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Citation: Hanh, P.T.H.; Phoungthong,
K.; Chantrapromma, S.; Choto, P.;
Thanomsilp, C.; Siriwat, P.;
Wisittipanit, N.; Suwunwong, T.
Adsorption of Tetracycline by
Magnetic Mesoporous Silica Derived
from Bottom Ash—Biomass Power
Plant. Sustainability 2023,15, 4727.
https://doi.org/10.3390/su15064727
Academic Editors: Ken Kawamoto,
Tomonori Ishigaki and Hoang
Giang Nguyen
Received: 1 January 2023
Revised: 26 February 2023
Accepted: 27 February 2023
Published: 7 March 2023
Copyright: © 2023 by the authors.
Licensee MDPI, Basel, Switzerland.
This article is an open access article
distributed under the terms and
conditions of the Creative Commons
Attribution (CC BY) license (https://
creativecommons.org/licenses/by/
4.0/).
sustainability
Article
Adsorption of Tetracycline by Magnetic Mesoporous Silica
Derived from Bottom Ash—Biomass Power Plant
Phan Thi Hong Hanh 1, Khamphe Phoungthong 1, Suchada Chantrapromma 2, Patcharanan Choto 3,4,
Chuleeporn Thanomsilp 3, Piyanuch Siriwat 3, Nuttachat Wisittipanit 3,5 and Thitipone Suwunwong 3,4 ,*
1Industrial Ecology in Energy Research Center, Faculty of Environmental Management, Prince of Songkla
University, Songkhla 90112, Thailand
2Division of Physical Science, Faculty of Science, Prince of Songkla University, Songkhla 90112, Thailand
3School of Science, Mae Fah Luang University, Chiang Rai 57100, Thailand
4Center of Chemical Innovation for Sustainability, Mae Fah Luang University, Chiang Rai 57100, Thailand
5Department of Materials Engineering, School of Science, Mae Fah Luang University,
Chiang Rai 57100, Thailand
*Correspondence: thitipone.suw@mfu.ac.th; Tel.: +66-869658254
Abstract:
In recent years, the contamination of the aquatic environment with antibiotics, including
tetracyclines, has drawn much attention. Bottom ash (BA), a residue from the biomass power plant,
was used to synthesize the magnetic mesoporous silica (MMS) and was utilized as an adsorbent
for tetracycline (TC) removal from aqueous solutions. The MMS was characterized by Fourier
transform-infrared (FTIR), X-ray diffraction (XRD) pattern, and scanning electron microscopy (SEM).
Optimum conditions were obtained in overnight incubation at 60
◦
C, a pH of 6–8, and an adsorption
capacity of 276.74 mg/g. The isotherm and kinetic equations pointed to a Langmuir isotherm model
and pseudo-first-order kinetic optimum fitting models. Based on the very low values of entropy
changes (
∆
S
◦
), the negative value of enthalpy changes (
∆
H
◦
) (
−
15.94 kJ/mol), and the negative Gibbs
free-energy changes (∆G◦), the adsorption process was physisorption and spontaneous.
Keywords: antibiotic contamination; tetracycline; solid waste; adsorption mechanisms
1. Introduction
Antibiotics have been commonly used for the prevention and treatment of infectious
diseases, both in humans and animals. According to Cycon et al., the situation analyzed
in 75 countries, indicates that the antibiotics usage climbed by 65%, and the prediction is
that in 2030, the consumption of antibiotics will be higher by 200% compared with 2015 [
1
].
Among antibiotics, tetracycline (TC) is a conventional, inexpensive antibiotic having wide-
ranging antibacterial action that is frequently used in both human and veterinary medicine
to prevent infection [
2
]. Additionally, TC antibiotics are utilized as a growth stimulant
for animals and for agricultural purposes [
3
,
4
]. Moreover, TC constitutes one of the most
important antibiotic families, ranking second in production and usage worldwide [
5
,
6
].
Because of its high aqueous solubility, TC has been found in ecological communities such
as surface water and groundwater (ranging from 5.4 to 8.1 ng/L) [
7
], municipal solid waste
(greater than 100 ng/L) [
8
], and soil (ranging from 86.2 to 198.7
µ
g/kg) [
9
], and it can be eas-
ily transferred to other environments via aqueous matrixes [
10
], and like other antibiotics, it
has a structural framework called naphthol ring that makes it difficult to degrade [
11
]. Thus,
Cyco´n et al. [
1
] suggested that “the reason is that antibiotics are not completely metabolized
by humans and animals, and a large proportion of the administered drug is released as
the parent compound through feces and urine, discharging into domestic wastewater and
into the pits where slurries/sludges are deposited”. Gu and Karthikeyan [
12
] also found
that “in the statewide survey of wastewater treatment plants, the compound TC was the
most frequently detected antibiotic (among 25 antibiotics), being present in 80% of the
Sustainability 2023,15, 4727. https://doi.org/10.3390/su15064727 https://www.mdpi.com/journal/sustainability
Sustainability 2023,15, 4727 2 of 13
wastewater influent and effluent samples”. However, standard aqueous solution water
treatment and spontaneous biodegradation are ineffective in removing TC from aqueous
solution [
12
,
13
]. Therefore, techniques for the secure and efficient removal of TC from
liquid have received a great deal of interest. Methods for TC removal include oxidation [
14
],
photo electrocatalytic [
15
], degradation [
16
], membrane processing [
17
], adsorption [
18
],
permeation [
19
], flocculation [
19
], ozonation oxidation [
19
]. Among these, adsorption has
several advantages over other procedures, including ease of operation, inexpensive operat-
ing costs, and significant removal efficiency at extremely low TC content in wastewater and
water [
20
]. Particularly, no generation of toxicity of intermediated and by-products during
adsorption makes it a more attractive appealing method of treating TC [
21
]. Until now,
some studies investigated the adsorption and removal of tetracyclines on several materials
such as graphene oxide [
22
], activated carbon [
23
], kaolinite [
24
], single-walled carbon
nanotubes, and multi-walled carbon nanotubes [
13
]. Carbon nanotubes and graphene,
which have a high graphite structure, have a strong TC elimination capacity [
25
]. However,
the high manufacturing, disposal, and regeneration costs of the aforementioned materials
would pose significant barriers in their practical application for TC adsorption, as well as
the capacity for large-scale application. As a result, there is still a considerable desire for
the development of efficient and affordable, simple-to-operate, high-selectivity devices.
Mesoporous silicas are often employed in a variety of applications, including catalysis,
separation, drug administration, chemical sensing, optic and electrical devices, rubber
reinforcement, and as a template in the production of nanomaterials [
26
]. Because they
possess a large surface area, big pore size, pore volume, and regular channel-type structures,
these properties are potential advantages that suit the adsorbate. A renewed interest in the
design of adsorbents [
27
] and catalysts has been sparked by the discovery of hexagonally
organized mesoporous silica [
28
], which has a distinctively large surface area, well-defined
pore size, and pore-shaped pores. For the production of surfactant-template silica materials,
which need an organic structure-directing agent or template is typically a single surfactant,
such as: the Pluronic-type surfactant; poly(ethylene oxide)-b-poly(propylene oxide)-b-
poly(ethylene oxide)-Pluronic P123 (EO
20
PO
70
EO
20
) or Pluronic F127 (EO
106
PO
60
EO
106
)
and cetyltrimethylammonium bromide (CTMAB) [
29
–
32
]. Some studies have magnetized
the silicates to allow for rapidly and effectively separating the adsorbents from the aqueous
solution by an external magnetic field, and incorporating magnetic elements within or on
mesoporous silicates has also enhanced the adsorption properties [33].
Some previous researches have investigated magnetic mesoporous silica (MMS) with
a variety of structures: including embedded [
31
], core-shell [
32
], and yolk-shell [
34
]. The
MMS could be synthesized by combining the mesoporous silica (MS) and magnetic particles.
Tetraethyl orthosilicate (TEOS) was used as the silica source, and magnetic particles were
prepared from iron (II) chloride tetrahydrate (FeCl
2
,4H
2
O), iron (III) chloride hexahydrate
(FeCl
3·
6H
2
O) [
35
]. Because TEOS is expensive and magnetic particles require time and
chemicals to make, searches for alternative cheap source are necessary.
Bottom ash (BA)—biomass power plant was a byproduct from the combusting of agri-
cultural waste to produce energy. Approximately 85–95 percent of BA was generated from
biomass power plants [
36
], with SiO
2
making up the majority of this ash [
37
]. Thus, these
residues were an ideal silica source. Currently, Thailand has an abundance of BA—biomass
power plants that may be utilized to create magnetic mesoporous silica as an appropriate
supply. Generally, utilizing the residues from incineration and combustion plant to produce
silica materials is feasible. Some studies have been successful in converting this byproduct
to zeolite in alkali solution [38], synthesizing zeolitic material and successfully separating
SiO
2
from municipal solid waste incineration (MSWI) ash [
39
], MCM-41, SBA-15, and
SBA-16 mesoporous silica from power plant bottom ash were successfully synthesized
for the first time in 2007 [
40
]. Mesoporous silica was prepared by the sol-gel method
from municipal solid waste incineration bottom ash [
41
] and industrial fly ash [
42
]. This
application would increase the usage of BA—biomass power plants, minimize pollution,
and effectively improve the quality of the wastewater process. However, according to our
Sustainability 2023,15, 4727 3 of 13
knowledge, there has not been any published research on the synthesis of magnetic silica
nanoparticles derived from BA for specific applications in the adsorption of tetracycline.
Therefore, the purpose of this study is to apply magnetic mesoporous silica synthesized
from biomass power plant ash to absorb TC from aqueous solutions. The properties of
the MMS including surface area, function groups, and crystal structure were analyzed.
The research was conducted on the isotherm model, kinetics model, and thermodynamic
characteristics of TC’s adsorption onto the MMS. According to the findings of adsorption
tests, the MMS demonstrated promising potential for TC removal from water. This study
provides a detailed technique for achieving the goal of treating trash with waste in addition
to describing how solid waste is utilized as a resource.
2. Materials and Methods
2.1. Materials
In this study, the bottom ash (BA) was obtained from the rubber biomass power
plant of Gulf Yala company, Yala, located in southern Thailand. The composition of
BA was determined by X-ray fluorescence spectrometer (XRF) as listed in Table S1 (in
the Supplementary material). Tetracycline hydrochloride, 96%, Alfa Aesar were used
without further purification. Hexadecyltrimethyl ammonium bromide (CTAB)
≥
98%,
Sigma, Cibolo, TX, USA; Sodium hydroxide (NaOH) 98%, Loba Chemie PVT.LTD, India;
Hydrochloric acid (HCl) 37%, Qrec, Newzealand; Ethanol 99.9%, Qrec, Newzealand.
2.2. Methods
2.2.1. Synthesis of Magnetic Mesoporous Silica
The biomass power plant ash (BA) which includes 55% SiO
2
content as the main
component (Table S1 in the Supplementary Material) was used as raw material for extraction
of SiO
2
compound. Briefly, BA sample was ground into small particles (<45
µ
m) and then
washed with DI water several times and dried in the oven at 90
◦
C for 24 h. The permanent
magnets were employed to separate the magnetic and nonmagnetic ashes. The magnetic
ash was served as the magnetic component in the magnetic mesoporous silica, and the
nonmagnetic ash was utilized to extract silica. The weight ratio between the magnetic and
nonmagnetic ashes was 1:10.
A total of 4 g of nonmagnetic ash was refluxed with 100 mL NaOH solution with the
concentration 4 M at 90
◦
C for 16 h in a round bottom flask. Then, the reaction solution was
filtered through a membrane of 0.45
µ
m and the supernatant was collected for adjusting
pH at 7 by 5 M HCl and aging at the room temperature for 24 h. The precipitated product
was washed with distilled water several times and dried at 90
◦
C for 24 h. To prepare the
sodium silicate solution, the obtained white power and NaOH (4:5 w/w) were precisely
weighed and then dissolved in 250 mL of distilled water at 80
◦
C for 1 day [
43
]. The white
power is SiO
2
and was extracted from BA, which was analyzed by X-ray fluorescence
spectrometer (XRF), as shown in Table S2 (in the Supplementary material).
Deionized water (20 mL) was added to 1.2 g hexadecyltrimethyl ammonium bromide
(CTAB) and then mixed at 40
◦
C until a clear solution was observed. The aqueous CTAB
solution was slowly mixed with sodium silicate solution and continuously stirred for
15 min. This mixture was added slowly into the magnetic ash; the ratio of the mass of
maghemite to Na
2
SiO
3
10% was 1:20, and it was stirred while being heated to 80
◦
C. After
further stirring for 30 min, the pH of the mixture was adjusted to 11 by 5 mol/L HCl
solution and then continuously stirred for 6 h. The mixture was kept in a water bath at
80
◦
C for 72 h. The ethanol 99.9%, 100 mL, was added to the precipitated product. To
eliminate CTAB, the mixture was sonicated for 30 min at 60
◦
C [
44
] and then washed with
DI water several times and dried at 80 ◦C for 24 h.
The experimental procedure was repeated more than three times to verify that the
results are reproducible.
Sustainability 2023,15, 4727 4 of 13
2.2.2. Characterization
The mineral compositions in BA and MMS were determined by X-ray fluorescence
(XRF), (S2175 Ranger, Bruker, Burladingen, Germany). Fourier transform infrared spec-
troscopy FT-IR (Perkin Elmer Model Spectrum GX) was used to analyze the specific func-
tional groups of the adsorbents by compressed samples into KBr pellets and then analyzed
with a Nicole IS10 spectrometer over the wavelength ranged from 400 to 4000 cm
−1
. X-ray
diffraction (XRD) pattern (PAN analytical, X’Pert Pro MPD) was performed on a Bruker
AXS Advance instrument for confirmation the structure. The surface morphology (SEM) of
magnetic mesoporous silica was examined by scanning electron microscopy (SEM, Apreo,
FEI, South Moravian Region, Czech Republic) running by Schottky field emission at the
accelerating voltage of 20 kV in high vacuum mode. The elemental mapping analysis
of the sample was recorded by energy-dispersive X-ray spectrometer (EDS)—(X-Max80,
Oxford, UK).
2.2.3. Adsorption Experiments
MMS was used for tetracyclines (TC) adsorption in an aqueous solution. To find the
ideal conditions, the impact of adsorption time and temperature on the unit adsorption
capacity and adsorption rate of TC was investigated. MMS (3 mg) was dispersed in a flask
containing 10.0 mL TC solution with various concentrations (10–100 mg/L), pH 6–8, and
then fully homogenized with a vortex mixer. The suspension was incubated overnight at
25
◦
C, 45
◦
C, and 60
◦
C and covered with aluminum foil to protect TC from the potential
photo degradation. MMS was then separated from the samples through a magnet after
centrifuging at 2000 rpm for 15, 30, 45, 60, and 90 min while maintaining the experiment
temperature including centrifugation. The residual concentration of TC in an aqueous
solution was determined by UV–vis absorbance at 357 nm, using a calibration curve
built. All the experiments were replicated thrice, and the averaged results were reported.
Equation (1) was used to calculate the percentage removal of TC, and Eq
uation (2
) was
used to determine the adsorption capacity:
Removal(%)=(Co−Ce)
Co
×100 (1)
Qe=(Co−Ce)×V
m(2)
where:
•Qeis the amount of TC adsorption per unit weight of adsorbent (mg/g);
•Cois the initial concentration of TCs (mg/mL);
•Ce(mg/L) is the equilibrium concentration of TC;
•Vis the solution volume (mL);
•mis the mass of adsorbent (g).
3. Results and Discussion
3.1. Characterization of MMS
The FT-IR peaks of the MMS before and after adsorption of TC were collected in the
range from 400 to 4000 cm
−1
(Figure 1), indicating the chemical bonds and functional
groups in the compound. The O–H stretching vibration was identified as the source of the
big broadband at 3440 to 3443 cm
−1
. The absorption peaks at 1640 cm
−1
were caused by
the symmetric and asymmetric bending vibration of C=O. Fe–O stretching was assigned to
the band below 700 cm
−1
. The characteristic absorption bands of the sample at 692 and
5
83 cm−1
were assigned to iron(II) oxide bending, and the band at 459 to 446 cm
−1
was
ascribed to the bending vibration mode of Fe
2
O
3
. The band at 1048 cm
−1
to 1051 cm
−1
,
which is associated with Si-O-Si antisymmetric stretching vibrations, is a sign that silicon
dioxide is present in the sample. The bands at 579 and 583 cm
−1
are also an indication of
the presence of Si–O–Fe [
45
]. Most of the adsorbent’s peaks remained constant after TC
Sustainability 2023,15, 4727 5 of 13
adsorption, revealing that the adsorption procedure did not modify the structure of material.
However, several distinctive peaks at 1478, 2852, and 2922 cm
−1
were characteristic peaks
of the C=C skeleton and C-H stretching vibration of CH
2
and CH
3
induced by aromatic
groups of TC [
46
,
47
]. The FT-IR spectrum proved that TC was adsorbed onto the adsorbent
and the sample’s structure is mostly stable after adsorption [48].
Sustainability 2022, 14, x FOR PEER REVIEW 5 of 14
groups in the compound. The O–H stretching vibration was identified as the source of the
big broadband at 3440 to 3443 cm—1. The absorption peaks at 1640 cm—1 were caused by
the symmetric and asymmetric bending vibration of C=O. Fe–O stretching was assigned
to the band below 700 cm—1. The characteristic absorption bands of the sample at 692 and
583 cm—1 were assigned to iron(II) oxide bending, and the band at 459 to 446 cm—1 was
ascribed to the bending vibration mode of Fe2O3. The band at 1048 cm—1 to 1051 cm—1,
which is associated with Si-O-Si antisymmetric stretching vibrations, is a sign that silicon
dioxide is present in the sample. The bands at 579 and 583 cm—1 are also an indication of
the presence of Si–O–Fe [45]. Most of the adsorbent’s peaks remained constant after TC
adsorption, revealing that the adsorption procedure did not modify the structure of ma-
terial. However, several distinctive peaks at 1478, 2852, and 2922 cm—1 were characteristic
peaks of the C=C skeleton and C-H stretching vibration of CH2 and CH3 induced by aro-
matic groups of TC [46,47]. The FT-IR spectrum proved that TC was adsorbed onto the
adsorbent and the sample’s structure is mostly stable after adsorption [48].
Figure 1. FTIR spectra of the adsorbent before and after adsorption of TC.
The XRD profile of the MMS in the FT-IR spectrum proved that TC was adsorbed
onto the adsorbent and the sample’s structure is mostly stable after adsorption [48].
Comparisons with those of SiO2 (extracted from biomass power plant Ash) and Fe2O3
peak [49] are shown in Figure 2. The XRD pattern of the MMS exhibited the characteristic
diffraction peaks of Fe2O3 with weak intensity due to the lower concentration of Fe2O3. In
addition, the relatively slightly broad peak observed at 15−30° arose from the SiO2,
demonstrating that the crystalline structure of the iron oxide was retained after the encap-
sulation in silica.
Figure 1. FTIR spectra of the adsorbent before and after adsorption of TC.
The XRD profile of the MMS in the FT-IR spectrum proved that TC was adsorbed onto
the adsorbent and the sample’s structure is mostly stable after adsorption [48].
Comparisons with those of SiO
2
(extracted from biomass power plant Ash) and
Fe
2
O
3
peak [
49
] are shown in Figure 2. The XRD pattern of the MMS exhibited the
characteristic diffraction peaks of Fe
2
O
3
with weak intensity due to the lower concentration
of Fe
2
O
3
. In addition, the relatively slightly broad peak observed at 15
−
30
◦
arose from the
SiO
2
, demonstrating that the crystalline structure of the iron oxide was retained after the
encapsulation in silica.
Sustainability 2022, 14, x FOR PEER REVIEW 6 of 14
Figure 2. XRD patterns of the MMS.
The morphological structures of the MMS were examined using SEM. As shown in
Figure 3a, there are many unobvious spherical particles (aggregated) with a size of around
50−80 nm. A rough surface, which mainly consists of not−well−crystallized Fe2O3, seems
not to be efficient to achieve homogeneous coating SiO2, but instead irregular and severely
agglomerated particles are observed in Figure 3b,c.
Figure 3. SEM images of MMS (a–c), and EDS curve of MMS (d).
Figure 2. XRD patterns of the MMS.
Sustainability 2023,15, 4727 6 of 13
The morphological structures of the MMS were examined using SEM. As shown in
Figure 3a, there are many unobvious spherical particles (aggregated) with a size of around
50
−
80 nm. A rough surface, which mainly consists of not
−
well
−
crystallized Fe
2
O
3
, seems
not to be efficient to achieve homogeneous coating SiO
2
, but instead irregular and severely
agglomerated particles are observed in Figure 3b,c.
Sustainability 2022, 14, x FOR PEER REVIEW 6 of 14
Figure 2. XRD patterns of the MMS.
The morphological structures of the MMS were examined using SEM. As shown in
Figure 3a, there are many unobvious spherical particles (aggregated) with a size of around
50−80 nm. A rough surface, which mainly consists of not−well−crystallized Fe2O3, seems
not to be efficient to achieve homogeneous coating SiO2, but instead irregular and severely
agglomerated particles are observed in Figure 3b,c.
Figure 3. SEM images of MMS (a–c), and EDS curve of MMS (d).
Figure 3. SEM images of MMS (a–c), and EDS curve of MMS (d).
The samples typically determined the elemental analysis or chemical properties using
energy-dispersive X
−
ray spectrometry (EDS). The various elements in the sample are rep-
resented by energy peaks. As-prepared nanocomposites were found to include significant
amounts of Fe, O, Si, Ti, Na, Ca, Al, etc. (Figure 3d), confirming the formation of additional
nanocrystals on the surface of Fe2O3@SiO2particles.
3.2. Adsorption Isotherms of TC on MMS
In order to elucidate the interactions between the adsorption ability of the MMS and
TC in solution at adsorption equilibrium, the data were modeled using four adsorption
isotherms, including Langmuir, Freundlich, Temkin, and Sips isotherm models at 25, 45,
and 60 ◦C. The isotherm models were listed below with Equations (3)–(6) as follows.
Qe=QmKLCe
1+KLCe;RL=1
1+KoCo(3)
Qe=KfC
1
n f
e(4)
Qe=RT
bT
ln(KTCe)(5)
Qe=Qm
(KSCe)a
1+ (KSCe)a(6)
where Q
m
= maximum adsorption ability (mg/g), C
e
= equilibrium concentration (mg/L),
Q
e
= adsorption capacity (mg/g) at equilibrium time, K
L
= Langmuir constant, R
L
= sepa-
ration constant, K
f
and n= Freundlich constant, R= universal gas constant (
8.314 J/mol
),
Sustainability 2023,15, 4727 7 of 13
T= tempe
rature in terms of Kelvin, b
T
= Temkin constant, K
T
= equilibrium bond constant
related to the maximum energy of bond, and Ksand aare Sips constants.
According to Figure 4a, as the equilibrium TC concentration climbed, the TC’s capacity
to adsorb on the MMS also dramatically increased. It was determined that the adsorption
process was endothermic because the adsorption capacity of TC increased as the adsorption
temperature rose. This is likely because the higher temperature supported the quantity of
activated functional groups on the surface of the MMS or sped up the diffusion rate of the
tetracycline molecules.
Sustainability 2022, 14, x FOR PEER REVIEW 8 of 14
Figure 4. Adsorption isotherms of TC onto MMS at 25, 45, and 60 °C (a), The nonlinear fits for Lang-
muir, Freundlich, Temkin, Sips models at 25 °C (b), 45 °C (c), 60 °C (d). (adsorbent dosage = 0.03
g/L, pH = 8, V = 10 mL).
Table 1. Parameters of adsorption isotherm model of TC onto MMS at 25, 45, 60 °C.
Temp (°C) 25 °C 45 °C 60 °C
Langmuir
Qm 18.11 ± 2.65 43.92 ± 15.67 276.74 ± 159.62
KL 0.031 ± 0.009 0.009 ± 0.004 0.002 ± 0.001
R2 0.9832 0.9876 0.9815
SSE 1.58 1.74 0.19
χ2 0.53 0.53 0.53
Freundlich
n 0.54 ± 0.003 0.79 ± 0.001 0.96 ± 3.25 × 10−4
KF 1.31 ± 0.01 0.64 ± 0.003 0.56 ± 6.20 × 10−4
R2 0.9867 0.9983 0.9999
SSE 138.27 32.07 1.66
χ2 0.14 0.03 0.002
Temkin
AT 0.58 ± 0.01 0.58 ± 0.01 0.88 ± 0.03
BT 790.61 ± 7.53 790.62 ± 7.53 909.52 ± 10.28
R2 0.9181 0.9181 0.8880
SSE 855.40 3806.77 8002.25
χ2 0.86 3.89 8.03
Sips
Qm 12.92 ± 0.11 14.52 ± 0.17 17.07 ± 0.24
KS 0.18 ± 0.02 0.35 ± 0.08 0.47 ± 0.13
A 2.09 ± 0.18 3.79 ± 0.77 3.63 ± 0.87
R2 0.9603 0.9415 0.9218
SSE 512.81 1055.65 2210.77
χ2 0.51 1.06 2.22
Figure 4.
Adsorption isotherms of TC onto MMS at 25, 45, and 60
◦
C (
a
), The nonlinear fits for
Langmuir, Freundlich, Temkin, Sips models at 25
◦
C (
b
), 45
◦
C (
c
), 60
◦
C (
d
). (adsorbent dosage =
0.03 g/L, pH = 8, V= 10 mL).
The starting TC concentration used in the modeling procedure varied from 10 to
1
00 mg/L
, and the adsorption temperature was adjusted at 25, 45, and 60
◦
C. Figure 4b–d
show the plot of the nonlinear fits for the aforementioned four isotherm models, and Table 1
lists the related parameter values. The Langmuir model was more appropriate to explain
the adsorption of TC onto the MMS, as shown by the correlation coefficients, R
2
, which
were consistently close to 1 and the sum of square error (SSE), chi-square (
χ2
) values, which
suggest that the surface of the MMS was homogeneous and the adsorption was monolayer.
Additionally, K
L
is a significant evaluation factor in relation to the binding site affinities.
The value of K
L
was between 0–1 and declined with the rising temperature, and separation
factor (R
L
) <1 reflected that the adsorption of TC onto the MMS was favorable [
50
]. The
nonlinear fitting of the Langmuir model (Figure 4b) showed the adsorption capacity (Q
m
)
of MMS for TC was 276.74 mg/g.
Sustainability 2023,15, 4727 8 of 13
Table 1. Parameters of adsorption isotherm model of TC onto MMS at 25, 45, 60 ◦C.
Temp (◦C) 25 ◦C 45 ◦C 60 ◦C
Langmuir
Qm18.11 ±2.65 43.92 ±15.67 276.74 ±159.62
KL0.031 ±0.009 0.009 ±0.004 0.002 ±0.001
R20.9832 0.9876 0.9815
SSE 1.58 1.74 0.19
χ20.53 0.53 0.53
Freundlich
n0.54 ±0.003 0.79 ±0.001 0.96 ±3.25 ×10−4
KF1.31 ±0.01 0.64 ±0.003 0.56 ±6.20 ×10−4
R20.9867 0.9983 0.9999
SSE 138.27 32.07 1.66
χ20.14 0.03 0.002
Temkin
AT0.58 ±0.01 0.58 ±0.01 0.88 ±0.03
BT790.61 ±7.53 790.62 ±7.53 909.52 ±10.28
R20.9181 0.9181 0.8880
SSE 855.40 3806.77 8002.25
χ20.86 3.89 8.03
Sips
Qm12.92 ±0.11 14.52 ±0.17 17.07 ±0.24
KS0.18 ±0.02 0.35 ±0.08 0.47 ±0.13
A2.09 ±0.18 3.79 ±0.77 3.63 ±0.87
R20.9603 0.9415 0.9218
SSE 512.81 1055.65 2210.77
χ20.51 1.06 2.22
The R
2
values of the Freundlich isotherms were from 0.98 to 0.99, but the 1/nvalues
were more than 1, at 1.85, 1.26, and 1.04, with increased temperatures at 25, 45, 60
◦
C,
respectively, suggesting the adsorption is not prone to occur. Therefore, the Freundlich
isotherm had a hard time accurately explaining the TC adsorption pattern.
The Temkin and Sips isotherms were a poor fit, although heterogeneity was also
assumed. The Temkin isotherms illustrate that the heat of adsorption is negatively pro-
portional to the surface area covered by the adsorbate molecules. The Temkin’s constant,
abbreviated B
T
, represents the heat generated during adsorption. The fact that B
T
was
positive (790.61–909.52 J/mol) for the TC adsorption from the aqueous solution indicates
that TC was endothermically adsorbed on the MMS. The equilibrium binding constant A
T
is the quantity that relates to the greatest binding energy. When the temperature was raised
from 25 to 60
◦
C, the values of A
T
also increased from 0.58 to 0.88 L/mg, indicating that the
adsorption of TC on the MMS was endothermic.
Equations (7) and (8) of the van’t Hoff equation were used to compute the Gibbs
free-energy change
∆
G
◦
enthalpy
∆
H
◦
and entropy
∆
S
◦
during the adsorption process (8).
The results are displayed in Table 2.
∆Go=−RTlnKL(7)
ln(KL)=∆So
R−∆Ho
RT (8)
where Ris the ideal gas constant 8.314 J/(mol
·
K), Tis Kelvin temperature (K), and K
L
is the
Langmuir isotherm equilibrium constant (L/mg) [50–53].
Sustainability 2023,15, 4727 9 of 13
Table 2. Thermodynamic parameters of TC adsorption onto MMS.
Temperature (K) ∆G◦(kJ/mol) ∆H◦(kJ/mol) ∆S◦(J/(mol.K)
298 −23.61
−7.62 −15.94
318 −21.92
333 −18.79
The absolute values of
∆
G
◦
are negative for all the parameter intervals, indicating that
the adsorption behavior is spontaneous.
∆
G
◦
also increased as the adsorption temperature
rose, suggesting that the adsorption process was impeded [
51
]. Moreover, the value of
∆
G
◦
in the range from
−
20 to
−
80 (kJ/mol) showed that physisorption was involved in the
process, but a minor chemical action effect could speed up the procedure. Furthermore, the
value of
∆
H
◦
lower than 20 (kJ/mol) revealed the physical adsorption with van der Waals
interaction implied in the process mechanism [
52
]. A very low
∆
S
◦
value
(−15.94 J/(mol.K))
proved a little change in entropy occurred during the adsorption of TC by the MMS [53].
3.3. Adsorption Kinetics of TC on MMS
The adsorption kinetics explained the impact of time and temperature on the TC
adsorption rate. The adsorption kinetics test was conducted by adding 0.03 g of adsorbent to
10 mL of 100 mg/L TC solution at 45
◦
C. The adsorption data were fitted using three popular
kinetic models, including the pseudo-first-order, pseudo-second-order, and Elovich models.
The kinetic models are listed in the Equations (9)–(11) respectively, the corresponding
nonlinear curves are shown in Figure 5, and Table 3contains the estimated and shown
kinetic parameters.
Qt=Qe1−e−K1t(9)
Qt=KeQ2
et
1+K2Qet(10)
Qt=1
bln(ab)+1
bln(t); (11)
Sustainability 2022, 14, x FOR PEER REVIEW 10 of 14
𝑄=(
)𝑙𝑛(𝑎𝑏) + (
)𝑙𝑛(𝑡); (11)
Table 3. Adsorption Kinetic Coefficients.
Model Parameter Value
pseudo-1st-order
K
1 0.14 ± 0.02
Qe 1.44 ± 0.02
R2 0.9971
SSE 0.005
χ2 0.00
pseudo-2nd-order
Qe 1.56 ± 0.003
K2 0.14 ± 0.002
R2 0.9646
SSE 2.44
χ2 0.00
Elovich
β 1.33 ± 0.05
α 3.90 ± 0.03
R2 0.9559
SSE 4.20
χ2 4.20
Figure 5. Adsorption kinetics of TC onto MMS at 45 °C, with initial TC concentration = 100 mg/L.
For the pseudo-second-order and Elovich models, R2 was considerably less than 1,
and SSE, χ2 were higher. It was determined that neither model could adequately explain
the experimental findings. However, the pseudo-first-order kinetic model showed a high
value of R2 (0.9971) and a small value of SSE, χ2.The closer fitting curves imply that the
pseudo-first-order adsorption process was dominant for TC adsorption on the MMS.
As presented in Table 4, the MMS has a high adsorption capacity when compared to
the Qm values of a variety of adsorbents utilized from waste for TC antibiotics in aquatic
environments. From this table, the MMS can be considered as an alternative adsorbent
material for TC adsorption.
Figure 5. Adsorption kinetics of TC onto MMS at 45 ◦C, with initial TC concentration = 100 mg/L.
Sustainability 2023,15, 4727 10 of 13
Table 3. Adsorption Kinetic Coefficients.
Model Parameter Value
pseudo-1st-order
K10.14 ±0.02
Qe1.44 ±0.02
R20.9971
SSE 0.005
χ20.00
pseudo-2nd-order
Qe1.56 ±0.003
K20.14 ±0.002
R20.9646
SSE 2.44
χ20.00
Elovich
β1.33 ±0.05
α3.90 ±0.03
R20.9559
SSE 4.20
χ24.20
For the pseudo-second-order and Elovich models, R
2
was considerably less than 1,
and SSE,
χ2
were higher. It was determined that neither model could adequately explain
the experimental findings. However, the pseudo-first-order kinetic model showed a high
value of R
2
(0.9971) and a small value of SSE,
χ2
.The closer fitting curves imply that the
pseudo-first-order adsorption process was dominant for TC adsorption on the MMS.
As presented in Table 4, the MMS has a high adsorption capacity when compared to
the Q
m
values of a variety of adsorbents utilized from waste for TC antibiotics in aquatic
environments. From this table, the MMS can be considered as an alternative adsorbent
material for TC adsorption.
Table 4. The capacity adsorption of TC with other different adsorbents’ utilization from waste.
Adsorbent Qm(mg g−1) Refs.
Alkali modified magnetic biochar (MSBC) 97.962
[54]
Alkali-acid modified magnetic biochar (MSABC) 98.334
The raw biochar (RBC) 37.803
Alfalfa-derived biochar 372 [55]
Pinus taeda-derived activated biochar 274.8 [56]
Waste chicken-feather-derived multilayered
graphene-phase biochar 388.33 [57]
Clay-biochar composites 77.962 [58]
Spent coffee-ground-derived biochar 39.22 [59]
Biomass ash pyrolyzed from municipal sludge 50.75 [60]
Shrimp Shell Waste 229.98 [61]
Magnetic mesoporous silica from BA-Biomass
Power Plant 276.74 In this study
4. Conclusions
Magnetic mesoporous silica that was fabricated from rubber biomass power plant ash
was considered as a low-price adsorbent for tetracycline adsorption in aqueous solution.
The adsorption of tetracycline onto the magnetic mesoporous silica was a monolayer en-
dothermic process. The maximum tetracycline adsorption capacity of magnetic mesoporous
silica in aqueous solutions was 276.74 mg/g at 60
◦
C. With the adsorption, the pseudo-first-
order model accurately reflects the data on adsorption kinetics, while the Langmuir model
for adsorption matches the data on adsorption isotherms well. Tetracycline adsorption
Sustainability 2023,15, 4727 11 of 13
by the MMS was spontaneous due to the negative values of the Gibbs free-energy and
enthalpy change, and at the higher temperature the adsorption was adversely affected.
The adsorbent showed exceptional properties, including tetracycline removal efficiency
and easy separation from aqueous media by a magnet. Therefore, magnetic mesoporous
silica could be used as a potential adsorbent for tetracycline from an aqueous solution.
Furthermore, this adsorbent can be utilized as an environmentally friendly adsorbent for
the treatment of wastewater.
Supplementary Materials:
The following supporting information can be downloaded at: https:
//www.mdpi.com/article/10.3390/su15064727/s1, Table S1: Chemical compositions of Bottom Ash -
Biomass Power Plant by XRF; Table S2: Chemical composition of SiO
2
extracted from Bottom Ash -
Biomass Power Plant by XRF.
Author Contributions:
Conceptualization, K.P., S.C., P.C., C.T., P.S., N.W. and T.S.; methodology,
P.T.H.H., P.S. and N.W.; validation, S.C., P.S. and N.W.; formal analysis, P.T.H.H., K.P., C.T. and P.S.;
investigation, P.C., C.T., P.S., N.W. and T.S.; resources, S.C., C.T., P.S. and N.W.; data curation, P.T.H.H.;
writing—original draft preparation, P.T.H.H.; writing—review and editing, P.T.H.H., K.P., S.C., P.C.,
N.W. and T.S.; visualization, K.P., C.T. and P.S.; project administration, N.W. and T.S. All authors have
read and agreed to the published version of the manuscript.
Funding:
This research was funded by Thailand Science Research and Innovation (TSRI), grant
number 652A01012 and Mae Fah Luang University.
Institutional Review Board Statement: Not applicable.
Informed Consent Statement: Not applicable.
Data Availability Statement: Not applicable.
Acknowledgments:
We would like to acknowledge the Graduate School and Faculty of Environmen-
tal Management, Prince of Songkla University for this study and carrying out this research.
Conflicts of Interest: The authors declare no conflict of interest.
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