Enhanced sorption of trivalent antimony by chitosan

Journal of Hazardous Materials 425 (2022) 127971
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Journal of Hazardous Materials
journal homepage: www.elsevier.com/locate/jhazmat
Enhanced sorption of trivalent antimony by chitosan-loaded biochar in
aqueous solutions: Characterization, performance and mechanisms
Hanbo Chen a, b, Yurong Gao a, b, Ali El-Naggar c, Nabeel Khan Niazi d, Chenghua Sun e,
Sabry M. Shaheen f, g, h, Deyi Hou i, Xing Yang b, f, Zhiyuan Tang j, Zhongzhen Liu k, Hong Hou l,
Wenfu Chen a, Jörg Rinklebe f, m, Michael Pohořelý n, o, Hailong Wang a, b, *
a
Agronomy College, Shenyang Agricultural University, Shenyang, Liaoning 110866, China
Biochar Engineering Technology Research Center of Guangdong Province, School of Environmental and Chemical Engineering, Foshan University, Foshan, Guangdong
528000, China
c
Department of Soil Sciences, Faculty of Agriculture, Ain Shams University, Cairo 11241, Egypt
d
Institute of Soil and Environmental Sciences, University of Agriculture Faisalabad, Faisalabad 38040, Pakistan
e
Department of Chemistry and Biotechnology, Center for Translational Atomaterials, Swinburne University of Technology, Hawthorn, VIC 3122, Australia
f
University of Wuppertal, School of Architecture and Civil Engineering, Institute of Foundation Engineering, Water, and Waste-Management, Laboratory of Soil, and
Groundwater-Management, Pauluskirchstraße 7, 42285 Wuppertal, Germany
g
King Abdulaziz University, Faculty of Meteorology, Environment, and Arid Land Agriculture, Department of Arid Land Agriculture, Jeddah 21589, Saudi Arabia
h
University of Kafrelsheikh, Faculty of Agriculture, Department of Soil and Water Sciences, 33 516 Kafr El-Sheikh, Egypt
i
School of Environment, Tsinghua University, Beijing 100084, China
j
Foshan Xincheng Landscaping Engineering Co., Ltd., Huakang Road, Lecong, Shunde District, Foshan, Guangdong 528315, China
k
Institute of Agricultural Resources and Environment, Guangdong Academy of Agricultural Sciences, Guangzhou 510640, China
l
State Key Laboratory of Environmental Criteria and Risk Assessment, Chinese Research Academy of Environmental Sciences, Beijing 100012, China
m
University of Sejong, Department of Environment, Energy and Geoinformatics, Guangjin-Gu, Seoul 05006, Republic of Korea
n
Institute of Chemical Process Fundamentals of the Czech Academy of Sciences, v. v. i., Rozvojová 135, 165 02 Prague 6-Suchdol, Czech Republic
o
Department of Power Engineering, Faculty of Environmental Technology, University of Chemistry and Technology Prague, Technická 5, 166 28 Prague 6, Czech
Republic
b
A R T I C L E I N F O
A B S T R A C T
Keywords:
Adsorption
Heavy metal
Biochar modification
Theoretical calculation
Contaminated water
Contamination of aquatic systems by antimony (Sb) is a worldwide issue due to its risks to eco-environment and
human health. Batch sorption experiments were conducted to assess the equilibrium, kinetics and thermody­
namics of antimonite [Sb(III)] sorption by pristine biochar (BC) and chitosan-loaded biochar (CHBC) derived
from branches of Ficus microcarpa. Results showed the successful loading of chitosan onto biochar surface,
exhibiting more functional groups (e.g., C–
–O, –NH2, and –OH). Langmuir model well described the Sb(III)
sorption isotherm experimental data, and the maximum sorption capacity of Sb(III) by CH1BC (biochar loaded
with chitosan at a ratio of 1:1) was 168 mg g− 1, whereas for the BC it was only 10 mg g− 1. X-ray photoelectron
spectroscopy demonstrated that CH1BC oxidized 86% of Sb(III) to Sb(V), while BC oxidized 71% of Sb(III).
Density functional theory calculations suggested that the synergistic effect of exogenous hydroxyl and inherent
carbonyl contributed to the enhanced removal efficiency of Sb(III) by CHBC. Key mechanisms for Sb(III) sorption
onto CHBCs included electrostatic interaction, chelation, surface complexation, π-π interaction, and hydrogen
bonding. Overall, this study implies that CHBC can be a new, viable sorbent for the removal of Sb(III) from
aquatic systems aiding their safe and sustainable management.
1. Introduction
Antimony (Sb) has been extensively employed in various industrial
applications, including bullets, pigments, batteries, semi-conductors,
and flame-retardant materials (Wei et al., 2020). Various anthropo­
genic activities such as smelting, mining, fuel combustion, and the
widespread use of Sb-containing compounds (e.g., rubbers, alloys) have
led to the transfer of Sb into water bodies, thus posing a high health risk
* Correspondence to: School of Environmental and Chemical Engineering, Foshan University, Foshan, Guangdong 528000, China.
E-mail address: hailong.wang@fosu.edu.cn (H. Wang).
https://doi.org/10.1016/j.jhazmat.2021.127971
Received 4 August 2021; Received in revised form 5 November 2021; Accepted 29 November 2021
Available online 1 December 2021
0304-3894/© 2021 Elsevier B.V. All rights reserved.
H. Chen et al.
Journal of Hazardous Materials 425 (2022) 127971
to humans (Jia et al., 2020). The global production of Sb reached 160,
000 t/year in 2019 (USGS, 2020), with more than 60% of the global Sb
produced in China (Nishad and Bhaskarapillai, 2021). Antimony occurs
naturally in aquatic systems, and the primary oxidation states of Sb are
inorganic antimonite (Sb(III)) and antimonate (Sb(V)); the former is
10-time more toxic than the latter (Xiong et al., 2020). Chronic exposure
at high concentration poses multiple detrimental risks to human beings,
including endocrinological, cardiovascular and neurological diseases
(Wei et al., 2020). Considering water is one of the major mediators of Sb
into the human body, the maximum allowable concentration of Sb in
drinking water was set as 20 μg L− 1 by the World Health Organization
(WHO) (Nishad and Bhaskarapillai, 2021).
Given these low allowable limits of Sb in water, various remediation
strategies have been developed for the removal of Sb from aquatic
ecosystems, including biological process (Wu et al., 2012), membrane
filtration (Zeng et al., 2021), and sorption (Jia et al., 2020; Wei et al.,
2020; Xiong et al., 2020). Sorption has gained significant attention as
the most sustainable method among different strategies owing to its
relative cost-effective, high efficiency, and easy-to-operate trait. Biochar
is an effective sorbent for green remediation of environments contami­
nated with various pollutants (Chen et al., 2020) such as the immobi­
lization/removal of toxic metals (Yin et al., 2020; Albert et al., 2021;
Yang et al., 2021a, 2021b), metalloids (Pan et al., 2021; Wen et al.,
2021), and organic compounds (Qin et al., 2020; Hoang et al., 2021; Guo
et al., 2022). Furthermore, biochar could be modified via various
advanced means, obtaining biochar with high surface functionality
exploiting greater sorption capacity for toxic elements in contaminated
water (Zhou et al., 2020a, 2020b; Bolan et al., 2021).
Chitosan, one of the most commonly utilized biopolymers, has
demonstrated high efficiency for metal(loid)s sorption owing to the
abundance of functional groups, chemical stability, nontoxicity, and
chelation behavior (Huang et al., 2020; Palansooriya et al., 2021).
Chitosan is an effective surface modification agent to functionalize host
material such as biochar with a strong bonding ability for metal(loid)
like Sb (Huang et al., 2020). Introducing chitosan onto biochar can
merge the merits of both bio-materials as functional groups on chitosan
surface (e.g., –NH2, and –OH) offers active binding sites for stable
chemical chelation/complexation with metal(loid) ions, offering a route
to enhance the sorption efficiency (Cui et al., 2017). Meanwhile,
positively-charged chitosan molecules can capture negatively-charged
metal(loids) like oxyanions of Sb(III) through electrostatic attraction
(Zhang et al., 2020).
Several studies have proved the potential of chitosan-modified bio­
chars in the removal of organic/inorganic pollutants, including tetra­
cycline (Liu et al., 2019), thionine dye (Jawad et al., 2021), phosphate
(Huang et al., 2020), and heavy metals (Zhang et al., 2020; Song et al.,
2021; Fan et al., 2022) from aqueous solutions. For instance, Fan et al.
(2022) reported that the cotton biochar-chitosan biomass-based
hydrogel showed superior sorption capacity for Pb2+ (1106 mg g− 1),
Cu2+ (678 mg g− 1), and methylene blue (591 mg g− 1), mainly due to a
strong chelating interaction between the adsorbent and pollutants.
Although several studies have reported the chitosan modification on
biochar for metal removal, most of them focused on the cationic metals
like Cd2+, Pb2+ and Cu2+. However, the sorption efficiency and mech­
anisms may differ in case of oxyanions such as Sb(III) by
chitosan-modified biochar, which remains least understood in the
literature. To our knowledge, this study is the first attempt to investigate
the potential of chitosan-loaded biochar to remove Sb(III) from
contaminated aqueous systems.
Ficus microcarpa is a typical urban green plant with an extensive
geographical distribution spanning from south China to other parts of
the world, including Southeast Asia, America, Europe, and Oceania (Li
et al., 2019b). It constitutes a significant component of urban green
waste due to the periodical trimming of plant branches, a frequent
management measure to maintain the urban landscape (Cui et al.,
2017). Utilizing green waste to produce biochar as a bio-sorbent has
been considered a sustainable solution for managing the ever-increasing
waste biomass in the modern urban community (Li et al., 2019b).
Therefore, in this study, we fabricated a new chitosan-loaded biochar
using F. microcarpa branch waste and tested its potential for Sb(III)
removal from aqueous solutions under a batch sorption experimental
setup. We hypothesize that loading the biochar surface with chitosan
may intensify the existence of oxygen- and nitrogen-containing func­
tional groups on the biochar surface, hence improving the sorption for
Sb(III). Therefore, the specific objectives were to: (1) identify the chi­
tosan modification-induced alteration on biochar characteristics which
may determine its ability for Sb(III) sorption; (2) compare the Sb(III)
sorption capacity of pristine biochar (BC) and chitosan-loaded biochars
(CHBCs) under varying conditions; (3) reveal the potential sorption
mechanisms related to pH, biochar dosage, initial concentration, reac­
tion time, and temperature.
2. Materials and methods
2.1. Chemicals and reagents
Potassium antimonyl tartrate trihydrate (C8H4K2O12Sb2⋅3 H2O),
glutaraldehyde (25%, w/v), acetic acid (CH3COOH), sodium chloride
(NaCl), sodium hydroxide (NaOH), and hydrochloric acid (HCl) were of
analytical grade and obtained from Macklin Bio-Chem Technology Co.,
Ltd. (Shanghai, China). Chitosan (Degree of deacetylation ≥ 75%)
derived from shrimp shells was purchased from Sigma-Aldrich Reagent
Co., Ltd. (Shanghai, China). Deionized water (18.2 MΩ cm− 1, ULPHW-I,
Ulupure Co. LTD., China) was used for the experiment.
2.2. Preparation, modification and characterization of biochars
2.2.1. Pristine biochar preparation
The branches of F. microcarpa were sampled from a park in Foshan
City, China, crushed to ~5 mm and oven-dried at 80◦ C for 24 h and used
as biomass for biochar production. The pristine biochar (BC) was pro­
duced from the F. microcarpa branches via pyrolyzing the dried biomass
at 500 ± 5◦ C and held for 2 h with a heating rate of 15 ± 2◦ C min− 1 in a
tubular furnace (SLG1100–100, Shanghai Litres Instrument Co., LTD,
China). The BC was ground and sieved (< 2 mm) for chitosan modifi­
cation and further batch sorption experiments.
2.2.2. Chitosan-loaded biochar preparation
The chitosan loading procedure was adopted from previous studies
with slight modification (Zhang et al., 2020; Palansooriya et al., 2021).
Briefly, 0.4, 1.0, 2.0 g of chitosan were dissolved in parallel in 200 mL
2.5% acetic acid solutions, then each solution was supplemented by 2.0
g of BC, and homogenized together using a magnetic stirring for 1 h at
25◦ C. Afterward, 3 mL of glutaraldehyde was added to the above solu­
tions. In the following step, 300 mL of NaOH solution (1%) was slowly
decanted into the mixture, and consecutively ultrasonic shaken for 1 h.
Subsequently, the above solution was sealed and kept still at 25◦ C for 18
h. Finally, CHBCs were collected, washed three times using deionized
water, vacuum dried (70◦ C) until completely dry, and ground
(100-mesh) for further analysis. Based on the ratio of chitosan to bio­
char, the obtained chitosan-loaded biochar (CHBC) treatments were
labeled as CH0.2BC, CH0.5BC, and CH1BC.
Glutaraldehyde is one of the frequently used cross-linking agents,
and the quantity adopted is a pivotal role in the cross-linking with chi­
tosan (Jawad et al., 2021). The amounts of glutaraldehyde would affect
the subsequent modification process of chitosan loading onto biochar
and affect the sorption efficiency of pollutants. In this respect, 2.0 g of
chitosan and 2.0 g of BC were added in 200 mL of 2.5% acetic acid so­
lutions, and stirring at 25◦ C for 1 h. Thereafter, various volumes of
glutaraldehyde (i.e., 0, 0.1, 0.2, 0.5, 1, 2, 3, 4, and 5 mL, named as
treatments G0-G5) were supplemented dropwise to the above solutions.
The following modification procedures were performed as described
2
H. Chen et al.
Journal of Hazardous Materials 425 (2022) 127971
above. The modified biochars differing in glutaraldehyde dosage were
used in further sorption experiments.
Characterization of pristine and modified biochars using various
spectroscopic techniques is presented in the supporting information.
equilibrium data. Thermodynamic parameters were further employed to
elucidate the thermodynamics of Sb(III) sorption onto the biochars.
Detailed information on these models and parameters is presented in the
supplementary material.
2.5. DFT calculations
2.3. Batch sorption experiments
Spin-polarized density functional theory calculations were carried
out under the scheme of generalized gradient approximation (Kohn and
Sham, 1965), with the use of PBE functional (Hammer et al., 1999) and
double numerical polarized (DNP) basis, as embedded in DMol3 package
(Delley, 1990, 2000). A global orbital cutoff with a radius of 4.5 Å was
employed, under which the adsorption geometry was fully relaxed with
total energy and atomic force converged to 10− 6 Ha and 0.005 Ha/Å
(1 Ha ≈ 27.2114 eV). DNP basis has been extensively tested, whose size
is comparable to Gaussian basis 6–311 +G* * sets (Inada and Orita,
2008). Solution effect has been considered using explicit model
(COSMO, with a dielectric constant of 78.3, as the default in DMol3). All
calculations were based on cluster model (described below); therefore,
no k-space sampling and vacuum layers are requested.
Sb(III) sorption over biochar was simulated based on three steps: (i) a
graphene cluster with C6v symmetry, containing 54 carbon atoms, has
been generated from graphene monolayer as a functional group-free
substrate; (ii) 7 functional O- and N-groups (Basal Plane, Ring–COOH,
– O, –COOH–HO–, where ‘–’ indicates
–COOH, –NH2, –OH–NH2, –OH–C–
the group coupled by hydrogen bond) have been introduced at the edge
of carbon framework, representing the rich chemistry associated with
edging groups; (iii) Sb(III) has been introduced to these sites one by one,
followed by full relaxation and energy calculation, based on which the
adsorption energy ΔE has been calculated as an indicator of adsorption
strength. Specifically, ΔE = E(Sb(III)* )-E(*)-E(Sb(III)), in which E(*), E
(Sb(III)) and E(Sb(III)* ) are the calculated total energies of clean bio­
char model (see step (ii)), free Sb(III) and Sb(III) sorbed over biochar. To
deliver a reliable comparison, Sb(III) sorption over basal plane (far away
from edging groups) has been employed as a reference when the role of
O-/N-groups has been discussed. Under this scheme, large negative ΔE
means strong sorption cacpacity.
A stock solution containing 1000 mg L− 1 of Sb(III) was prepared by
dissolving potassium antimonyl tartrate trihydrate (C8H4K2O12Sb2⋅3
H2O) in deionized water. The working solutions were prepared by
diluting specific volumes of the stock solution with 0.01 M NaCl as the
background electrolyte to maintain ionic strength.
Batch sorption experiments were carried out following the reported
conditions of previous methods (Jia et al., 2020; Nie et al., 2021).
Briefly, 0.05 g of selected biochars (i.e., BC, CH0.2BC, CH0.5BC, and
CH1BC) were added into 25 mL working solution under various influ­
encing conditions. The mixture was oscillated for 24 h, with a rotational
speed of 180 rpm prior to analysis. The influence of biochar and
glutaraldehyde dosages were respectively studied with biochar dosages
of 0.2–2.5 g L− 1 and glutaraldehyde volumes from 0 to 5 mL, with an
initial Sb(III) concentration of 40 mg L− 1, at 25◦ C. In particular, the
initial Sb(III) concentration was referred to the previous studies (Jia
et al., 2020; Wan et al., 2020; Wei et al., 2020), which were selected as
30–50 mg L− 1. The Sb(III) sorptions affected by initial solution pH and
ionic strength were evaluated with a fixed initial Sb(III) concentration
(40 mg L− 1), with initial pH ranging from 2 to 12, and NaCl concen­
trations ranging from 0.01 to 0.25 M, at 25◦ C. The initial solution pH
was adjusted using 1 M HCl and/or 1 M NaOH. The ubiquitously found
ions (NO3-, Cl-, SO42-, and PO43-) in the aqueous system were chosen as
the model co-existing ions to investigate their influence on Sb(III)
removal. The initial concentration of these anions was selected as 40 mg
L− 1, acting as a typical concentration of anions in wastewater (Wang
et al., 2018). In addition, humic substance is a major component of
natural organic matter (NOM) in natural water (Wan et al., 2020).
Therefore, humic acid was introduced into the solution to test its in­
fluence on Sb(III) removal at two dose levels, i.e., 5 mg L− 1 and 20 mg
L− 1 (Wan et al., 2020). Kinetic sorption experiments were conducted by
adding 0.05 g of used biochars to 25 mL 40 mg L− 1 Sb(III) solution, and
then oscillated (180 rpm) for 24 h, at 25◦ C. Samples were collected at
various time intervals to determine the Sb(III) concentrations. Sorption
isotherm experiments were carried out by adding used biochars (0.05 g)
into 25 mL Sb(III) solution with various initial concentrations, and then
oscillated (180 rpm) for 24 h, at 25◦ C. Moreover, the thermodynamics of
Sb(III) removal by biochars was also investigated with specific tem­
peratures at 25◦ C, 35◦ C, and 45◦ C.
After sorption, all samples were filtered (0.45-μm) prior to the
quantitative analysis. Residual Sb(III) concentration in the supernatant
was measured within 24 h using an atomic absorption spectrometer
(ZA3300, Shimadzu, Japan). The biochar-sorbed amount of Sb(III) and
the removal efficiency were calculated as follows (Rahman et al., 2021):
Qe = (Co − Ce )V/m
(1)
η = (Co − Ce )/Co × 100%
(2)
2.6. Desorption and sorbent regeneration
After the batch sorption was completed, the Sb(III)-loaded biochars
were filtered, washed by ultrapure water, and oven-dried. Then, the
biochars were added into 25 mL NaCl solution (0.01 M), shaken
(180 rpm) for 24 h at 25◦ C, and this process was repeated three times.
Samples were collected and filtered (0.45 µm) every 24 h, and Sb con­
centration was quantified at each cycle using the atomic absorption
spectrometer.
The reusability of chitosan-loaed biochar was evaluated using NaOH
solution (0.5 M) as the desorption agent (Deng et al., 2020). The re­
generated CH1BC was rinsed with deionized water and then added into
25 mL 40 mg L− 1 Sb(III) solution, and four consecutive Sb(III) sorp­
tion/desorption cycles were conducted to investigate the sorption ca­
pacities of the regenerated biochars at each cycle.
2.7. Data quality control and statistical analysis
where Qe (mg g− 1) is the sorption capacity at equilibrium time; Co (mg
L− 1) represents the initial Sb(III) concentration; Ce (mg L− 1) represents
the final Sb(III) concentration after equilibrium; V (L) is the solution
volume; m (g) is the biochar mass and; η (%) is the removal percentage.
All sorption experiments were conducted in triplicate, and the rela­
tive standard deviation of triplicate analysis was set to < 5%. The
plasticware and glassware used in the experiment and analysis were
soaked in 3% nitric acid for 24 h and rinsed with deionized water. The
atomic absorption spectrometer was recalibrated after the measurement
of each 25 samples.
The statistical analyses were performed using SPSS 26.0. The
experimental data were expressed as mean ± standard error (n = 3).
The significant differences (P < 0.05) were evaluated using the analysis
of variance (ANOVA) and Duncan’s multiple range t-tests. Origin 2021
2.4. Modeling for Sb(III) sorption
The kinetics sorption data were fitted using four models, i.e., pseudofirst-order, pseudo-second-order, Elovich and intra-particle diffusion
models. In addition, three sorption isotherm models, namely the Lang­
muir, Freundlich, Temkin models, were adopted to fit experimental
3
H. Chen et al.
Journal of Hazardous Materials 425 (2022) 127971
and R studio software were used in the data graphing. XPS data were
analyzed and deconvoluted using the Thermo Avantage program.
Fig. S3. In the spectrum of chitosan, the occurrence of the peak at
3417 cm− 1 could be ascribed to the stretching vibrations overlapping of
N–H and O–H bonds (Ren et al., 2013), and the peaks at 2930 and
2875 cm− 1 were assigned to the stretching vibrations of C–H bond, i.e.,
–CH– and –CH2–, respectively (Chen et al., 2019). The axial stretching
bands around 1651 cm− 1, 1377 cm− 1, 1076 cm− 1 were due to the C–O
stretching band (amide I), –CH3 symmetrical angular deformation, and
C–O stretching from β(1→4) glycosidic bonds, respectively (Monier
et al., 2010; Ren et al., 2013).
–O
For all biochars, various bands including carbonyl C–
– O (1400 cm− 1),
(1585 cm− 1), symmetrical stretching carboxyl O–C–
and aromatic C–H bond (875 cm− 1) in the spectra were identified
(Zhang et al., 2020; Mahmoud et al., 2021), indicating the original
functional group types of biochar were maintained after the chitosan
modification, despite with a slight shift of stretching peaks (Fig. S3).
Emerging absorption peaks around 1654 cm− 1, 1070 cm− 1, 3421 cm− 1
were detected in the CHBCs, which corresponded to the stretching bands
from chitosan, indicating that the abundant functional groups of chito­
san were introduced to biochars (Ren et al., 2013). Overall, biochars,
especially for the CH1BC, are covered with large active functional
groups, rendering them potentially superior sorbents for Sb ions.
The XPS spectra showed that the principal elements on the BC sur­
face were C 1 s (85.3%), O 1 s (12.2%), N 1 s (1.5%), while increased O
1 s (22.8%), N 1 s (3.9%) and decreased C 1 s (73.3%) were detected in
the CH1BC (Fig. S4). This was consistent with the aforementioned
conclusion of atomic ratio change, further confirming that the chitosan
had adhered onto the biochar surface. The C 1 s of BC (Fig. 1A) could be
deconvoluted into three peaks, i.e., 284.80, 285.71, and 289.36 eV,
which were respectively assigned to the C–C (68.2%), C–O (27.2%), and
– O (4.6%) (Zeng et al., 2019; Palansooriya et al., 2021). After
HO–C–
chitosan modification, the C 1 s of CH1BC showed four peaks at 284.04,
284.80, 286.52, and 287.91 eV (Fig. 1B), corresponding to the C–H
– O (13.9%), respectively
(15.2%), C–C (45.9%), C–O (25.1%) and N–C–
(Palansooriya et al., 2021; Zhang et al., 2021).
As for O 1 s, two peaks centered at 531.98 (C–O, 69.6%) and 533.53
(–OH, 30.4%) eV were found in BC (Fig. 1C, D); whereas two peaks at
531.65 (C–O, 13.6%) and 533.26 eV (–OH, 86.4%) were noted in CH1BC
with slight location shift (Xiong et al., 2020; Zhang et al., 2021). In N 1 s
spectra, as for BC, four peaks at 397.49, 398.34, 400.00, and 403.74 eV
were respectively attributed to metal nitrides, –NH, –NH2, and
CH3CO–NH (Fig. 1E, F); while CH1BC showed only two peaks corre­
sponding to 399.94 eV (–NH2) and 402.92 eV (CH3CO–NH) (Pal­
ansooriya et al., 2021; Song et al., 2021). The peaks intensity and
location shifts in C 1 s, O 1 s and N 1 s after modification indicate that
interaction occurred between biochar and chitosan; the CHBC was
functionalized with more oxygen/nitrogen-containing groups like hy­
droxyl, carboxyl and amino groups which is consistent with the FTIR
results (Fig. S3).
3. Results and discussion
3.1. Characterization of the pristine and chitosan-loaded biochars
3.1.1. Physicochemical characteristics
The biochar modification with chitosan resulted in higher H, N and O
contents, and lower C content than that in the BC (Table 1). The
increased contents of H, O, and N in CHBCs indicate the successful
loading of chitosan onto biochar surfaces. Meanwhile, the decrease in C
content is attributed to surface exposure to the abundant oxygen/
nitrogen-containing groups, reducing the C proportion in the total
mass of modified biochars (Zhou et al., 2013). The increased ratio of
chitosan to biochar led to a decrease of pH following the order: CH1BC <
CH0.5BC < CH0.2BC < BC (Table 1), which was attributed to (1) the
decrease of alkaline minerals contents on the biochar’s surface, (2)
formation of acidic functional groups on surfaces of CHBCs, and (3) the
neutralization effect with chitosan (pH 6.80). The first assumption can
be supported by the EDS results (Fig. S2A), demonstrating the decrease
of ash and alkaline element (Ca, Mg) after chitosan loading. In partic­
ular, the ash content decreased as the chitosan loading ratio increased,
mainly due to the thermal instability of chitosan. Furthermore, the
chitosan modification enhanced the CEC of biochars, which could be
ascribed to chitosan’s strong cation exchange nature (Table 1). On the
other hand, the chitosan loading simultaneously decreased the biochar
specific surface area and pore volume (Table 1), owing to blocking some
biochar pore structures (Zhou et al., 2013).
3.1.2. Morphological and qualitative surface characteristics
The surface morphology of CH1BC did not change significantly
compared to the BC (Fig. S1, A-F). The BC showed an angular and
honeycomb structure with multiple pore channels (Fig. S1, B, E), while a
sugar-coating-like layer was noted on the CH1BC, covering some biochar
pores and passivating the edges of the biochar (Fig. S1C, F). Accordingly,
the surface area and pore volume of CHBCs were decreased (Table 1).
The TEM images indicated the layer-mesostructure of BC, while the
CH1BC showed a coccoid-shaped cluster morphology due to the crosslinking between biochar and chitosan (Fig. S1G-J).
The EDS spectra demonstrated that the CH1BC had higher O and N
contents and lower C and alkaline mineral element content (Mg, K, Ca)
than the BC (Fig. S2A). This result is consistent with the result of
elemental analysis, hinting that the chitosan successfully loaded onto
the biochar, and further confirming the modification-induced pH
decrease is related to the loss of alkaline mineral elements. The promi­
nent characteristic peaks of chitosan at ~ 11.0◦ and 20.0◦ displayed in
the XRD pattern (Fig. S2B) were assigned to the chitosan crystal forms I
and II, respectively (Zhang et al., 2019b). In addition, calcium oxide and
calcite were observed in the BC, and an enhanced calcium oxide peak
and an emerging chitosan-specific peak at 20◦ were detected in the
CH1BC (Fig. S2B). This indicates that the CH1BC has possessed the
crystals of both biochar and chitosan.
The FTIR spectra of chitosan, BC, and CHBCs are illustrated in
3.1.3. Swelling performance and weight loss
The CH1BC had a greater swelling ratio (SR) than BC, increased time
and temperature positively affected on swelling performance of both
biochars (Fig. S5, A, B). For instance, the maximum SR of CH1BC (130%)
and BC (93%) was obtained at 240 min (Fig. S5A), which was due to that
Table 1
Physico-chemical characteristics of chitosan and biochars.
Samplea
C
H
O
N
S
pH
(H2O)
Ash content
(%)
Cation exchange capacity (cmol
kg− 1)
Specific surface area (m2
g− 1)
Pore volume (cm3
g− 1)
7.3
0.6
0.9
1.9
3.6
0.18
0.08
0.04
0.05
0.76
6.8
9.3
5.9
5.5
4.8
0.4
8.4
7.6
3.4
1.4
4.61
1.01
1.06
1.34
1.87
1.96
4.96
4.14
3.18
2.23
0.002
0.013
0.010
0.006
0.004
(%)
Chitosan
BC
CH0.2BC
CH0.5BC
CH1BC
a
41.6
74.3
63.9
56.2
53.8
7.4
3.0
4.0
4.7
5.4
42.5
20.2
29.2
35.2
35.2
BC: pristine biochar; CH0.2BC, CH0.5BC, CH1BC: biochar loaded with chitosan at a ratio of 0.2:1, 0.5:1 and 1:1, respectively.
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Journal of Hazardous Materials 425 (2022) 127971
Fig. 1. XPS spectra of C 1 s (A and B), O 1 s (C and D), and N 1 s (E and F) for BC and CH1BC. BC: pristine biochar; CH1BC: biochar loaded with chitosan at a ratio
of 1:1.
water molecules gradually occupied the internal pores of the sorbent
until equilibration. In Fig. S5B, enhancement of hydration over tem­
perature could be ascribed to the disturbance of H-bonding between the
sorbent networks (Mahmoud et al., 2021). The higher SR of CH1BC
might be attributed to the electrostatic repulsion between adjacent
ionized amino groups carried by chitosan (Qu et al., 2000).
As for the TGA results, the weight percent of BC, CH0.2BC, CH0.5BC
and CH1BC decreased over temperature, by 21%, 32%, 48% and 55%,
respectively (Fig. S5C). Accordingly, the degradation of all the biochars
is based on two stages. The first stage for all biochars showed sharp
weight loss from 25◦ C to 105◦ C, which is attributed to the escape of
adsorbed water (Ren et al., 2013). As for the second stage, the weight
loss of BC gradually increased from 105◦ C to 800◦ C, due to the degra­
dation of functional groups and graphitic char (Mahmoud et al., 2021);
the weight loss for CHBCs from 210◦ C was due to the decomposition of
chitosan polymer over temperature (Fadaiea et al., 2019).
3.2. Chitosan-loaded biochar for Sb(III) sorption
3.2.1. Effect of chitosan-biochar ratio and biochar dosage
The impact of biochar dosage on Sb(III) sorption by BC and CHBCs is
shown in Fig. 2A. The higher dosages of different biochars led to higher
removal efficiency of Sb(III). At the same dosage, CH1BC showed the
highest Sb(III) removal efficiency, with maximum removal of 88% at
2.5 g L− 1 dose (Fig. 2A), indicating the critical role of chitosan-biochar
ratio in Sb(III) sorption. The sorption efficiencies increased with an in­
crease in sorbent dosage can be ascribed to the enhancement of binding
sites (Yadaei et al., 2018). The biochar dosage of 2 g L− 1 was adopted for
all the further sorption experiments in this study, considering the cost,
effectiveness and easy-to-operate conditions compared to other biochar
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Journal of Hazardous Materials 425 (2022) 127971
Fig. 2. Impact of chitosan-biochar ratio and biochar dosage on Sb(III) removal efficiency (main plot) and residual Sb(III) concentration in solution (the inset plot)
(A); glutaraldehyde dosage influence on Sb(III) sorption capacity of CH1BC (B). Treatments: G0-G5: added glutaraldehyde volume with 0–5 mL. BC: pristine biochar;
CH0.2BC, CH0.5BC, CH1BC: biochar loaded with chitosan at a ratio of 0.2:1, 0.5:1 and 1:1, respectively.
Fig. 3. Impact of initial solution pH (A), ionic strength (B), and co-existing substances (C) on Sb(III) sorption on the BC and CHBCs. BC: pristine biochar; CH0.2BC,
CH0.5BC, CH1BC: biochar loaded with chitosan at a ratio of 0.2:1, 0.5:1 and 1:1, respectively; HA-5: humic acid dose = 5 mg L− 1; HA-20: humic acid
dose = 20 mg L− 1.
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Journal of Hazardous Materials 425 (2022) 127971
dosages.
0.25 M (Fig. 3B). The decrease of sorption capacity at lower ionic
strength (0.01–0.05 M) could be attributed to the Sb(III)-Cl- competition
for the sorption sites on the biochar surface (Xi et al., 2010; Zhang et al.,
2020). Moreover, the increased Na+ ions led to an elevated degree of
charge screening (Huynh and Chen, 2011), hence inhibiting the Sb(III)
sorption. When the ionic strength increased beyond 0.10 M, the Sb(III)
might only form an inner-sphere complex with biochar and thus the
sorption process was not dependent on the ionic strength (Xi et al.,
2010). The responsive Sb(III) sorption depending on the ionic strength
in this study suggested the formation of both inner-sphere and
outer-sphere Sb(III) complexes on biochars, as indicated by Rahman
et al. (2021).
As illustrated in Fig. 3C, the presence of NO3-, Cl-, and SO42- showed
an insignificant influence on Sb(III) sorption by biochars, which was in
agreement with the conclusion by Xiong et al. (2020) and Wan et al.
(2020). However, the co-existing PO43- significantly (P < 0.05)
decreased the Sb(III) sorption by 19% on the CH1BC (Fig. 3C). In this
respect, PO43- has a similar molecular structure with antimonite (Wang
et al., 2018), and they may share the same active reaction sites. Huang
et al. (2020) reported that the chitosan-modified biochar showed a
strong affinity to phosphate, with a maximum sorption capacity up to
109 mg g− 1. Therefore, the existence of PO43- could significantly impede
the Sb(III) sorption performance of CH1BC via competitive adsorption
(Xiong et al., 2020). Moreover, the introduced humic acid (HA) at
20 mg L− 1 enhanced the sorption of Sb(III) in the BC and CH1BC by
41.8% and 12.5%, respectively, as compared to the control (Fig. 3C). We
assume that the humic substance might be sorbed in the biochar surface,
and thereby providing additional sorption sites for Sb(III)-HA complexes
as suggested by Deng et al. (2020). Wan et al. (2020) also indicated that
the formation of Sb(III)-HA complexes between biochar and HA
increased the Sb(III) sorption on the manganese modified biochar than
that of manganese oxides.
3.2.2. Effect of glutaraldehyde dosage
Glutaraldehyde volume less than 0.5 mL in CH1BC did not signifi­
cantly influence the Sb(III) sorption (Fig. 2B). This was mainly due to the
poor stability of glutaraldehyde in acidic solutions and partial dissolu­
tion, thereby hindering the further cross-linking of low-dosage glutar­
aldehyde and chitosan (Huang et al., 2020). At increased volumes of
glutaraldehyde (> 0.5 mL), the Sb(III) sorption capacity of CH1BC
improved; the maximum sorption capacity was observed in the G4
treatment, increased by 6.33 folds, as compared to the G0 treatment
(Fig. 2B).
A Schiff-base reaction could occur between the carboxyl group
– O) in glutaraldehyde and hydroxy (–OH) and amine (N–H) bonds in
(C–
chitosan during cross-linking, and this process facilitated the stability
and mechanical strength of chitosan in acidic media (Jawad et al.,
2021). We regard that the increase of glutaraldehyde concentration
contributed to the enhancement of the cross-linking effect, thus
improving the binding between biochar and chitosan, and ultimately
caused a positive impact on Sb(III) sorption. In this respect, Huang et al.
(2020) found that strengthening the cross-linking via increasing the
mass ratio of glutaraldehyde to chitosan favored the phosphate sorption
to chitosan-wheat straw biochar composite. However, the
non-significant (P > 0.05) difference of Sb(III) sorption capacity be­
tween G3, G4 and G5 suggested that 3 mL of glutaraldehyde was suffi­
cient to enhance the cross-linking with chitosan. Therefore, from the
perspective of an environmentally friendly option, the optimal volume
of glutaraldehyde was selected as 3 mL in the subsequent batch
experiments.
3.2.3. Effect of initial solution pH
All treatments showed the maximum Sb(III) sorption when the initial
solution pH was 2 (Fig. 3A). The highest Qe value was up to 17.2 mg g− 1
for CH1BC, whereas Sb(III) sorption sharply declined as the initial so­
lution pH decreased from 3 to 6. At an initial pH range of 6–11, the Sb
(III) sorption on biochars seems to be relatively stable. It is worth
mentioning that Sb(III) sorption capacities by CHBCs increased
(P < 0.05) by 8.9%− 16.8% at pH 12 as compared to those at initial pH
11, whereas the sorbed Sb(III) onto the surface of BC decreased by
51.1% (Fig. 3A).
For Sb(III) sorption, the initial solution pH impact could be ascribed
to distinct predominant species of Sb(III) within a range of pH (Cui et al.,
2017). The dominating speciation of Sb(III) is Sb(OH)2+, Sb(OH)30 and
Sb(OH)4- in strongly acidic, weak acidic-neutral, and alkaline pH ranges,
respectively (Wan et al., 2020). At initial solution pH of 2, Sb(III) exists
as Sb(OH)2+ species, hence the high Sb(III) sorption capacity was
attributable to the electrostatic interaction between the
negatively-charged biochar surface and Sb(OH)2+. The decrease of Sb
(III) sorption efficiency at initial pH 3–6 might be ascribed to the
weakened electrostatic interaction caused by the decrease in Sb(OH)2+
(Cui et al., 2017). At initial pH above 6, Sb(OH)30 was the predominant
form and was stable over an initial 6–11 pH range (Xiong et al., 2020),
which was hard to be sorbed by biochar. When the initial solution pH
was > 11, the major form of Sb(III) was Sb(OH)4-, the deprotonated
groups (e.g., –NH2, and –OH) on the CHBCs favored capture of Sb(OH)4through complexation/chelation and hydrogen bonding (Yadaei et al.,
2018). For the BC at initial pH 12, the electrostatic repulsion between
negatively-charged surface of biochar and Sb(OH)4- was responsible for
the significant decrease of Sb(III) sorption, and the competition for
sorption sites between hydroxyl and Sb(OH)4- might be another possible
reason (Iqal et al., 2013).
3.3. Sorption kinetics
Results of pseudo-first-order, pseudo-second-order, Elovich and
intra-particle diffusion models and their fitting parameters are shown in
Fig. 4A and Table S1. The Sb(III) sorption capacity reached 73% (BC),
65% (CH0.2BC), 77% (CH0.5BC), and 62% (CH1BC) of the maximum
sorption capacity within 5 min, and these sorption processes completed
93%− 96% within 60 min (Fig. 4A). The results suggest that the BC and
CHBCs may be potentially excellent Sb(III) sorbent with a prominent
sorption rate, as fast sorption processes are cost-effective for sorption
facilities (Wan et al., 2020).
In this study, Sb(III) sorption process conformed to a typical twostage sorption (Iqal et al., 2013; Liu et al., 2019), where Sb ions
rapidly occupied the available active sites in the fast-stage (0–60 min),
and gradually being sorbed in the slow-stage (60–1440 min) until the
consequent equilibrium approached. Kinetic data were well fitted by the
pseudo-second-order model, with R2 spanned 0.92–0.99 (Table S1).
Meanwhile, the values of derived Qe were more consistent with the
experimental data, indicating that the rate-determining step of Sb(III)
sorption was governed by the chemisorption process (Rahman et al.,
2021). Given the lower model-fitted R2, Sb(III) sorption kinetics data
were poorly described by the pseudo-first-order kinetic model (R2 =
0.56–0.88) and Elovich model (R2 = 0.47–0.76) (Table S1). We assume
that the possible reason could be the desorption of Sb(III) in the biochars
due to the Sb(III) oxidation phenomenon, as indicated by Cui et al.
(2017). We also employed the intra-particle diffusion model to expound
the diffusion mechanism in the sorption process, and the fitting results
are shown in Fig. 4B. As the fitted curve did not go through the original
point (C‡0) (Table S1), thus, the intra-particle diffusion was not the sole
dominating step in our study, more than one step governed the sorption
process (Rahman et al., 2021). All the sorption processes can be divided
into 3 linear stages (Fig. 4B). At the first stage, the fitted K1 values were
higher than the other stages (Table S1), suggesting the highest sorption
3.2.4. Effect of ionic strength and co-existing substances of aquatic system
The sorption of Sb(III) onto various biochars was suppressed as the
ionic strength ranged from 0.01 to 0.10 M, whereas Sb(III) sorption was
relatively unaffected by ionic strength changes at the range from 0.10 to
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Journal of Hazardous Materials 425 (2022) 127971
Fig. 4. Sorption kinetics of Sb(III) on the BC and CHBCs using pseudo-first-order, pseudo-second-order, Elovich (A) and intra-particle diffusion models (B); sorption
isotherms of Sb(III) on the BC and CHBCs (C). BC: pristine biochar; CH0.2BC, CH0.5BC, CH1BC: biochar loaded with chitosan at a ratio of 0.2:1, 0.5:1 and 1:1,
respectively.
rate; this could be due to the boundary layer diffusion of Sb(III) by
external surface sorption (Wan et al., 2020). At the second stage,
intra-particle diffusion has become the rate-limiting step leading Sb
ions’ gradual occupation onto sorption sites, whereas the third stage was
due to the equilibrium (Liu et al., 2019; Xiong et al., 2020).
The maximum sorption capacity (Qm) derived from the Langmuir
model for CH0.2BC, CH0.5BC, and CH1BC was 86, 115, and 168 mg g− 1,
respectively (Table S2). Enhanced values of the affinity coefficient, KL,
also suggested that higher chitosan content in biochar caused higher Sb
(III) affinity (Jia et al., 2020). Separation factor (RL) was a parameter
defined from the Langmuir model for Sb(III) sorption to sorbents (Sup­
plementary material). Typically, the sorption process is unfavorable (RL
> 1), linear (RL = 1), favorable ( 0 < RL < 1), and irreversible (RL = 0)
(Shakoor et al., 2019). The RL of BC and CHBCs ranged between 0 and
0.98 (Table S2), indicating favorable Sb(III) sorption by the different
treatments. Likewise, the range of RL values gradually decreased as the
chitosan loading rate increased, demonstrating the role of
chitosan-loading in promoting the Sb(III) sorption. The Temkin model
also fitted the sorption data well (R2, 0.86–0.98), suggesting that Sb(III)
sorption process may be affected by the interactions between sorbent
and sorbate (Rahman et al., 2021). The comparative Sb(III) sorption
capacities of other potential sorbents and CHBCs in the aqueous solution
are presented in Table S3. Generally, CHBC exhibited high Sb(III)
sorption capacity in addition to its facile and low-cost synthesis process,
making it an environmentally-friendly sorbent for Sb-polluted water
remediation.
3.4. Sorption isotherms
Sorption isotherms of Sb(III) on biochars were fitted using the
Langmuir, Freundlich and Temkin models (Fig. 4C), and the associated
parameters are listed in Table S2. The Freundlich model fitted better for
Sb(III) sorption on BC with R2 = 0.97, and Langmuir model provided
higher R2 values (0.94–0.99) for Sb(III) sorption on CHBCs. Freundlich
model is deemed better for the sorption on heterogeneous surfaces,
while the Langmuir model describes it better for homogeneous surfaces
(Chen et al., 2021). Hence, the modeling results suggested that the
sorption of Sb(III) on the BC was multi-site heterogeneous sorption
(Chen et al., 2021). However, chitosan modification enhanced the ho­
mogeneity of biochar surface, indicating the occurrence of monolayer
sorption of Sb(III) on the CHBCs’ surface (Chen et al., 2021). These re­
sults also agreed with the observed distribution of BC- and CH1BC-sor­
bed Sb in HR-TEM images (Fig. 6).
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Journal of Hazardous Materials 425 (2022) 127971
Fig. 6. High-resolution transmission electron microscope (HR-TEM) images and associated elemental mapping of BC (A), Sb-loaded BC (B), CH1BC (C) and Sb-loaded
CH1BC (D). BC: pristine biochar; CH1BC: biochar loaded with chitosan at a ratio of 1:1.
3.5. Sorption thermodynamics
diffusion (Zeng et al., 2019). As described in Section 3.1.3, the
increasing temperature enhanced the swelling ratio of biochars
(Fig. S5B), which can support the second assumption.
The van’t Hoff plot for the Sb(III) sorption onto CHBCs is presented
in Fig. S6B, and the associated thermodynamic parameters are shown in
Table S4. All the values of ΔG◦ were negative, ranging between − 0.28
and − 4.17 kJ mol− 1, asserting that the Sb(III) sorption process was
spontaneous and thermodynamically favorable, and verifying the
occurrence of physical interactions (Zeng et al., 2019), and the
decreased value of ΔG◦ endows the sorption process more beneficial (Lin
As the temperature increased from 288 K to 308 K, the sorption ca­
pacities of CH0.2BC, CH0.5BC, and CH1BC significantly (P < 0.05)
increased from 10.6 to 11.7, 12.5–13.3, 16.0–17.0 mg g− 1, respectively
(Fig. S6A). We assume two possible reasons: (1) increasing temperature
promoted the activation of sorption sites and enhanced the collisions
with Sb ions in solution (Dong et al., 2017); (2) high
temperature-induced swelling effect on the particle porosity favored Sb
ions sorption on CHBCs via outer boundary layer and inner pore
Fig. 5. Effect of desorption cycles on the Sb(III) sorption capacity (main plot) and Sb(III) removal efficiency (the inset plot) of BC and CHBCs (A), and the removal
efficiencies of Sb(III) by the regenerated CH1BC at each cycle (B). BC: pristine biochar; CH0.2BC, CH0.5BC, CH1BC: biochar loaded with chitosan at a ratio of 0.2:1,
0.5:1 and 1:1, respectively.
9
H. Chen et al.
Journal of Hazardous Materials 425 (2022) 127971
et al., 2018). The positive ΔH◦ values imply that the sorption was
endothermic, as supported by the enhanced Sb(III) sorption capacity
onto CHBCs with a temperature rise (Fig. S6A). Furthermore, the posi­
tive ΔS◦ values reveal the randomness at the liquid-solid interface dur­
ing the sorption processes (Zhang et al., 2020).
particular, the sorption efficiency of CH1BC for Sb(III) remained 92.3%
after 3 adsorption-desorption cycles. It indicated that chitosan-loaded
biochar had great reusability performance, suggesting its promising
potential for practical application.
3.7. Sb(III) removal mechanisms
3.6. Desorption and sorbent regeneration
3.7.1. BET and HR-TEM analyses
The specific surface area, pore volume and pore diameter of the Sbsorbed BC and CHBCs significantly decreased, compared to pre-sorption
(Table S5), which highlights the role of pore filling in the Sb(III) sorption
by biochar. However, given the small specific surface area of biochar in
this study (Table 1), pore filling may not be the most predominant
sorption mechanism. Zhang et al. (2019a) indicated that the adsorbents’
surface area and porous structure may have a weaker influence on heavy
metals adsorption than oxygen-containing functional groups.
The HR-TEM images and associated elemental mapping are pre­
sented in Fig. 6. The major elements of BC before Sb(III) sorption were C,
O, N, S, Ca, and K, and those of CH1BC were C, O, N and S. TEM
elemental images of Sb-loaded BC confirmed the heterogeneous pres­
ence of Sb onto biochar surfaces (Fig. 6B), while the CH1BC displayed
greater intensity and brightness with a homogeneous distribution of Sb
(Fig. 6D), which was in agreement with the conclusion summarized in
The stability of biochar sorbing Sb(III) was investigated by desorp­
tion experiments, to test their potential secondary release into the so­
lution. After three desorption cycles, about 70% (BC) and 81% (CH1BC)
of Sb(III) were still retained (Fig. 5A). As expected, the removal effi­
ciencies of BC and CHBCs declined as the desorption cycles increased,
while the CH1BC showed a 70% removal efficiency for Sb(III) after the
third desorption cycle (Fig. 5A, the inset plot). The Sb(III) desorption
rates of the CH0.2BC (20%), CH0.5BC (21%), CH1BC (19%) were lower
than that of the BC (30%) (Fig. 5A), which indicated that the Sb(III)
sorption by CHBCs mitigated the secondary release and was more stable
than the BC. The anti-desorption abilities of BC and CHBCs may stress
the importance of chemisorption, such as inner-sphere surface
complexation (Rahman et al., 2021).
Furthermore, the regeneration results demonstrated the removal
ability of CH1BC decreased slightly after each cycle (Fig. 5B). In
Fig. 7. The FTIR spectra (A), XPS full-scan spectra (B) along with the spectra of C 1 s (C and D), and N 1 s (E and F) of the BC and CH1BC after Sb(III) sorption. BC:
pristine biochar; CH1BC: biochar loaded with chitosan at a ratio of 1:1.
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Journal of Hazardous Materials 425 (2022) 127971
Section 3.4.
during the sorption process. The formation of N-containing groups-Sb
complexes could explain the lower binding energy value of these
peaks, and the electron-donating of the sorbed Sb thus generated a
higher electron density (Hao et al., 2019). Furthermore, two new peaks
noted at ~ 407 eV for both biochars (Fig. 7E, F) were assigned to nitrate
(Addaria et al., 2010). This may suggest a possible formation of
nitrate-like substance (i.e., Sb–N–O bond) after Sb-O/N-containing
groups reaction.
As shown in the O 1 s + Sb 3d spectra (Fig. 8A, B), the proportion of
– O peak decreased from 69.6% in BC to 24.9% in Sb-loaded BC and
C–
the O–H peak decreased from 86.4% in CH1BC to 43.1% in Sb-loaded
–O
CH1BC with simultaneous peak location shifts, indicating that C–
and O–H are the respective dominating oxygen-containing group
involved in the sorption process for BC and CH1BC. The deconvolution
for Sb 3d was used to determine the oxidation state of Sb(III) on bio­
chars. However, Sb 3d5/2 scanning regions were overlapped by O 1 s.
Hence, 3d3/2 was chosen and deconvoluted into two peaks for both BC
and CH1BC (Fig. 8C, D). The binding energy for Sb 3d3/2 ranging from
538.8 to 539.4 eV, 539.8–541.7 eV was respectively assigned to Sb(III)
and Sb(V) within Sb oxides (Wu et al., 2019). The Sb(V) accounted for
86.1% in Sb-loaded CH1BC, whereas the proportion of Sb(V) in BC was
70.8% (Fig. 8C, D), suggesting the stronger oxidative ability of CH1BC
for Sb(III). It seems convincible that the presence of reductive groups, e.
g., –OH and –NH2 in the CH1BC favored the direct oxidation process
through a redox reaction as indicated by Zhang et al. (2021), which was
the major reason for the greater Sb(III) oxidation ability for CH1BC. We
also assume that the persistent free radicals (PFRs) that existed on the
biochar surface may also contribute to the oxidation of Sb(III). In
particular, the lignin-rich materials such as the F. microcarpa branch are
supposed to generate the PFRs via homolytic cleavage of weak linkage
bonds (e.g., C–C and C–O) during pyrolysis (Yuan et al., 2022). Li et al.
(2019a) noted that the PFRs directly activated H2O2 to generate⋅OH and
further degraded naphthalene in water. The strong electron-donating
capacity of PFRs may reduce the molecular oxygen in solution into
reactive oxygen species (ROs,⋅O2-), and the presence of ROs facilitated
the Sb(III) oxidation, as reported by Cui et al. (2017).
3.7.2. FTIR analyses
– O position was noted in the Sb-sorbed BC and
A sharp shift of C–
– C–O and aromatic C–H bond (Fig. 7A). For Sb-loaded
weakened O–
CH1BC, the overlapping peak density of N–H/O–H and C–O bonds
weakened with position shifts. Moreover, the C–H bonds at 2932, 2873,
1560 cm− 1 and C–O at 1070 cm− 1 vanished. These results indicated that
these multiple functional groups were involved during the Sb(III)
sorption (Cui et al., 2017). Instead, four new peaks at 1026, 819, 756
and 563 cm− 1 (Fig. 7A) were observed. The emerging peak (1026 cm− 1)
could be assigned to the [δ(O–H)] deformation bands of solid hexahy­
droxy antimonate salts (Jia et al., 2020). The observable peak at
819 cm− 1 corresponds to the stretching vibration of O–Sb, while the
peak at 756 cm− 1 is assigned to Sb–O–Sb peak (Wei et al., 2020), and the
latter was similar to the Sb–O–Sb vibrations in Sb2O3. Additionally, the
new peak (563 cm− 1) might be the characteristic peak of O–Sb
stretching vibration (Cui et al., 2017). Therefore, we conclude that the
Sb sorbed on CH1BC mainly existed in the form of both Sb(III) and Sb(V)
according to the emerging stretching vibrations (Jia et al., 2020; Wei
et al., 2020).
The FTIR results suggested that the sorption mechanisms for BC and
CH1BC differed. As in BC, π-π interaction and surface complexation are
the most responsible mechanisms for Sb(III) sorption (Cui et al., 2017;
Chen et al., 2021). First, a π electron in the aromatic C–H bonds on BC
can be combined with another π electron in Sb(III), and hence sorb Sb
(III) onto BC. Second, the formation of surface complexes between the
– O and Sb(III)
protonated oxygen-contained functional groups such as C–
favored its sorption in BC.
The Sb(III) sorbed onto CH1BC also included these mechanisms via
π-π interaction with C–H bonds and surface complexation with carboxyl
(Cui et al., 2017; Wei et al., 2020). Chelation interaction between N–H
bonds in the CH1BC and Sb(III) also contributed to the sorption process
as indicated by Wei et al. (2020). Additionally, the hydroxyl groups on
the biochar surface have been recognized as hydrogen donors, forming
hydrogen bonding with oxygen atoms on the Sb(III) molecule (Xiong
et al., 2020). The decline of –OH peaks suggested that hydrogen bonding
might be a crucial factor influencing the Sb(III) sorption by CH1BC.
Furthermore, the presence of Sb(V) manifested the CH1BC oxidized Sb
(III) to Sb(V) and thus facilitated the sorption.
3.7.4. DFT calculations
Density functional theory (DFT) calculations were employed to
clarify the adsorption capacity of various functional groups towards Sb
(III) fixation, as shown in Fig. 9A, in which 7 sites (Basal Plane, Ring­
– O, –COOH–HO–). These
–COOH, –COOH, –NH2, –OH–NH2, –OH–C–
groups are particularly investigated because they are dominating species
on the biochar (Fig. S3). According to calculated electron density (see
Fig. 9A), O- and N-groups at the edges present high density, which are
negatively charged and offer strong capacity to sorb Sb(III). Using
adsorption energy (ΔE) as an indicator, active sites with large negative
ΔE can be identified as promising structure for Sb(III) sorption (Zhang
et al., 2019a). According to Fig. 9B, three features can be summarized:
(i) both O- and N-groups are beneficial to strengthen Sb(III) sorption
because calculated ΔE over these sites is much higher than that over
Basal Plane (ΔE = − 0.97 eV); (ii) for BC, Ring-COOH can slightly
improve Sb(III) sorption with respect to sole -COOH, highlighting the
key role of Ring–COOH in the removal of Sb(III); (iii) for the
chitosan-loaded biochar, calculated ΔE = − 2.01 eV over amino groups
(–NH2) is ranked at the second place (see Fig. 9B), confirming the che­
lation interaction between N–H bonds with Sb(III) during the sorption
on CHBC as revealed by XPS analysis (Fig. 7F).
Noteworthily, newly-loaded hydroxyl on the CHBC could trigger a
synergistic effect on Sb(III) sorption with the carbonyl group in the
– O coupled through hydrogen
pristine biochar. Particularly, the –OH–C–
bonding offers the strongest adsorption (ΔE = − 2.52 eV), which vividly
demonstrates a robust evidence for the huge enhancement of Sb(III) in
the chitosan-loaded biochar. Based on batch sorption experiments,
characterization analyses, and theoretical calculations, the sorption
mechanisms of Sb(III) by BC and CHBCs were depicted in Fig. 10,
3.7.3. XPS analyses
The appearance of Sb 3d peak on the Sb-loaded BC/CH1BC
confirmed the sorption of Sb on the biochar surface (Fig. 7B). The Sb 3d
state could be split into Sb 3d5/2 and Sb 3d3/2 states corresponding to the
two peaks at the binding energy of 531.8 and 540.8 eV, respectively
(Luo et al., 2015; Xiong et al., 2020). The peak of 540.8 eV could be
assigned to Sb(V) (Wu et al., 2019), indicating the oxidation ability of
biochars to Sb(III). The peaks of C 1 s and N 1 s of BC/CH1BC were
weakened after Sb(III) sorption, and this implied the C/N-containing
functional groups might be involved.
The C 1 s spectrum of BC demonstrated that the C–C peak proportion
decreased from 68.2% to 59.0% after Sb(III) sorption (Fig. 7C), which
– C bond. The
could be due to the π-π interaction between Sb(III) and C–
peak shift for C–O implied that the formation of C–O–Sb might occur.
The C 1 s spectrum of CH1BC showed the disappearance of the C–H peak
after sorption that may link the importance of C–H bonds in complex­
ation with Sb ions (Fig. 7D). In this respect, Cui et al. (2017) found that
the C–H and –CH2 disappeared after Sb(III) sorption owing to the Sb-π
interaction. The alteration of peaks at 286.67 and 288.61 eV also indi­
– O–N to Sb(III) sorption (Fig. 7D).
cated the contribution of C–O and C–
As shown in N 1 s spectra for Sb-sorbed BC/CH1BC (Fig. 7E, F), the
proportion of –NH2 reduced simultaneously from 68.9% to 35.0%,
77.7–47.0% for BC and CH1BC, respectively; and the peaks of –NH2 and
CH3CO–NH both shifted, indicating the consumption of free –NH2 and
CH3CO–NH, and further confirming the occurrence of chelation reaction
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H. Chen et al.
Journal of Hazardous Materials 425 (2022) 127971
Fig. 8. The XPS spectra of O 1 s + Sb 3d (A and B) and Sb 3d3/2 (C and D) of Sb-loaded BC and CH1BC. BC: pristine biochar; CH1BC: biochar loaded with chitosan at a
ratio of 1:1.
Fig. 9. Optimized geometry and electron density of carbon framework with O- and N-containing functional groups (A), in which O, H, N and C are shown as red,
white, blue and grey colors; DFT-calculated binding energies of Sb(III) over different adsorption sites.
including pore filling, electrostatic interaction, π-π interaction,
hydrogen bonding, surface complexation with surface functional groups
– O), and oxidation reaction. Besides,
(e.g., C–H bonds, hydroxyl, C–
chelation was involved in Sb(III) sorption by CHBCs due to the presence
of amino groups.
4. Conclusions
Biochar loaded with chitosan at a ratio of 1:1 (i.e, CH1BC) showed
much higher Sb(III) sorption capacity than that of pristine biochar (168
vs 10 mg g− 1), mainly due to enrichment of functional groups (e.g.,
– O, –OH, and –NH2) on biochar surface. Our results demonstrated
C–
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Journal of Hazardous Materials 425 (2022) 127971
Fig. 10. Diagram of the potential mechanisms of Sb(III) sorption by pristine and chitosan-loaded biochars.
that Sb(III) sorption mechanisms onto CHBCs included electrostatic
attraction, π-π interaction, surface complexation/chelation, hydrogen
bonding as well as redox transformation of Sb(III). Density functional
theory (DFT) calculations verified the key role of –NH2 in the Sb(III)
sorption by CHBCs, and also suggested that the enhancement of removal
efficiency could be ascribed to the synergistic effect of exogenous hy­
droxyl and inherent carbonyl. Desorption and regeneration experiments
of CH1BC indicated that CH1BC was a feasible sorbent with robust sta­
bility for Sb(III) and great reusability performance. Future studies are
needed to understand the contribution of electron-donating and
electron-mediating capacities of biochars in the Sb(III) oxidation pro­
cess. Moreover, the feasibility of using alternative low-cost and ecofriendly industrial-grade chitosan (chitin) for enhancement of biochar
needs to be explored to strengthen its practical application.
the work reported in this paper.
CRediT authorship contribution statement
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This study was financially supported by the National Key Research
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Appendix A. Supporting information
Supplementary data associated with this article can be found in the
online version at doi:10.1016/j.jhazmat.2021.127971.
Declaration of Competing Interest
The authors declare that they have no known competing financial
interests or personal relationships that could have appeared to influence
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