Article Type: Research Article Article Citation: Mehmet Uğurlu, Huseyn
Osman, Ali imran Vaizoğullar,
and Abdul Chaudhary. (2021). FLUOROQUINOLONES ANTIBIOTICS ADSORPTION ONTO
POLYMER COATED MAGNETIC NANOPARTICULAR ACTIVATED CARBON. International Journal
of Engineering Science Technologies, 5(2), 81-104. https://doi.org/10.29121/IJOEST.v5.i2.2021.172 Received Date: 13 March 2021 Accepted Date: 11 April 2021 Keywords: Activated Carbon Adsorption Magnetic
Adsorbent Fluoroquinolone
(FLQ Isotherm The present study investigated the adsorption of molecular fluoroquinolone (FLQ) from aqueous solution onto active carbon (AC), magnetic activated carbon (MagAC), styrene-butadiene styrene magnetic activated carbon (SBS/MagAC) and poly charbon magnetic activated carbon (PC/MagAC) as adsorbent materials. The process optimization was carried by investigating the effects of pH, temperature, solid-liquid ratio, adsorbent type and initial concentration of FLQ. The data showed that adsorption reached equilibrium in as little as one hour. The adsorption cacapcity was comparatively less at low pH values than at approximately pH 5.0. The results also showed that the polymer coated magnetic materials did not perform very well at high pH values. However, all the materials performed well at room temperature when the situation was examined in terms of kinetics. It was also observed that AC, SBS/MagAC and PC/MagAC are more effective than MagAC to remove FLQ from aqueous medium. The kinetic data support pseudo-second-order model (r2 ⩾ 0.95) but showed very poor fit for pseudo-first-order model (r2 ≤ 0.90). Intra-particle model also showed that there were two separate stages in sorption process, namely, external diffusion and the diffusion of inter-particle. Adsorption isotherms for all adsorbends were fitted to Langmuire models more effectively than Freundlich models (r2 ⩾ 0.98). Thermodynamics parameters such as; free energy (ΔG0), enthalpy (ΔH0) and entropy (ΔS0) were also calculated. In conclusion, our results revealed that FLQ can be removed more easily from the aqueous medium by using magnetic and polymeric material.
1. INTRODUCTIONAntibiotics are widely used to treat or prevent
human and animal diseases, but antibiotics resistance has become a major global
public health issue in recent years [1] Fluoroquinolone antibacterial agents (FLQs)
are a group of potent synthetic antibiotics that are widely used in human and
veterinary medicines [2]. The widespread presence and potential
toxicities of FLQs necessitate a better understanding of their environmental
fate to properly assess their risks. In this sense,
the identification and determination of these compounds play a
very important role on the development of new wastewater treatment
technologies in order to reduce the micropollutants concentration levels in
wastewaters. Previously, biodegradation [3], photodegradation [4] and adsorption [5], [6] processes have been used for the removal of
many organic chemicals, including FLQ, from aquoes
streams. However, some of these methods are not effective in the removal
of FLQ as the occurrence of this
chemical in natural environment can affect the selection of genetic variants of
microorganisms, result in the development of drug resistant bacteria or
pathogens, and induce a risk to the ecosystem and human health. Therefore, the
conventional primary and secondary water treatment plants can
not remove or degrade FLQ efficiently. Hence, there is a need to develop
and optimise a feasible and effective method for the
removal of FLQ from aqueous effluent streams [7], [8]. In the literature study, fluoroquinolone
antibacterial agents (FLQs) were found to adsorb strongly to goethite with
50–76% of the added FLQ adsorbed under the experimental conditions [9], [10]. Another study involveed
the removal of flumequine from water by activated carbon fixed-bed columns and
represents the first study of this emerging contaminant in fixed bed adsorption
[11]. The adsorption of flumequine (fluoroquinolone
antibiotic) and copper (II) on an alkaline soil sample was studied at
macroscopic and molecular scales. The retained antibiotic amount onto the soil
surface increases (from 2- to >20-fold) with the copper concentration due to
the formation of a Cu(II)–flumequine ternary surface complex, which leads to
the accumulation of flumequine into soils [12]. The glycerol-based carbon materials (GBCM200,
GBCM300 and GBCM350) was used as adsorbents for the
removal of the antibiotic compounds flumequine and tetracycline from aqueous
solution [13]. As can be seen briefly from the explanations
above, many different methods can be used for the removal of FLQ from aqueous
solution. Among these, adsorption is one of the most effective methods and is
suitable for developing countries due to low chemicals and materials
consumption [14]. In particular, adsorption
has become a most widely used method for removing micropollutants, especially
hydrophilic compounds such as FLQ, because the adsorbent can be easily modified
with increased surface charge. However, the selectivity of the adsorption
method is not very good, requiring a specific
adsorbent of the surface modification of the common adsorbent [15]. The modification of the adsorbent is needed
to increase the efficiency and selectivity of FLQ removal by natural adsorbents
with a small specific surface area. The most problematic and undesirable
situation in the adsorption process is the removal and recovery of nano-based
adsorbent materials from the reaction systems. These nano-based materails can be realesed into
the ecosystem and must be removed for safe disposal. It has been reported that
this difficulty can be solved by simply and effectively exposing the reaction
system to external magnetic fields. However, this appaocah
can only be used for the separation of magnetic nanoscale materials [16], [17]. In literature studies, iron-based magnetic nanopowders are most widely used and preferred as material
candidates due to their magnetic and other distinctive properties [18], [19]. The application of modified activated carbon,
containing iron particles, is the commonly used adsorbent to remove a varirty of environmental contaminats
from various waste streams. For example,
activated carbon (AC) containing
iron components (FeO, Fe2O3 or
Fe3O4) have been recently used for the treatment of food waste [20], effluent streams containing various organic
compounds [21], Cr(VI) [22], antibiotics [23], different heavy metals and As(V) [24]. In addition, AC can also be used for
different purposes after processing with certain polymeric materials. In the
literature studies, effective mercury removal was carried out using polymer
coated activated carbon. In this study, polysulfide-rubber (PSR) polymer, a
sulfur-rich compound, was used to increase affinity to remove mercury with
activated carbon. Hg-Cl and Hg-S interactions on the activated carbon surface
of the chemical bond with mercury have been reported [25]. In all previous studies, it is observed that
there is no electrical charge on AC powders and it
cannot be controlled by electric or magnetic fields. It can also create
secondary contamination after adsorption processes and the presence of magnetic
charge would help to trap, restore and recycle AC and prevent its release into the
environment. In literature study, AC was obtained from rice husk and then
modified using magnetic material. The resultant adsorbent material has high
surface area (770 m2/g) and 2.78 emu/g saturation magnetization (Ms) with 23 % Fe3O4
coating. The material has high adsorption capacity and was successfully used
for the removal of Methylene Blue (MB) [26]. In another study, it has been reported that
activated carbon modifede using magnetic nanoparticle
(AC-Fe3O4 MNPs) have additional advantages over the
conventional AC materials. These modified AC materials are effective adsorbent
materials for the removal of various contaminants such as aniline from both
water and wastewater streams
[27]. We now report on the synthesis, characterisation and application of activated carbon
modified with magnetic nanoparticles for the removal of FLQ from aqueous
systems. Firstly. samples of active carbon (AC), magnetic acticated
carbon (MagAC), styrene-butadiene styrene magnetic
activated carbon (SBS/MagAC) and poly charbonat magnetic activated carbon (PC/MagAC)
were synthesized and characterized as new adsorbents. In FLQ adsorption
experiments, temperature, solid-liquid ratio, adsorbent type, initial
concentrations of FLQ and solution pH were investigated by carrying out
experiments on a constant shaker under similar experimental condistions.
The efficiency of different adsorbents were calculated
and compared by inverstigating adsorption kinetics
and thermodynamic parameters. 2.
Experimental
2.1. MATERIALS AND PREPARATION OF ADSORBENTS
2.1.1. ACTIVATED CARBON(AC) The activated carbon samples used in the study
were obtained commercially, Sigma-Aldrich, 242276-300MG, Other iron samples
were also obtained commercially (ZAG, ZK.100380.1000) and prepared according to
literature studies [28]. 2.1.2. MAGNETIC ACTIVATED CARBON (MAGAC) FeCl3 (1.08 g) and FeCl2
(2.40 g) were added to 150 ml of distilled water and shaken at 60-65 °C for 1
hour. A 5 g of activated carbon was then added to these samples and shaken at
the same temperature for 2 hours. In the final step, 5 g of NaOH was added
while stirring the mixture for 1 h and aged at overnight. After washing several
times with distilled water. it was dried in an oven to a constant weight. 2.1.3. STYRENE-BUTADIENE STYRENE MAGNETIC
ACTIVATED CARBON (SBS/MAGAC) SBS (1.0g) + THF/DMF 50ml (30:20) samples were
prepared and stirred at 60-65 °C for 2 hours. Then 5 g of activated carbon was
added to the same temperature for one hour. To this mixture, FeCl3
(1.08 g) + FeCl2 (2.40 g) was added and the
mixture was stirred at 60-65°C for 2 hours. In the final step, 5 g of NaOH was
added to the samples stirred for 1 hour and allowed to stand overnight. These
samples were washed several times with distilled water, then filtered and dried
to a constant weight in an oven. 2.1.4. POLY CHARBONAT MAGNETIC ACTIVATED
CARBON (PC/MAGAC) PC (1.0 g) and THF/DMF 50 ml (30:20) samples
were taken and stirred at 60-65 °C for 2 h. 5g of activated carbon was added
and stirred at the same temperature for one hour. FeCl3 (1.08 g) +
FeCl2 (2.40 g) was then added and the
mixture was stirred for 2 hours at 60-65°C. In the final step, 5 g of NaOH was
added and the mixture was stirred for 1 hour. After washing several times with
distilled water, the filtered samples were dried to a constant weight in an
oven. 2.2. CHARACTERIZATION PROCESSES
The characterization of the adsorbents were carried out using Perkin Elmer Lambda 35 UV-Vis
Spectrophotometer, Perkin Elmer Pyris 1 for
Thermogravimetric Analyzer, Perkin Elmer Diamond for Differential Scanning
Calorimeter, Jsm-7600f for Scanning Electron Microscope, Rigaku-Smart-Lab-X-Ray
Diffractometer for XRD, Fourier Transform Infrared (FTIR) Spectroscopy, Thermo
Scientific Nicolet Is10 and
Micromeritics TriStar II PLUS for BET analysis. 2.3. ADSORPTION EXPERIMENTS AND DETERMINATION OF FLUOROQUINOLONE (FLQ)
The standard stock FLQ solution was prepared by
dissolving 0.02g of FLQ in 100ml EtOH using a magnetic stirrer and the final
volume was made to 1,000 ml by adding 900 ml of pure water. Adsorption
experiments were carried out by using 200 ml of FLQ solution containing 50
mg/L. Changes in FLQ concentration were determined before and after the
adsorption process with a UV spectrophotometer at a constant wavelength of 244
nm (lmax
for FLQ). All concentrations were determined by using the calibration curve.
Commercially activated granular activated carbon (AC), MagAC,
SBS/MagAC and PC/MagAC
magnetic adsorbents were used in the experimental study. All adsorption
experiments (except experiments examining the pH effect) were carried out at
natural pH. Temperature, solid-liquid ratio,
adsorbent type, initial concentrations and solution pH
were selected as parameters to be investigated for the adsorption studies. In
experiments examining the pH effect, the pH of the solution was adjusted using
dilute HCl and NaOH solutions and monitored throughout the exoperiment
using a pH meter (WTW-Germany. PH 330i). Adsorption experiments were carried
out on a constant shaker with a cooling effect. The structure of all
synthesized materials, adsorption test apparatus and FLQ are shown in Figure 1,
Figure 2a and 2b. Figure 1: Synthesis of magnetic materials and
general flow chart of experimental design
Figure 2: Adsorption experiment setup (a) and
the chemical structure of FLQ (b) 2.4. ADSORPTION KINETICS
There are various kinetic models that
characterize the adsorption process which determine what kind of mechanism
plays a role in the adsorption of the pollutant to the adsorbent surface. The
kinetic study was performed using various kinetic models. The pseudo-first-order
(Eq.(1)) and pseudo-second-order (Eq.(2)) kinetic models proposed by Lagergren [29] and Ho and McKay [30] are described in the following equations
respectively [31]. (1) (2) In the above equations, qe
is the amount of adsorbed substance (mg/g) per gram of adsorbent at
equilibrium, qt is the amount of adsorbed substance (mg/g) per gram of
adsorbent at any instant. The k1 and k2 (min-1) are the adsorption rate constants of the
pseudo 1st and 2nd order kinetic models respectively. The intraparticle diffusion equation proposed
by Weber and Morris is given as follows [32].
(3) Here c and ki (in
mg/g min1/2) represent the intercept and the
intra-particle-diffusion rate constant. The half-time of adsorption t1/2
is defined as the time required for adsorption to reach half the equilibrium
value. This time is generally used as a measure of adsorption rate and is
calculated with the help of Equation (4). (4) 2.5. ADSORPTION ISOTHERMS
It is extremely important to understand the
surface properties and affinity of adsorbent and its interaction with the
adsorbate. In this study, the adsorption isotherms were conducted at optimized
experimental conditions. The physicochemical data to recognize the adsorption
mechanism was interpreted using both Langmuir and Freundlich isotherm models.
The following equations (Eq.(5) and Eq.(6)) represent the Langmuir and
Freundlich isotherm models [33]. Linear state of this equation
(5) (6) The above equations. Qo (in mg/g) is
the maximum monolayer adsorption capacity. Ce (in mg/L) is the
equilibrium concentration. b. KF and n are respectively called the
Langmuir and Freundlich constants. 2.6. ADSORPTION THERMODYNAMICS
For the adsorption process, enthalpy, entropy and free energy changes can be determined by the
equilibrium constant. These thermodynamic parameters are shown in the following
equations (Eq. (7) and Eq. (8) [33]. (7)
(8) Here. ΔGº standard Gibbs free energy.
ΔHº standard enthalpy and ΔSº standard entropy. ΔHº and ΔSº
are calculated from the slope of the graph of 1/T versus lnKe
and the cut-off point. respectively. Adsorption equilibrium constant can be
calculated using the following equation.
(9) Here, Cads is the concentration
(mg/L) of the adsorbed substance at equilibrium and Ce is the
concentration of the substance remaining in solution at equilibrium (mg/L). 3.
DISCUSSION
AND CONCLUSION
3.1. CHARACTERIZATION OF THE SYNTHESIZED ADSORBENT MATERIALS
All adsorbent materials (AC, MagAC, SBS/MagAC and PC/MagAC) were characterised using
various analytical techniques, for example, FTIR, SEM, EDS, XRD, TGA, DSC and
BET. All the related evaluations and
comparisons for different adsorbent materials are given in the following
subsections. 3.1.1. SEM/EDS IMAGES AND BET ANALYSIS The morphological structure of adsorbents can
give specific information about the adhesion characteristics and mechanism. The
SEM and EDS results of different adsorbents are given in Fig 3. From Fig 3, when the SEM images of the activated carbon are examined, a distinct porous structure is seen, with many cavities, and its outer surfaces are recessed and protruding. In MagAC and PC/MagAC materials, it is seen that the iron and polymers species attach to the porous surface of the activated carbon and the particles inside. SEM image of SBS/MagAC material show a clearer appearance than the other samples in which a significant change is observed. This is thought to be the result from the interaction of polymers with the pores of the activated carbon.
Figurer 3: SEM and EDS images of adsorbent
materials a) AC, b) MagAC, c) SBS/MagAC,
d) PC/MagAC In order to obtain the information about the
element structure of adsorbent materials, EDS analysis was performed from SEM
image. As shown in the EDS graph given in Fig 3, it is seen that Ca, Fe, C, and
O elements are present predominantly on the surface of the adsorbents except
activated carbon and with small amounts of Na and Mg elements. Trace amounts of
Cu and Si elements are also partially observed in SBS/MagAC
material. In addition, when the EDS analysis was examined the highest amount of
Ca and C elements were found in these samples. In addition, activated carbon
samples obtained by using various polymeric materials and surface and pore size
changes are given in Table 1. Table 1: Changes in surface and pore size by
loading magnetic and polymeric materials on the surface of activated carbon.
As can be seen in Table 1,
considering the surface area of AC and synthesized materials, it is
observed that all surface area parameters follow as AC>MagAC>PC/MagAC> SBS/MagAC. A similar
situation is observed when pore volume is examined. When the pore size data are
evaluated, the opposite situation is observed. This is attributed to the
polymer material reaching and penetrating into the
interior of the pores. 3.1.2. XRD IMAGES XRD analysis of magnetic materials obtained
with various polymeric materials using activated carbon samples are given in
Fig 4. Figure 4: XRD images of different adsorbent materials Fig 4 shows the XRD spectra of activated carbon
and activated carbon based catalysis samples. When XRD
spectra of pure activated carbon are examined, the characteristic 2 theta
degree with 20.82°, 26.42°, 29.37°, 42.82° and 62.26° are seen with sharp
peaks. This shows that AC has a regular crystal structure. The diffraction
peaks at 26.42 and 42.82 shows (002) and (100) planes
respectively. The highest peak intensity was observed at 26.42°. This shows
that AC grows in the direction of the surface (002) surface [34]. In addition, when compared with other
synthesized samples, significant decreases in peak intensity are observed. This
shows that AC samples with partially crystalline structure tend to turn into
amorphous structure over time. In other words, the diffraction peaks in the
other samples except AC were slightly wider. The decrease in the sharp peak
intensities of AC indicates that the composite components shift to a little
more amorphous structure and also indicates that the
particle size decreases to some extent [35]. When the diffraction peaks of the MagAC sample with magnetic content were examined, a
diffraction peak at 35.45° which iron components was observed besides the
characteristic activated carbon peaks. In all synthesized samples except
activated carbon, it is observed that iron peaks are formed as
a result of treatment with strong base to ensure complete precipitation
of iron ions. The angle of about 35° is seen in each sample. This shows that
there is no change in the crystal structure of the PC and SBS used on the Iron
components. When we look at SBS/MagAC sample, it is
seen that the main diffraction peak intensity decreases and Full Width at Half
Maximum (FWHM) value increases. This suggests that SBS provides an effective
dispersion on AC. This result shows that the particles of polymeric properties
are synthesized with high efficiency of activated carbon samples and that an
active dispersion of other magnetic and polymeric components on AC surface is
provided. 3.1.3. TGA ANALYSIS The relationship between temperature and mass
loss was investigated by using TGA analysis of magnetic materials obtained with
various polymeric materials using activated carbon samples and is given in Fig
5. Figure 5: TGA results of AC, MagAC, SBS/MagAC, and PC/MagAC adsorbents In short, TGA is a method used to examine the
ability of a substance to maintain its mass (thermal stability) under various
conditions. In other words, it is the continuous monitoring of the changes in
the mass of the substance depending on the temperature and evaluating this as a
function of the temperature. The data in Fig 5 show that the TGA values of AC
generally lose mass with increasing temperature and a significant peak change
occurs. Here, a significant peak change was observed at MagAC
and PC/MagAC samples at approximately 800°C and at
the other SBS/MagAC samples at 850°C. This shows that
the thermogravimetric method is dynamic that the system will never reach
equilibrium and that changes in the amorphous and crystalline structure can
occur with increasing temperature [36]. In addition, starting from 800 °C, mass loss
was evident and sharp in all adsorbents. This is particularly related to the
decomposition and change of the carbon skeleton found in polymer-coated
materials [37] 3.1.4. DSC ANALYSIS DSC analysis results of magnetic materials
obtained with various polymeric materials by using activated carbon samples are
given in Fig 6. Figure 6: DSC results of AC, MagAC, SBS/MagAC and PC/MagAC adsorbents The most important applications of
thermogravimetric methods are for polymers. The decomposition mechanisms of
various polymeric materials can be explained by the information obtained from
thermograms. In addition, the investigation behaviour
characteristic of each type polymer is used in the
identification of polymers. In the DSC process, the temperature of the sample
and reference is increased at a regular rate by measuring the amount of energy
absorbed or released while the sample is heated and cooled or maintained at a
constant temperature. In endothermic reactions heat goes into the sample but in
exothermic reactions heat flows out of the sample. The heat lost or recovered as a result of endothermic or exothermic reactions in the
sample is recovered. Furthermore, the heating rate is recorded as a function of
the sample temperature. When examined in Fig 6 it is seen that there are
endothermic peaks around 100°C in AC samples. In MagAC
sample, it is observed that one exothermic peak is formed
and it is increased affected by temperature. Exothermic peaks at 500°C in SBS/MagAC samples exothermic peaks at 100°C and 300°C and
endothermic peaks at PC/MagAC samples. In addition,
it is seen that all the samples of temperature resistance are generally lower [37]. 3.1.5. FT-IR IMAGES FT-IR spectra of polymeric coated magnetic
materials using activated carbon samples are given in Fig.7. (a) (b) Figure 7: FT-IR results a) AC vs MagAC, b) AC vs PC/MagAC, c) AC
vs SBS/MagAC adsorbents 1)
When
the FT-IR spectra of pure activated carbon were examined. it was observed that
the peak of –OH group observed at 3041.85 cm-1 shifted to 3013.96 cm-1
as a result of the interaction with Fe3O4
nanoparticles. Therefore, we can infer that adsorption occurs by physical
interaction with OH groups on the surface. However, the C-O single bond to
which the –OH group is bound and observed at 1216.36 cm-1 has been
strengthened by the weakening of the –OH group and shifted to 1223.16 cm-1.
This is another proof of the interaction of nanoparticles with this group. The
peaks of the –CH group observed at 2880.90 cm-1 of activated carbon in the free
state were significant in the peaks of the C = O group observed at 1714.43 cm-1
and the C = C tensile peaks at 1536.03 cm-1. the absence of a change
indicates that there was no interaction with these groups. 2)
When
the FT-IR spectrum of the product obtained by the interaction of activated
carbon coated with Fe2Fe3O4 nanoparticles with
polycarbonate polymer. the –OH peak observed at 3019.96 cm-1 shifted
to 3421.76 cm-1. This is due to the hydrogen bond between the -COO
group and the –OH molecules in the structure during the bonding of
polycarbonate to the MagAC composite material.
Another proof of this is the C=O bond peak observed at 1648.86 cm-1 in
polycarbonate bonded MagAC composite material. This
peak was weakened by hydrogen bond formation and appeared at 1648.86 cm-1. Fe-C bonds not observed in MagAC
composite material were observed at 603.39 cm-1 in polycarbonate bonded MagAC composite material which explains the bond formed
between the surface-coated Fe metal and polycarbonate polymer. C-O bond
observed at 1223.16 cm-1 in MagAC composite material
weakened during the interaction between polycarbonate and MagAC
composite material and appeared at 1171.22 cm-1. 3)
When
the IR spectrum of the product obtained by the interaction of Fe2Fe3O4
coated activated carbon with SBS polymer was observed. no change was observed
in the spectrum [38], [39]. 3.2. ADSORPTION RESULTS
Magnetic and polymer coated adsorbents were
synthesized by using activated carbon (AC) as a base material. The process optmisation was done by investigating the effects of
solid-liquid ratio, temperature, initial FLQ concentrations and solution pH
values under identical experimental conditions. The variation of the adsorbed
amount per gram weight of adsorbents over time was evaluated separately for
each parameter. 3.2.1. EFFECT OF SOLID-LIQUID RATIO To investigate the effect of solid/liquid
ratio, the removal rates of FLQ using adsorbent materials were plotted in Fig.
8.
Figure 8: The amount of FLQ adsorbed over
time in adsorption with different adsorbents 0.5g/L (a), 1.0 g/L (b), and 2.0
g/L (c) Conditions: Temperature = 298K; natural pH ; FLQ conc., = 37 mg/L;
constant mixing ) The adsorbent dosage influences the total
specific surface area of binding sites and thus
is an important factor affecting adsorption [40]. Fig 8 shows that the removal efficiency of
FLQ increased sinigicantly with an ıncrease in adsorbents dosage. However, due to the increase in all
solid-liquid ratios, there is a significant reduction in the amount of
adsorbed, while the amount of adsorbed per gram decreases as the amount of adsorbent increases. This situation is thought to
arise from the interaction of the magnetic material in the solution and the
properties of the adsorbent surface. In addition, although the adsorption
performance is high at low concentration, low removal at the end point, the
higher the adsorption rate per gram, the higher the amount of adsorbed
substance per gram. The best results in terms of performance are seen in AC,
PC/MagAC, SBS/MagAC and MagAC examples, respectively. In addition, since 1.0 g/L
adsorbent provided a significant increase in removal, subsequent experiments
were conducted taking this solid-liquid ratio into account. Table 2 compares
the adsorption capacities for FLQ on different types of adsorbents reported in
the literature. As can be seen, different adsorbents exhibit different
adsorption capacities for adsorption which indicates that adsorbent's
properties have a significant effect on its efficacy. The results show that the
adsorption capacity of the mangnetice compounds used in this study is higher in most cases
than reported in the literature. Table 2: Literature results of the adsorption
capacities of FQs onto different adsorbents
3.2.2. TEMPERATURE EFFECT Temperature is one of the most significant
factors affecting the adsorption of micro-aqueous media and hance it was
studied, considering three different temperature of 291 K, 298 K, 308 K. The removal values of FLQ using different
adsorbent materials are showed in Fig 9. The data in Figure 9 showed that the adsorption
rate increases in the first 5 minutes with the increase of temperature, and the
system reaches equilibrium at the end of about 90 minutes. It was observed that
the amount of material adsorbed at almost all temperatures reached equilibrium. When the data of all
adsorbents were examined at 291K temperature, the best performance was observed
in AC (37 mg/g) and the lowest efficiency was observed in SBS/MagAC samples (11 mg/g). In the experimental study
performed at room temperature, higher removal with AC was 37 mg/g and the
lowest removal with MagAC was 35 mg/g. At high
temperature (308K), AC was provided with 38 mg/g, again showing better results,
while the lowest rate was observed in SBS/MagAC (16
mg/g). This situation can be explained by the decrease in adsorption with the
decrease in temperature exothermically after a slight increase due to the
decrease in viscosity with the increase in temperature [51].
Figure 9: The amount of FLQ adsorbed over
time at different temperatures (291 K
(a), 298 K (b), 308 K (c) (1.0 g / L, natural pH, constant mixing speed,
37ppm)) 3.2.3. THE EFFECT OF INITIAL CONCENTRATION To examine the initial concentration effect,
changes in the amount of FLQ adsorbed per gram using adsorbent materials are
plotted in Fig. 10.
Figure 10:
The amount of FLQ adsorbed over time in adsorption at different
concentrations 37 mg/L (a), 60 mg/L (b),
100 mg/L (c) (1.0 g/L, 298 K, natural pH, constant mixing speed) The data in Figure 10 showed that, under
identical experimental conditions, the adsorption efficiency increases with increasing
the initial concentration of FLQ. It is seen that in the first 5 minutes a
rapid adsorption occurs at all initial concentrations and equilibrium is
reached in almost all species. In addition, it is observed that more pronounced
and superior adsorption performance is obtained in AC and SBS/MagAC adsorbent samples. In addition, the adsorption rate
is low with high concentration of PC/MagAC. In
general, the % yield value decreased although the amount of FLQ adsorbed
increased when the initial concentrations were increased. In addition, The adsorption reaching equilibrium in a short time can be
associated with the physical character of the interaction [52]. 3.2.4. THE EFFECT OF PH In order to investigate the effect of pH, the
changes that occur in the adsorption capacity of FLQ using AC and other
synthesized magnetic adsorbents are given in Fig.11.
Figure 11: The amount of FLQ adsorbed per gram
of adsorbent depending on the pH effect pH:3.0 (a), pH:5.0 b), pH:7.0 (c), pH 9 (d) (1.0 g/L, 298K, 37
mg/L, constant mixing speed) The
data in Fig. 11 showed that the removal efficiency decreased as the solution pH
increased from 3 to 9. In the adsorption with all adsorbents, it was determined
that the same behaviors were observed in the adsorbents at low pH values
except SBS/MagAC. Here, values
of 37 mg/L were obtained in both pH ranges. It was observed that
in neutral pH values, respectively, AC, MagAC, PC/MagAC and lowest SBS/MagAC
values. When the chemical structure of FLQ is examined, it is thought that it
can be significantly affected by the pH change, since it has different
functional groups and its solubility is higher in the
basic environment. Considering these features, it is
estimated that adsorption to different adsorbents may occur at different pH's. Figure 12: Aqueous phase acid–base speciation
of OFX, highlighting the predominance of the zwitterionic species at pKa1 <
pH < pKa2. It have been reported
flumequine pKa values of acidic [53]. So, it seems to exhibit weakly acid
properties for its carboxylic group and lacks of a
piperazine ring [54]. This originates from two phenomena: (i) flumequine was negatively charged above pH 6.5 (pKa = 6.2) [55] and (ii) the adsorbents surface carried more
negative charges with increasing pH. These two
phenomena led to electrostatic repulsions, which implies that variable charges
are implicated in the sorption mechanism. A similar trend was observed for
fluoroquinolones sorption onto soil constituents or whole soils [55]. Gu et al. [56] showed that ciprofloxacin sorption onto iron
and aluminum hydrous oxides was highly pH dependent in the 4–10 range and
followed the species distribution of the antibiotic. Therefore, the neutral
form of flumequine dominates at pH<pKa
and the anionic form is mainly present at pH > pKa.
Since, at pH. < pHPZC,
AC surface was determined been positively charged and the reverse for pH values
> pHPZC. Adsorption seems to decreased after at these pH ranges, which is near the
reported pKa values of FLQ (Fig. 12). In
other words, The typical environmental pH of
antibiotic water environment (weakly acidic) was favorable for FLQ adsorption [57]. Adsorption Isotherms Adsorption isotherms data were essential to
understand the interaction between adsorbate and adsorbent surface sites. The
equilibrium time and adsorption isotherm obtained for the FLQ –adsorbents
system are depicted in Table 3. Herein, the adsorption behavior of FLQ onto
adsorbents were evaluated by the most frequently used isotherm models.
Monolayer surface coverage and the equal availability of adsorption sites and
no transmigration of adsorbates in the plane of the surface are the idealized
assumptions associated with the Langmuir adsorption model [58]. On the other hand, the Freundlich isotherm is
an empirical model that assumes the presence of the interaction between the
molecules of the adsorbate, all surface sites are different and the multilayer
adsorption [59]. Obviously, it is apparent from the results
reported in Table 3 that the adsorption of FLQ onto adsorbents at all
temperatures is well described by the Langmuir isotherm model (R2
value of 0.95). The goodness of Langmuir model suggests monolayer adsorption of
FLQ at the outer surface of adsorbnets at all temperatures
[60]. Thus,
comparison of the R2 values for both models (Table 3) showed that
adsorption of the mixed pollution by all adsorbents were more consistent with
the Langmuir adsorption model. This indicated that monolayer adsorption of FLQ
on to the surface of adsorbents were
significant and most likely the dominant mechanism. Table 3: Freundlich and Langmuire
values for AC, Mag AC, SBS/MagAC and PC/Mag AC
Adsorption kinetics
Adsorption kinetic data is very
useful in understanding the general mechanism associated with the
adsorption of FLQ to all adsorbents. To investigate the kinetics of FLQ
adsorption process in all adsorbents, three kinetic models were used (ie, pseudo first order, pseudo second order and intraparticle
diffusion kinetic models). Graphics were placed using linear regression and
kinetic parameters are shown in Tables 4a and 4b. It was observed that FLQ adsorption to all
adsorbents followed the pseudo-second reaction with high correlation coefficients
(R2 : 0.94 - 1.00) generally and the adsorption capacity (qe, cal) from the
pseudo-second-order model was much closer to experimental experience. data
indicating that the process controlling the speed may be a chemical sorption (qe, exp) [54]. These observations were consistent with
sorption mechanisms that suggested strong electrostatic interaction or ion
exchange on the absorbent surface. In this case, the adsorption process of FLQ
can be defined as chemisorption and it is suggested that the rate determination
step is surface adsorption [61], [62]. As the initial concentration of the FLQ
solution changes from 37 mg/L to 100 mg/L, the second order rate constants (k2)
of the adsorption reaction where the adsorbed amount appears increasingly
different for each adsorbent (Tables 4a and 4b). Two dominant factors have been
proposed: (i) an increase in the initial FLQ
concentration may increase the driving force of the concentration gradient
between the adsorbent in the solution and the adsorbent in the solution, and
may result in higher FLQ intake, and (ii) higher concentrations of adsorbate
molecules were greater for the active regions [63]. In addition, the appropriateness of the
particle diffusion model was examined in the presented study. Here, R2
value was observed to be significant. Given the intraparticle diffusion model,
it can be said that there is a close agreement between the experimental and
theoretical qe value, which confirms the
applicability of this model by all adsorbents to explain phenol adsorption.
This means that surface adsorption and intraparticle diffusion occur
simultaneously. This can be attributed primarily to the outer surface of the
adsorbents in two stages, and then one by diffusion into the pores [64]. Table
4a: In FLQ adsorption of AC and Mag AC and I. II.
degree kinetic model and particle diffusion data
Table 4b: In
FLQ adsorption of SBS/MagAC and PC/MagAC
and I. II. degree
kinetic model and particle diffusion data
3.2.5. ADSORPTION THERMODYNAMICS Thermodynamic parameters of FLQ adsorption on
different temperatures and adsorbents are given in Table 5. ΔGo
values Eq. (7) and the temperature range of 291-308 K showing the spontaneous
nature of the adsorption process. The negative values of ΔGo
indicated the applicability of the sorption process and the positive values of ΔHo and ΔSo
showed that the sorption process was endothermic in nature and had a random
increase in the solid / liquid interface, respectively, during the sorption
process. The positive value of ΔS showed good affinity of the material
with dye molecules and showed an increase in the degree of freedom of the
adsorbed species [65]. Table 5: Thermodynamic parameters of FLQ adsorption on
different temperatures and adsorbents
Using the kinetic. adsorption and thermodynamic
data. it was concluded that physical adsorption in the polymer network and
chemical interaction of the OTC molecules via ion-exchange were both involved
in the adsorption process. 4.
CONCLUSION
The data obtained as part of this study show
that AC, MagAC, SBS/MagAC
and PC/MagAC can be used as adsorbents to remove FLQ
from wastewater streams. Scanning electron microscopy (SEM), X-ray diffraction
(XRD) analysis and Brunauer Emmett Teller (BET)
measurements were applied for the structural characterization of all adsorbent
materials. Adsorption isotherms, kinetic and thermodynamic calculations were
conducted to investigate the adsorption potential of all adsorbents for the
removal of FLQ from aqueous solution. The optimisation
of adsorption process was conducted by insetigating
the effects of different parameters, for example, solid-liquid ratio, solution
temperature, intial FLQ concentration and solution pH. It was observed that the adsorption capacity, acidic
and neutrality positively affected. During the adsorption process, the main
adsorption mechanisms are thought to be electrostatic interaction, ion
exchange, π-π bond interaction, aromaticity
and H-bonding and hydrophobic interaction. In conclusion, the adsorbent
materials which were prepared and used in this study can be used effectively
for the removal of FLQs from wastewater streams due to their chemical
structure, high adsorption capacity and environmental friendliness. SOURCES OF FUNDINGThis research received no specific grant from any funding agency in the public, commercial, or not-for-profit sectors. CONFLICT OF INTERESTThe author have declared that no competing interests exist. ACKNOWLEDGMENTThis study was financially supported as a project (19/081/01/1/1) by Research Project Coordination Unit. Muğla Sıtkı Koçman Üniversity. Mehmet Uğurlu declares that he has no conflict of interest. Mehmet Uğurlu has received research grants from Research Project Coordination Unit Muğla Sıtkı Koçman Üniversity. Mehmet Uğurlu is currently working as a Professor in the Department of Chemistry, Faculty of Science, Muğla Sitki Kocman University, 48000 Muğla, TURKEY. This article does not contain any studies with human or animal subjects performed by the any of the authors. The authors declare that they have no conflict of interest and they are academic staff members of Higher Education Institutions (HEI) in Turkey and UK. REFERENCES[1] Murray., A. Örmeci., B. Environ. Sci. Pollut. R (2012) vol 193, p. 820–3830. [2] Terzić, S., Senta, I., Ahel, M., Gros, M., Petrović, M., Barcelo, D., ... & Jabučar, D. (2008). Occurrence and fate of emerging wastewater contaminants in Western Balkan Region. Science of the total environment, 399(1-3, 66-77. [3] Parshikov á I., Freeman á A. J. P., Williams A. J., Moody á J. D., Sutherland J. B. Appl Microbiol Biotechnol (1999) vol. 52 p. 553-557. [4] Nieto, J., Freer, J., Contreras, D., Candal, R.J., Sileo, E.E., Mansilla, H.D., J. Hazard. Mater. (2008) vol. 15545 p. 50 [5] Nowara, A., Burhenne, J., Spiteller, M., J. Agr. Food Chem. 1997 vol. 45, p. 1459–1463. [6] Zhang, H., Huang, CH. Chemosphere (2007) vol. 66 p. 1502–1512. [7] Sotelo, J.L., Rodríguez, A., Mestanza, M., Díez, S., Álvarez, S., García, J. J. Environ. Sci. Health Part B (2012) vol. 47640 p. 652. [8] Juan L., Acero, F., Javier Benitez., Francisco J., Real, Fernando Teva, Chemical Engineering Journal 210 (2012)1-8 [9] Zhang, H., & Huang, C. H. Chemosphere, (2007). Vol. 66(8, p. 1502-1512. [10] Paul, T., Liu, J., Machesky, M. L., & Strathmann, T. J. Journal of colloid and interface science, (2014) vol. 428, p. 63-72. [11] Sotelo, J. L., Ovejero, G., Rodríguez, A., Álvarez, S., & García, J. Chemical engineering journal, (2013). Vol. 228, p. 102-113. [12] Guaita, D. P., Sayen, S., Boudesocque, S., & Guillon, E. Journal of colloid and interface science, (2011) vol. 357(2, p. 453-459. [13] Álvarez-Torrellas, S., Ribeiro, R. S., Gomes, H. T., Ovejero, G., & García, J. Chemical Engineering Journal, (2016) vol. 296, p. 277-288. [15] Adams, C., Wang, Y., Loftin, K., & Meyer, M. (2002) vol. 128(3, p. 253-260. [16] Asfaram, A., Ghaedi, M., Goudarzi, A., & Rajabi, M. Dalton Transactions, (2015) vol. 44(33, p. 14707-14723. [17] Shaker, M. A., & Yakout, A. A. Spectrochimica Acta Part A: Molecular and Biomolecular Spectroscopy, (2016). Vol. 154, p. 145-156. [18] Ghaedi, M., Khafri, H. Z., Asfaram, A., & Goudarzi, A. Spectrochimica Acta Part A: Molecular and Biomolecular Spectroscopy, (2016) vol. 152, p. 233-240.. [19] Davarnejad, R., & Panahi, P. Separation and Purification Technology, (2016) vol. 158, p. 286-292. [20] Akın, D., Yakar, A., & Gündüz, U. Synthesis of Magnetic Fe3O4‐Chitosan Nanoparticles by Ionic Gelation and Their Dye Removal Ability. Water Environment Research, (2015). vol. 87(5, p. 425-436. [21] Bagheri, A. R., Ghaedi, M., Asfaram, A., Bazrafshan, A. A., & Jannesar, R. Comparative study on ultrasonic assisted adsorption of dyes from single system onto Fe3O4 magnetite nanoparticles loaded on activated carbon: experimental design methodology. Ultrasonics sonochemistry, (2017) vol. 34, p. 294-304. [22] Bhatia, D., Datta, D., Joshi, A., Gupta, S., & Gote, Y. Adsorption study for the separation of isonicotinic acid from aqueous solution using activated carbon/Fe3O4 composites. (2018). Journal of Chemical & Engineering Data, vol. 63(2, p. 436-445. [23] Zhang, S., Wang, Z., Chen, H., Kai, C., Jiang, M., Wang, Q., & Zhou, Z. Polyethylenimine functionalized Fe3O4/steam-exploded rice straw composite as an efficient adsorbent for Cr (VI) removal. Applied Surface Science, (2018). Vol. 440, 1p. 277-1285. [24] Badi, M. Y., Azari, A., Pasalari, H., Esrafili, A., & Farzadkia, M. Modification of activated carbon with magnetic Fe3O4 nanoparticle composite for removal of ceftriaxone from aquatic solutions. Journal of Molecular Liquids, (2018). Vol. 261, p. 146-154. [25] Wen, T., Wang, J., Yu, S., Chen, Z., Hayat, T., & Wang, X. Magnetic porous carbonaceous material produced from tea waste for efficient removal of As (V, Cr (VI, humic acid, and dyes. ACS Sustainable Chemistry & Engineering, (2017). Vol. 5(5, p. 4371-4380. [26] Kim, E. A., Seyfferth, A. L., Fendorf, S., & Luthy, R. G. Immobilization of Hg (II) in water with polysulfide-rubber (PSR) polymer-coated activated carbon. water research, (2011). Vol. 45(2, p. 453-460. [27] Yang, N., Zhu, S., Zhang, D., & Xu, S. Synthesis and properties of magnetic Fe3O4-activated carbon nanocomposite particles for dye removal. Materials Letters, (2008). Vol. 62(4-5, p. 645-647. [28] Kakavandi, B., Jonidi, A., Rezaei, R., Nasseri, S., Ameri, A., & Esrafily, A. Synthesis and properties of Fe 3 O 4-activated carbon magnetic nanoparticles for removal of aniline from aqueous solution: equilibrium, kinetic and thermodynamic studies. Iranian journal of environmental health science & engineering, (2013). Vol. 10(1, p. 1-9. [29] Fard, M. A., Vosoogh, A., Barkdoll, B., & Aminzadeh, B. Using polymer coated nanoparticles for adsorption of micropollutants from water. Colloids and Surfaces A: Physicochemical and Engineering Aspects, (2017). Vol. 531, p. 189-197. [30] Lagergren, S. K. (1898). About the theory of so-called adsorption of soluble substances. Sven. Vetenskapsakad. Handingarl, 24, 1-39. [32] Acemioğlu, B. Batch kinetic study of sorption of methylene blue by perlite. Chemical Engineering Journal, (2005). Vol. 106(1, p. 73-81. [33] Uğurlu, M. Kinetic of the adsorption of reactive dyes by using sepiolite mineral. Microporous and Mesoporous Materials, (2009). Vol. 119, p. 276-283. [34] Ugurlu, M. Adsorption studies and removal of nitrate from bleached kraft mill effluent by fly-ash and sepiolite. Fresenius Environmental Bulletin, (2009). Vol. 18(12, p. 2328-2335. [35] Girgis, B. S., Temerk, Y. M., Gadelrab, M. M., & Abdullah, I. D. X-ray diffraction patterns of activated carbons prepared under various conditions. Carbon letters, (2007). Vol. 8(2, p. 95-100. [36] Vaizoğullar, A. İ. TiO2/ZnO supported on sepiolite: preparation, structural characterization, and photocatalytic degradation of flumequine antibiotic in aqueous solution. Chemical Engineering Communications, (2017). Vol. 204(6, p. 689-697. [37] Oliveira, G. F. D., Andrade, R. C. D., Trindade, M. A. G., Andrade, H. M. C., & Carvalho, C. T. D. Thermogravimetric and spectroscopic study (TG–DTA/FT–IR) of activated carbon from the renewable biomass source babassu. Química Nova, (2017). Vol .40(3, p. 284-292. [38] Uğurlu, M., Gürses, A., & Açıkyıldız, M. Comparison of textile dyeing effluent adsorption on commercial activated carbon and activated carbon prepared from olive stone by ZnCl2 activation. Microporous and Mesoporous Materials, (2008). Vol. 111(1-3, p. 228-235. [40] Mazloomi, F., & Jalali, M. Ammonium removal from aqueous solutions by natural Iranian zeolite in the presence of organic acids, cations and anions. Journal of Environmental Chemical Engineering, (2016). Vol. 4(1, p. 240-249. [41] Wang, Y., Lu, J., Wu, J., Liu, Q., Zhang, H., & Jin, S. Adsorptive removal of fluoroquinolone antibiotics using bamboo biochar. Sustainability, (2015). Vol. 7(9, p. 12947-12957. [42] Duan, W., Li, M., Xiao, W., Wang, N., Niu, B., Zhou, L., & Zheng, Y. Enhanced adsorption of three fluoroquinolone antibiotics using polypyrrole functionalized Calotropis gigantea fiber. Colloids and Surfaces A: Physicochemical and Engineering Aspects, (2019). Vol. 574, p. 178-187. [43] Gu, C., & Karthikeyan, K. G. Sorption of the antimicrobial ciprofloxacin to aluminum and iron hydrous oxides. Environmental science & technology, (2005). Vol. 39(23, p. 9166-9173. [47] Wang, B., Jiang, Y. S., Li, F. Y., & Yang, D. Y. Preparation of biochar by simultaneous carbonization, magnetization and activation for norfloxacin removal in water. Bioresource technology, (2017). Vol. 233, p. 159-165. [48] Yang, W., Lu, Y., Zheng, F., Xue, X., Li, N., & Liu, D. Adsorption behavior and mechanisms of norfloxacin onto porous resins and carbon nanotube. Chemical engineering journal, (2012). Vol. 179, p. 112-118. [49] Fu, H., Li, X., Wang, J., Lin, P., Chen, C., Zhang, X., & Suffet, I. M. Activated carbon adsorption of quinolone antibiotics in water: Performance, mechanism, and modeling. Journal of environmental sciences, (2017). Vol. 56, p. 145-152. [51] Gao, Y., Li, Y., Zhang, L., Huang, H., Hu, J., Shah, S. M., & Su, X. Adsorption and removal of tetracycline antibiotics from aqueous solution by graphene oxide. Journal of colloid and interface science, (2012). Vol. 368(1, p. 540-546. [53] Barbosa, J., Barron, D., Cano, J., Jimenez-Lozano, E., Sanz-Nebot, V., & Toro, I. Evaluation of electrophoretic method versus chromatographic, potentiometric and absorptiometric methodologies for determing pKa values of quinolones in hydroorganic mixtures. Journal of pharmaceutical and biomedical analysis, (2001). Vol. 24(5-6, p. 1087-1098. [55] Conkle, J. L., Lattao, C., White, J. R., & Cook, R. L. Competitive sorption and desorption behavior for three fluoroquinolone antibiotics in a wastewater treatment wetland soil. Chemosphere, (2010). Vol. 80(11, p. 1353-1359. [57] Sotelo, J. L., Ovejero, G., Rodríguez, A., Álvarez, S., & García, J. Analysis and modeling of fixed bed column operations on flumequine removal onto activated carbon: pH influence and desorption studies. Chemical engineering journal, (2013). Vol. 228, p. 102-113. [58] Ötker, H. M., & Akmehmet-Balcıoğlu, I. Adsorption and degradation of enrofloxacin, a veterinary antibiotic on natural zeolite. Journal of Hazardous Materials, (2005). 122(3, 251-258. [59] He, X., Wang, B., & Zhang, Q. Phenols removal from water by precursor preparation for MgAl layered double hydroxide: Isotherm, kinetic and mechanism. Materials Chemistry and Physics, (2019). 221, 108-117. [60] Ahmed, M. J., & Theydan, S. K. Fluoroquinolones antibiotics adsorption onto microporous activated carbon from lignocellulosic biomass by microwave pyrolysis. Journal of the Taiwan Institute of Chemical Engineers, (2014). Vol. 45(1, p. 219-226. [61] Leal, R. M. P., Alleoni, L. R. F., Tornisielo, V. L., & Regitano, J. B. Sorption of fluoroquinolones and sulfonamides in 13 Brazilian soils. Chemosphere, (2013). Vol. 92(8, p. 979-985. [63] Renault, F., Morin-Crini, N., Gimbert, F., Badot, P. M., & Crini, G. Cationized starch-based material as a new ion-exchanger adsorbent for the removal of CI Acid Blue 25 from aqueous solutions. Bioresource technology, (2008). Vol. 99(16, p. 7573-7586. [64] Duan, W., Li, M., Xiao, W., Wang, N., Niu, B., Zhou, L., & Zheng, Y. Enhanced adsorption of three fluoroquinolone antibiotics using polypyrrole functionalized Calotropis gigantea fiber. Colloids and Surfaces A: Physicochemical and Engineering Aspects, (2019). Vol. 574, p. 178-187. [65] El Yacoubi, A., Rahali, N., Elmerras, D., Rezzouk, A., & El Idrissi, B. C. Removal of Methylene Blue Dye by Adsorption on Natural Sand. American Journal of Environment and Sustainable Development, (2019). Vol. 4(2, p. 84-88.
This work is licensed under a: Creative Commons Attribution 4.0 International License © IJOEST 2016-2020. All Rights Reserved. |