. Also, this was consistent with 13CDP/NMR data (i.e. a growing aromatic structure after NaOH activation of pine BC) which reported by Park et al. [35]. Furthermore, FTIR spectrum of A-PBC exhibited the negligible functional groups and more similar to commercial ACs which have high hydrophobicity (Fig. S1).
After NaOH activation, changes in the morphology of BC surface were observed (Fig. S2). The sponge-like structure in A-PBC indicated the development of micro/meso-pore on the surface. This is highly supported by the data from the textural analysis (i.e., decreased in average of pore width and increased in volume of pores) (Fig. 1). With the well-developed structure, high BET surface area (959.9 m2/g) was achieved in A-PBC. Thus, the activation of PBC increased hydrophobicity and surface area which can lead to more heterogeneous adsorption characteristics [21, 36].

3.2. Effects of Initial pH on Adsorption Capacity

Recently, a strong relationship between SMX species and surface charge of CMs has been reported in the literatures [8, 37]. To take the SMX property into consideration, adsorption capacities (Qe, mg/g) of SMX on the R-PBC and A-PBC were investigated at various initial solution pH (Fig. 3). Both BCs exhibited similar pH-dependent adsorption patterns. The Qe values increased with increasing pH from 1 to 4 and then reached the maximum at pH 4 for both BCs (58.91 mg/g for R-PBC and 437.36 mg/g for A-PBC). Meanwhile, the significantly decreased Qe values were observed at pH higher than 7. Similar results were found from the pH-dependent SMX adsorption on bamboo-derived BC [19].
The higher Kd0 than both Kd+ and Kd at the pH of 2–5 (for R-PBC) and 1–6 (for A-PBC) was observed in both R-PBC and A-PBC (Table S1), indicating the importance of SMX0 species to the overall adsorption. At the pH ranges between 1 and 5, the contribution of SMX0 species to the overall adsorption was higher than 51% (Fig. S3). This was consistent with the data from Zheng et al. [8] who reported significant contribution of SMX0 species to the overall adsorption on giant reed BC. Interestingly, between pH 1 and 5, the contribution of SMX0 species to the overall adsorption on A-PBC was higher than that of R-PBC, suggesting that activation process enhanced hydrophobic interaction as well π-π interaction which was predominant adsorption mechanisms for SMX species on A-PBC [16, 38].
As shown in Fig. 3, the SMX+ and SMX are the dominant species at pH < pKa1 = 1.6 and pH > pKa2 = 5.7, respectively, and SMX0 is dominant at pH between pKa1 and pKa2. In addition, it is assumed that the surface charge of BC was protonated (i.e., positive charge) when the solution pH was lower than the pHPZC of BC, otherwise it was the negative. Considering the pHPZC of R-PBC (5.08) and A-PBC (6.83) (Table S1), a strong electrostatic repulsion between the SMX species and surface of BCs would occur at high pH (> 7) or low pH (< 1) resulting in relatively low value of Qe. However, it is interesting to note that some portion of SMX (72–98 mg/g) was still adsorbed on the A-PBC at high pH (> 7), although a strong electrostatic repulsion would be expected between the SMX and negative charged surface of BCs. This was explained by a strong negative charged assistant hydrogen bond (CAHB) along with oxygen-containing group on surface [20]. Teixidó et al. [33] and Yu et al. [16] reported that -NH2 and -SO2NH- groups in sulfonamides can interact with oxygen-containing group on surface of CMS. Thus, it is hypothesized that A-PBC has a high potential for hydrogen bonding accepting or donating due to its plenty of oxygen-containing group on surface of A-PBC (Table S1). As expected, the significant decrease in Qe values (about 291 mg/g at pH 2 and 104 mg/g at pH 1) were also observed at low pH due to the much less hydrophobicity of A-PBC with abundant ionized forms and the inhibited Lewis acid-base interaction [1].
π-π electron donor-acceptor (EDA) interaction is one of the possible mechanisms for the adsorption of sulfonamides on BC surface [19, 39]. SMX has a strong π-electron acceptor nature due to its p-amino rings (donates loan pair electrons to the benzene ring) and N and/or O-hetero-aromatic rings (contribution to electronic resonance), while the graphite structure π-electron of BC can act as electron donor [16, 19]. The ability for π-π EDA interaction of different SMX species follows the order of SMX+ > SMX0 > SMX due to the decrease in π-withdrawing ability resulted from deprotonation of -SO2NH- groups [40], while increased by protonation of –NH2- groups [41]. At pH ranges between 4 and 6, this was not consistent with the order of Kd (Kd0 > Kd > Kd+) in this study. Thus, π-π EDA interaction was not the major mechanism for adsorption of SMX onto A-PBC at these pH ranges. This was also consistent with the previous research [16] which reported the strong hydrophobic interaction between TC and CNTs rather than π-π EDA interaction. In contrast, previous researches [42, 43] reported the possible π-π EDA interaction between SMX and CMs. Recently, Yoo [44] reported the increasing π-π* transition (plasmon) intensity after the chemical activation of PBC. Similarly, Peng et al. [45] reported that adsorption is determined mainly by the number or areas of aromatic rings in both antibiotics and adsorbent when the main adsorption mechanism is the π-π interaction. Collectively, it is reasonable to conclude that the A-PBC in this study with the high surface area (959.9 m2/g), hydrophobicity and sp2 after the NaOH activation resulted in high SMX adsorption capacity via possible adsorption mechanisms including hydrophobic interaction, π-π EDA interaction and hydrogen bonding.

3.3. Adsorption Kinetics and Isotherms

To understand the possible adsorption mechanisms and rate-limiting step, the kinetic study for adsorption of SMX onto A-PBC was conducted as described in section 2.5 (Fig. 4). Investigation of adsorption kinetics of SMX and A-PBC interaction using PFO, PSO, Elovich and two-compartment first order are summarized in Table 2. Most of studies reported that kinetic data from sulfonamides adsorption experiment was well-fitted by PSO indicating chemisorptive interaction [7]. As shown in Fig. 4 and Table 2, PSO (R2 of 0.91–0.92) was more suitable than PFO (R2 of 0.65) to described the interaction between SMX and A-PBC. In addition, Elovich was the best fitted kinetic model with lowest SSE (3.25–7.96) and highest R2 (0.98) suggesting that strong chemisorptive interaction occurred on energetically heterogeneous surface of A-PBC [46].
Additionally, the time-courses of SMX adsorption data was analyzed using a two-compartment first order model (Fig. 4(a) and (b)). This model well fitted the dynamics of SMX (20 and 100 mg/L) on A-PBC with low SSE (3.18–7.38) and high R2 (0.98). The significantly higher Ffast (0.83–0.85) and Kfast (116.16–163.7) value than those of Fslow (0.15–0.17) and Kslow (0.25–0.26), indicating that the fast sorption stage was predominant during adsorption process. This was supported by the calculated values from Elovich model (i.e. the higher a (adsorption rate) than b (desorption rate)).
In general, transfer of solutes to adsorbents is usually characterized by liquid film diffusion or intraparticle diffusion. In this study, two diffusion models (intra-particle diffusion and film diffusion) were used to gain insight into the rate-limiting steps affecting the kinetics of SMX adsorption onto A-PBC. As shown in Fig. 4(c)–(f), intra-particle diffusion model was more fitted (R2 of 0.84) to the experimental results than liquid film model (R2 of 0.65). Thus, intra-particle diffusion would be considered as the major limitation for the adsorption of SMX onto A-PBC. More detailed corresponding rate constants (i.e., Ki1-Ki3) (Fig. 4(c) and (d)) derived from linear regression represented the highest constant at the first stage (4.27–10.55 g/mg min1/2), while rate constants of 2nd (1.45–3.91 g/mg min1/2) and 3rd stages (0.46–1.27 g/mg min1/2) attributed to the intra-particle diffusion showed low value. Especially, the rate constants of 3rd stage assigned to micropore diffusion was significantly lower than those of the previous two stages. This phenomenon is frequently observed during adsorption of organic contaminants resulted from decreased active adsorption sites on BCs [46].
Several literatures reported that isotherms of SMX adsorption onto various BCs could be fitted into either Langmuir [21] or Freundlich [8, 1820] model. Based on the type of isotherm of A-PBC (i.e., Type I) (Fig. 5), three phenomenological models including Langmuir (no decrease of heat), Freundlich (a logarithmic decrease of heat) and Temkin (a linear decrease of heat) were applied in this study. The results in Fig. 5 showed Freundlich model was more appropriate than Langmuir and Temkin models indicating that the adsorption may take place on a heterogeneous surface with varied affinities. Based on the nf values derived from Freundlich model, the adsorption of SMX was favorable (i.e. the nf value within the range 1–10) onto A-PBC with high energetic heterogeneity. Besides, when the results from the isotherm experiments were fitted to Langmuir and Temkin models, the Langmuir constant (KL = 0.82) and Temkin constant (bT = 57.63) suggested the SMX adsorption onto A-PBC would be favorable (i.e., 0 < RL <1) and adsorption reaction would occur exothermically (i.e., bT > 1) in the concentration range used in this study.

3.4. Comparison of the SMX Adsorption Capacities of Various CMs

Table 3 lists the maximum adsorption capacities for SMX and surface areas of CMs. The Qm value for A-PBC (397.29 mg/g) was higher than those of various CMs including BC, AC, CNT and graphene oxide. For example, Zheng et al. [8] and Sun et al. [21] reported the Qm value of 1.93–4.99 mg/g and 0.25–7.40 mg/g using the BCs from giant reed and crop residues, respectively. In addition, Çalışkan and Göktürk [15] and Chen et al. [14] achieved the Qm value of 185.19 mg/g and 240 mg/g using commercial AC and graphene oxide, respectively. From the statistical results (i.e., Pearson correlation) (Fig. S4), the listed surface area showed positive correlation (r = 0.904, p < 0.01) with Qm value, suggesting that high SMX adsorption capacity on A-PBC could be attributed to its highly increased surface area (959.9 m2/g) with well-developed porous structure after activation.
To evaluate the feasibility of PBC for adsorption of SMX, the batch adsorption experiments were conducted using four commercial ACs (Calgon F400, Darco® G-60, Norit® CA1 and Norit® GAC) under the same conditions (0.1 g/L, 100 mg/L of SMX and pH 4). As shown in Table 3, the Qm of A-PBC (397.29 mg/g) was comparable to four commercial ACs (312.14 mg/g for Calgon F400, 328.83 mg/g for Darco® G-60, 377.5 mg/g for Norit® GAC and Norit® CA1 for 399.94 mg/g). Thus, these results supported the A-PBC would have high potential for removal of SMX in wastewater and water after process optimization.
Recently, Ahmed et al. [19] reported that adsorption of SMX onto BCs decreased when containing competing compounds (i.e., sulfamethazine and sulfathiazole) while it was enhanced with low molecular weight organic acids (e.g., citric- and malic acid) [21]. Thus, further investigations will include adsorption of SMX onto A-PBC in various wastewater containing multiple contaminants. In addition, appropriate regeneration processes (e.g., chemical VS thermal) of the A-PBC should be studied to increase their cost effectiveness.

4. Conclusions

The NaOH-activated PBC (A-PBC) was prepared and it showed highest adsorption capacity for SMX at pH 4 at which SMX (mostly neutral form of SMX) would be adsorbed onto A-PBC via mainly π-π and hydrophobic interactions. The kinetic studies showed the Elovich and intraparticle diffusion models were the most appropriate ones indicating the chemisorption of SMX onto surface of A-PBC with diffusion limitation. The Freundlich model was found to be best fitted isotherm model indicating heterogeneous and multiple layer of adsorption. A-PBC demonstrated its high adsorption capacity for SMX which was comparable with commercial ACs.

Supplementary Information


This work was funded by Texas A&M University Chancellor Research Initiative Fund and the US-DOE Office of Energy Efficiency and Renewable Energy (Award Number DE-EE0006639).


1. Zhao H, Liu X, Cao Z, et al. Adsorption behavior and mechanism of chloramphenicols, sulfonamides, and non-antibiotic pharmaceuticals on multi-walled carbon nanotubes. J Hazard Mater. 2016;310:235–245.

2. Kümmerer K. The presence of pharmaceuticals in the environment due to human use ― Present knowledge and future challenges. J Environ Manage. 2009;90:2354–2366.

3. Qi C, Liu X, Lin C, et al. Degradation of sulfamethoxazole by microwave-activated persulfate: Kinetics, mechanism and acute toxicity. Chem Eng J. 2014;249:6–14.

4. Kumar A, Xagoraraki I. Pharmaceuticals, personal care products and endocrine-disrupting chemicals in US surface and finished drinking waters: A proposed ranking system. Sci Total Environ. 2010;408:5972–5989.

5. Liu J-L, Wong M-H. Pharmaceuticals and personal care products (PPCPs): A review on environmental contamination in China. Environ Int. 2013;59:208–224.

6. Luo Y, Xu L, Rysz M, Wang Y, Zhang H, Alvarez PJ. Occurrence and transport of tetracycline, sulfonamide, quinolone, and macrolide antibiotics in the Haihe River Basin, China. Environ Sci Technol. 2011;45:1827–1833.

7. Peiris C, Gunatilake SR, Mlsna TE, Mohan D, Vithanage M. Biochar based removal of antibiotic sulfonamides and tetracyclines in aquatic environments: A critical review. Bioresour Technol. 2017;246:150–159.

8. Zheng H, Wang Z, Zhao J, Herbert S, Xing B. Sorption of antibiotic sulfamethoxazole varies with biochars produced at different temperatures. Environ Pollut. 2013;181:60–67.

9. Akhtar J, Amin NS, Aris A. Combined adsorption and catalytic ozonation for removal of sulfamethoxazole using Fe2O3/CeO2 loaded activated carbon. Chem Eng J. 2011;170:136–144.

10. Boreen AL, Arnold WA, McNeill K. Photochemical fate of sulfa drugs in the aquatic environment: Sulfa drugs containing five-membered heterocyclic groups. Environ Sci Technol. 2004;38:3933–3940.

11. de Amorim KP, Romualdo LL, Andrade LS. Electrochemical degradation of sulfamethoxazole and trimethoprim at boron-doped diamond electrode: Performance, kinetics and reaction pathway. Sep Purif Technol. 2013;120:319–327.

12. Putra EK, Pranowo R, Sunarso J, Indraswati N, Ismadji S. Performance of activated carbon and bentonite for adsorption of amoxicillin from wastewater: Mechanisms, isotherms and kinetics. Water Res. 2009;43:2419–2430.

13. Gao J, Pedersen JA. Adsorption of sulfonamide antimicrobial agents to clay minerals. Environ Sci Technol. 2005;39:9509–9516.

14. Chen H, Gao B, Li H. Removal of sulfamethoxazole and ciprofloxacin from aqueous solutions by graphene oxide. J Hazard Mater. 2015;282:201–207.

15. Çalışkan E, Göktürk S. Adsorption characteristics of sulfamethoxazole and metronidazole on activated carbon. Sep Sci Technol. 2010;45:244–255.

16. Yu X, Zhang L, Liang M, Sun W. pH-dependent sulfonamides adsorption by carbon nanotubes with different surface oxygen contents. Chem Eng J. 2015;279:363–371.

17. Tabish TA, Memon FA, Gomez DE, Horsell DW, Zhang S. A facile synthesis of porous graphene for efficient water and wastewater treatment. Sci Rep. 2018;8:1817
crossref pdf

18. Calisto V, Ferreira CI, Oliveira JA, Otero M, Esteves VI. Adsorptive removal of pharmaceuticals from water by commercial and waste-based carbons. J Environ Manage. 2015;152:83–90.

19. Ahmed MB, Zhou JL, Ngo HH, Guo W, Johir MAH, Sornalingam K. Single and competitive sorption properties and mechanism of functionalized biochar for removing sulfonamide antibiotics from water. Chem Eng J. 2017;311:348–358.

20. Lian F, Sun B, Song Z, Zhu L, Qi X, Xing B. Physicochemical properties of herb-residue biochar and its sorption to ionizable antibiotic sulfamethoxazole. Chem Eng J. 2014;248:128–134.

21. Sun B, Lian F, Bao Q, Liu Z, Song Z, Zhu L. Impact of low molecular weight organic acids (LMWOAs) on biochar micropores and sorption properties for sulfamethoxazole. Environ Pollut. 2016;214:142–148.

22. Xie M, Chen W, Xu Z, Zheng S, Zhu D. Adsorption of sulfonamides to demineralized pine wood biochars prepared under different thermochemical conditions. Environ Pollut. 2014;186:187–194.

23. Shimabuku KK, Kearns JP, Martinez JE, Mahoney RB, Moreno-Vasquez L, Summers RS. Biochar sorbents for sulfamethoxazole removal from surface water, stormwater, and wastewater effluent. Water Res. 2016;96:236–245.

24. Yao Y, Zhang Y, Gao B, Chen R, Wu F. Removal of sulfamethoxazole (SMX) and sulfapyridine (SPY) from aqueous solutions by biochars derived from anaerobically digested bagasse. Environ Sci Pollut Res. 2018;25:25659–25667.
crossref pdf

25. Thangalazhy-Gopakumar S, Adhikari S, Ravindran H, et al. Physiochemical properties of bio-oil produced at various temperatures from pine wood using an auger reactor. Bioresour Technol. 2010;101:8389–8395.

26. Martínez F, GÓmez A. Estimation of the solubility of sulfonamides in aqueous media from partition coefficients and entropies of fusion. Phys Chem Liq. 2002;40:411–420.

27. Jang HM, Yoo S, Choi Y-K, Park S, Kan E. Adsorption isotherm, kinetic modeling and mechanism of tetracycline on Pinus taeda-derived activated biochar. Bioresour Technol. 2018;259:24–31.

28. ASTM. Standard test methods for proximate analysis of coal and coke by macro thermogravimetric analysis. West Conshohocken: ASTM International; 2016.

29. Marriott A, Hunt A, Bergström E, et al. Investigating the structure of biomass-derived non-graphitizing mesoporous carbons by electron energy loss spectroscopy in the transmission electron microscope and X-ray photoelectron spectroscopy. Carbon. 2014;67:514–524.

30. Yoo S, Kelley SS, Tilotta DC, Park S. Structural characterization of loblolly pine derived biochar by X-ray diffraction and electron energy loss spectroscopy. ACS Sustain Chem Eng. 2018;6:2621–2629.

31. Zhang Z-l, Brydson R, Aslam Z, et al. Investigating the structure of non-graphitising carbons using electron energy loss spectroscopy in the transmission electron microscope. Carbon. 2011;49:5049–5063.

32. Daniels H, Brydson R, Rand B, Brown A. Investigating carbonization and graphitization using electron energy loss spectroscopy (EELS) in the transmission electron microscope (TEM). Philos Mag. 2007;87:4073–4092.

33. Teixidó M, Pignatello JJ, Beltrán JL, Granados M, Peccia J. Speciation of the ionizable antibiotic sulfamethazine on black carbon (biochar). Environ Sci Technol. 2011;45:10020–10027.

34. Zhu X, Liu Y, Zhou C, Luo G, Zhang S, Chen J. A novel porous carbon derived from hydrothermal carbon for efficient adsorption of tetracycline. Carbon. 2014;77:627–636.

35. Park J, Hung I, Gan Z, Rojas OJ, Lim KH, Park S. Activated carbon from biochar: Influence of its physicochemical properties on the sorption characteristics of phenanthrene. Bioresour Technol. 2013;149:383–389.

36. Keiluweit M, Nico PS, Johnson MG, Kleber M. Dynamic molecular structure of plant biomass-derived black carbon (biochar). Environ Sci Technol. 2010;44:1247–1253.

37. Ji L, Liu F, Xu Z, Zheng S, Zhu D. Adsorption of pharmaceutical antibiotics on template-synthesized ordered micro-and mesoporous carbons. Environ Sci Technol. 2010;44:3116–3122.

38. Ji L, Chen W, Duan L, Zhu D. Mechanisms for strong adsorption of tetracycline to carbon nanotubes: A comparative study using activated carbon and graphite as adsorbents. Environ Sci Technol. 2009;43:2322–2327.

39. Chen W, Duan L, Wang L, Zhu D. Adsorption of hydroxyl-and amino-substituted aromatics to carbon nanotubes. Environ Sci Technol. 2008;42:6862–6868.

40. Ji L, Chen W, Zheng S, Xu Z, Zhu D. Adsorption of sulfonamide antibiotics to multiwalled carbon nanotubes. Langmuir. 2009;25:11608–11613.

41. Hansch C, Leo A, Taft R. A survey of Hammett substituent constants and resonance and field parameters. Chem Rev. 1991;91:165–195.

42. Reguyal F, Sarmah AK. Adsorption of sulfamethoxazole by magnetic biochar: Effects of pH, ionic strength, natural organic matter and 17α-ethinylestradiol. Sci Total Environ. 2018;628:722–730.

43. Chen H, Gao B, Li H. Functionalization, pH, and ionic strength influenced sorption of sulfamethoxazole on graphene. J Environ Chem Eng. 2014;2:310–315.

44. Yoo S. Structural characterizations of biomass-derived carbon materials and application as supercapacitor electrode [dissertation]. USA: North Carolina State Univ; 2018.

45. Peng B, Chen L, Que C, et al. Adsorption of antibiotics on graphene and biochar in aqueous solutions induced by π-π interactions. Sci Rep. 2016;6:31920
crossref pdf

46. Zhou Y, Liu X, Xiang Y, et al. Modification of biochar derived from sawdust and its application in removal of tetracycline and copper from aqueous solution: Adsorption mechanism and modelling. Bioresour Technol. 2017;245:266–273.

47. Enders A, Hanley K, Whitman T, Joseph S, Lehmann J. Characterization of biochars to evaluate recalcitrance and agronomic performance. Bioresour Technol. 2012;114:644–653.

48. Harvey OR, Herbert BE, Rhue RD, Kuo L-J. Metal interactions at the biochar-water interface: Energetics and structure-sorption relationships elucidated by flow adsorption microcalorimetry. Environ Sci Technol. 2011;45:5550–5556.

49. Yang X, Xu G, Yu H, Zhang Z. Preparation of ferric-activated sludge-based adsorbent from biological sludge for tetracycline removal. Bioresour Technol. 2016;211:566–573.

50. Inyang M, Gao B, Pullammanappallil P, Ding W, Zimmerman AR. Biochar from anaerobically digested sugarcane bagasse. Bioresour Technol. 2010;101:8868–8872.

51. Gauden PA, Szmechtig-Gauden E, Rychlicki G, Duber S, Garbacz JK, Buczkowski R. Changes of the porous structure of activated carbons applied in a filter bed pilot operation. J Colloid Interface Sci. 2006;295:327–347.

52. Oliveira LCA, Rios RVRA, Fabris JD, Garg V, Sapag K, Lago RM. Activated carbon/iron oxide magnetic composites for the adsorption of contaminants in water. Carbon. 2002;40:2177–2183.

53. Gicheva G, Yordanov G. Removal of citrate-coated silver nanoparticles from aqueous dispersions by using activated carbon. Colloids Surf A. 2013;431:51–59.

Fig. 1
Physical, chemical and textual characteristics of R- and A-PBCs.
Fig. 2
Van Krevelen plot of elemental ratios for R- and A-PBCs. Other BCs or ACs derived from paper mill sludge [18], giant reed [8], herb-residue [20], powered ACs [23], corn stover, dairy manure, poultry manure [47], cordgrass, pine wood [48] and sewage sludge [49] were used to fit in this study. Red arrow indicates the change in properties between R- and A-PBC.
Fig. 3
pH-dependent SMX species and adsorption capacities of Rand A-PBCs.
Fig. 4
Adsorption kinetic of SMX on A-PBC by fitting the pseudo-first order, pseudo-second order, Elovich and two-compartment first order ((a) and (b)), intra-particle model ((c) and (d)) and film diffusion model ((e) and (f)). Two different initial concentration 20 mg/L for (a), (c) and (e) and 100 mg/L for (b), (d) and (f).
Fig. 5
Adsorption isotherms of SMX on A-PBC by fitting the Freundlich, Langmuir and Temkin models.
Table 1
Physical and Chemical Characteristics of Sulfamethoxazole (SMX)
Molecular structure Formula Molecular weight Solubilitya pKa
/upload/thumbnails/eer-2018-358f6.gif C10H11N3O3S 253.28 0.37 g/L pK1 = 1.6, pK2 = 5.7

At ambient temperature [26].

Table 2
Summary of Kinetic Parameters of SMX Adsorption on A-PBC
Pseudo-first order Pseudo-second order

Qeb Qec K1 SSE R2 Qeb K2 SSE R2
20a 191.93 171.08 0.03 12.96 0.65 179.69 2.89 × 10−4 5.54 0.92
100a 430.82 380.13 0.03 32.49 0.65 400.30 9.74 × 10−5 14.3 0.91

Elovich Two-compartment first order

Qec a b t0 SSE R2 Qec Ffast Fslow Kfast Kslow SSE R2

20a 197.79 2.43 × 103 0.07 6.22 × 10−3 3.25 0.98 186.56 0.8281 0.1719 163.7 0.26 3.18 0.98
100a 445.08 1.05 × 103 0.03 3.78 × 10−2 7.96 0.98 417.11 0.8475 0.1525 116.16 0.25 7.38 0.98

Initial concentration of SMX (mg/L)

Observed value

Calculated value

Table 3
Maximum Adsorption Capacity (Qm) of SMX and the Surface Area of Various CMs
Carbonaceous materials Surface area (m2/g) Qm (mg/g) References
Giant reed BC (300°C) 2.09 4.99b Zheng et al. [8]
Giant reed BC (600°C) 58.75 1.93b
Bamboo BC (H3PO4 activated, 600°C) 1.12 88.10b Ahmed et al. [19]
Rice-straw BC (300°C) 5.76 4.21b
Rice-straw BC (600°C) 27.4 7.40b Sun et al. [21]
Wheat-straw BC (300°C) 7.62 6.75b
Wheat-straw BC (600°C) 38.1 0.25b
Primary paper mill sludge BC (800°C) 209.12 1.69b Calisto et al. [18]
Commercial AC (ChemViron Carbon) 848.22 118b
Anaerobically digested bagasse BC (600°C) 17.66c 54.38b Yao et al. [24]
Graphene Oxide - 240b Chen et al. [14]
Coal based AC (Norit®) 851 185.19b Çali kan and Göktürk [15]
Multi-walled CNTs 382 71.8b Zhao et al. [1]
R-PBC 1.4 58.91a
A-PBC 959.9 437.36a, 397.29b
Commercial AC (Calgon F400) 816.3d 312.14a In this study
Commercial AC (Darco® G-60) 933e 328.83a
Commercial AC (Norit® CA1) 980f 399.94a
Commercial AC (Norit® GAC) 1,200f 377.5a

Derived from experimental data;

Derived from isotherm model data;





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