Adsorption, photodegradation and antibacterial study of graphene-Fe 3O4 nanocomposite for multipurpose water purification application

Article abstract of DOI:10.1039/c4ra02913e

Graphene-Fe₃O₄ (G-Fe₃O₄) composite was prepared from graphene oxide (GO) and FeCl₃· 6H₂O by a one-step solvothermal route. The as-prepared composite was characterized by field-emission scanning electron microscopy, transmission electron microscopy, dynamic light scattering and X-ray powder diffraction. SEM analysis shows the presence of Fe₃O₄ spheres with size ranging between 200 and 250 nm, which are distributed and firmly anchored onto the wrinkled graphene layers with a high density. The resulting G-Fe ₃O₄ composite shows extraordinary adsorption capacity and fast adsorption rates for the removal of Pb metal ions and organic dyes from aqueous solution. The adsorption isotherm and thermodynamics were investigated in detail, and the results show that the adsorption data was best fitted with the Langmuir adsorption isotherm model. From the thermodynamics investigation, it was found that the adsorption process is spontaneous and endothermic in nature. Thus, the as-prepared composite can be effectively utilized for the removal of various heavy metal ions and organic dyes. Simultaneously, the photodegradation of methylene blue was studied, and the recycling degradation capacity of dye by G-Fe₃O₄ was analyzed up to 5 cycles, which remained consistent up to ～97% degradation of the methylene blue dye. Although iron oxide has an affinity towards bacterial cells, its composite with graphene still show antibacterial property. Almost 99.56% cells were viable when treated with Fe₃O₄ nanoparticle, whereas with the composite barely 3% cells survived. Later, the release of ROS was also investigated by membrane and oxidative stress assay. Total protein degradation was analyzed to confirm the effect of the G-Fe₃O₄ composite on E. coli cells.

Full text of DOI:10.1039/c4ra02913e

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PAPER
Adsorption, photodegradation and antibacterial
study of graphene–Fe3O4 nanocomposite for
multipurpose water purification application†
Cite this: RSC Adv., 2014, 4, 28300
a
b
Chella Santhosh, Pratap Kollu,‡ Sejal Doshi,c Madhulika Sharma,c
*
Dhirendra Bahadur, Mudaliar T. Vanchinathan,a P. Saravanan,d Byeong-Su Kime
and Andrews Nirmala Grace
c
*
a
*
Graphene–Fe3O4 (G–Fe3O4) composite was prepared from graphene oxide (GO) and FeCl3$6H2O by a
one-step solvothermal route. The as-prepared composite was characterized by field-emission scanning
electron microscopy, transmission electron microscopy, dynamic light scattering and X-ray powder
diffraction. SEM analysis shows the presence of Fe3O4 spheres with size ranging between 200 and 250
nm, which are distributed and firmly anchored onto the wrinkled graphene layers with a high density.
The resulting G–Fe3O4 composite shows extraordinary adsorption capacity and fast adsorption rates for
the removal of Pb metal ions and organic dyes from aqueous solution. The adsorption isotherm and
thermodynamics were investigated in detail, and the results show that the adsorption data was best fitted
with the Langmuir adsorption isotherm model. From the thermodynamics investigation, it was found that
the adsorption process is spontaneous and endothermic in nature. Thus, the as-prepared composite can
be effectively utilized for the removal of various heavy metal ions and organic dyes. Simultaneously, the
photodegradation of methylene blue was studied, and the recycling degradation capacity of dye by G–
Fe3O4 was analyzed up to 5 cycles, which remained consistent up to ꢀ97% degradation of the
methylene blue dye. Although iron oxide has an affinity towards bacterial cells, its composite with
graphene still show antibacterial property. Almost 99.56% cells were viable when treated with Fe3O4
nanoparticle, whereas with the composite barely 3% cells survived. Later, the release of ROS was also
investigated by membrane and oxidative stress assay. Total protein degradation was analyzed to confirm
the effect of the G–Fe3O4 composite on E. coli cells.
Received 2nd April 2014
Accepted 3rd June 2014
DOI: 10.1039/c4ra02913e
www.rsc.org/advances
reported to immobilize the enzymes in some plants and can
also cause energy deciency in the transmission of nerve
impulses in humans.1,2 Hence, it has become immediately
necessary to prevent water pollution caused by microbes as well
as by heavy metals and dyes. Most purication methods suffer
from some drawbacks such as high capital and operational
costs or disposal of residual metal sludge, and hence are not
suitable for small scale industries.3 In addition, these methods
are not very efficient at low concentrations in the range 1–100
Introduction
Water pollution resulting from biotic and abiotic pollutants has
been a major environmental threat attributable to both existing
and emerging industries. Heavy metals and dyes beyond the
permissible limits directly cause toxicity to humans and other
living organisms. Metal ions, such as Cd(II) and Pb(II) have been
aCentre for Nanotechnology Research, VIT University, Vellore 632014, India. E-mail:
anirmalagladys@gmail.com; anirmalagrace@vit.ac.in; Tel: +919791322311
bDST-INSPIRE Faculty, Department of Metallurgical Engineering and Materials
Science, Indian Institute of Technology Bombay, Mumbai 400076, India
cDepartment of Metallurgical Engineering and Materials Science, Indian Institute of
Technology Bombay, Mumbai 400076, India. E-mail: dhirenb@iitb.ac.in
dDefence Metallurgical Research Laboratory, Hyderabad 500 058, India
mg Lꢁ1 4,5 Silver has become the basis for antibacterial activity;
.
however, because silver is expensive and leaches into the envi-
ronment, the focus has shied to the exploration of inexpensive
materials for antibacterial activity.
Adsorption is one of the promising processes for the removal
of heavy metal ions from water. Some of the adsorbent mate-
rials for heavy metal ion removal are activated carbon (powder
or granular) and CNTs, etc.,6–8 for the removal of Pb, cadmium
and other heavy metal ions. However, the cost of adsorbent
becomes relatively high when pure sorbents are used. Photo-
degradation of dye has been a successful method for the
eDepartment of Chemistry and Department of Energy Engineering, Ulsan National
Institute of Science and Technology (UNIST), UNIST-gil 50, Ulsan 689-798, Korea
† Electronic supplementary information (ESI) available. See DOI:
10.1039/c4ra02913e
‡ Current address: Thin Film Magnetism group, Cavendish Laboratory,
Department of Physics, University of Cambridge, Cambridge CB3 0HE, UK.
Email: pk419@cam.ac.uk
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eradication of dyes from effluent water, which further helps in
water purication.
Preparation of graphene oxide
Graphene oxide (GO) was prepared from natural graphite
powder by the modied Hummers method.23 In a typical
synthesis, 1 g of graphite was added to 23 mL of 98% H2SO4,
followed by stirring at room temperature over a period of 24 h.
Then, 100 mg of NaNO3 was added to the mixture, stirred for 30
min and then maintained below 5 ꢂC in an ice bath. Finally, 3 g
of KMnO4 was added slowly into the mixture. Aerwards, the
mixture was heated to 35–40 ꢂC with continuous stirring for 30
min. Then, 46 mL of water was added to the abovementioned
mixture and mixed well for 25 min. Finally, 140 mL of water and
10 mL of 30% H2O2 were added to the mixture to stop the
reaction. Unreacted graphite in the resulting mixture was
removed by centrifugation and the resulting mixture was dried
at 60 ꢂC in a vacuum oven. The as-synthesized GO was dispersed
into individual sheets in distilled water at a concentration of 0.5
mg mLꢁ1 with the help of ultrasound for further use.
Graphene and Fe3O4 (ref. 9 and 10) individually have been
reported for their ltration and purication activity. For
standalone use they can be used as water-purifying materials,
thus it was considered to combine them as a composite and
efficiently use them as a nanomaterial for water purication.
Graphene is a single layer of carbon densely packed in a
honeycomb crystal lattice structure. In comparison with other
carbonaceous material, it has drawn much attention since its
discovery due to its unique electronic and mechanical proper-
ties.11,12 Its remarkable properties, such as large surface area
(2630 m2 gꢁ1), good chemical stability, at structure, imply that
it can be used as an excellent adsorbent.13–16 Its basal plane
structure allows for strong p–p interactions, assigned to the p–
p stacking between aromatic dyes and the p-conjugation
regions of the graphene layers.17This property of graphene
would serve as a good adsorbent for removing pollutants such
as metal ions, dyes, and biological materials. However, gra-
phene sheets usually suffer from serious agglomeration and
restacking due to their p–p interactions between neighbouring
sheets, leading to a signicant loss of effective surface area and
lower adsorption capacity.18,19 To overcome such difficulties,
magnetic adsorbents have emerged as new generation materials
for environmental application. The advantage of such materials
is their ability to effect magnetic separation by applying an
external magnetic eld to extract the adsorbent material from
the suspension.20,21
Synthesis of G-Fe3O4nanocomposites
The as-prepared GO was used as a precursor for the preparation
of G–Fe3O4 composites by a one-step solvothermal method. In a
typical synthesis, the as-prepared GO (0.5 g) was exfoliated by
ultrasonication in 80 mL of ethylene glycol for more than 3 h.
1.6 g FeCl3$6H2O and 3.2 g NaAc were then dissolved in GO/EG
solution at ambient temperature. Aer stirring for about 30
min, the solution was transferred to a 100 mL teon-lined
ꢂ
stainless-steel autoclave and maintained at 200 C for 6 h, fol-
lowed by naturally cooling to ambient temperature. The black
precipitate was centrifuged, washed with ethanol several times
and nally dried at 60 C in a vacuum oven.
Compared with traditional methods, such as ltration,
centrifugation and gravitational separation, magnetic separa-
tion requires less energy and better separation can be achieved.
Therefore, the synthesis of graphene with covalently attached
Fe3O4 offers an effective approach to overcome the separation
problem associated with graphene. At the same time, the
attachment of Fe3O4 with graphene provides a decrease in the
possibility of serious agglomeration. In this regard, the current
work is focused on the synthesis of G–Fe3O4 composites by a
solvothermal route. Further, the as-prepared composites were
tested for adsorption properties towards Pb ions and the pho-
todegradation of methylene blue. To better understand its
antimicrobial mechanism, we compared the antibacterial
activity of graphene and G–Fe3O4 towards a bacterial model,
Escherichia coli. Under similar concentration and incubation
conditions, G–Fe3O4 dispersion shows the highest antibacterial
activity. ROS22 and protein degradation were responsible for
bacterial cell wall rupture.
ꢂ
Characterization
The particle size and morphology of the material was studied by
FE-Scanning Electron Microscopy (FE-SEM) (HITACHI SU6600
SEM). The phase and crystallographic structure were identied
by X-ray diffraction (XRD; Philips X0Pert Pro, Cu-Ka: l ¼
0.1540598 nm). The adsorbed metal ion concentration was
examined by an atomic adsorption spectrophotometer (Perkin-
Elmer Analyst 240). Thermogravimetric analysis (TGA) of the
samples was performed on an SDT Q600 (TA Instruments,
Korea) with a heating rate of 5 ꢂC minꢁ1 from 0ꢂ to 1000 ꢂC. The
G–Fe3O4 samples were characterized with a high resolution
transmission electron microscope (HR-TEM) JEOL-2000EX
operated at 120 kV. Dynamic light scattering measurements
were performed using a Beckman Coulter Delsa™ Nano
instrument.
Experimental
Analytical measurement
Materials
For analytical studies, analytical grade lead nitrate was used to
All chemicals were analytical grade, purchased from Sigma prepare a stock solution containing 1000 mg Lꢁ1 of Pb(NO3)2,
Aldrich and used as received without purication. Graphite which was further diluted with double-distilled deionized water
powder, hydrogen peroxide (30 wt%), sodium nitrate (98%), to the required ionic concentration. Adsorption thermodynamic
sulphuric acid (98 wt%), potassium permanganate, Pb(NO3)2, experiments were conducted in a 100 mL glass ask containing
FeCl3$6H2O and sodium acetate (NaAc) were purchased from 25 mg of adsorbent and 100 mL of Pb ion solution at varying
SD-Fine and used as received.
concentrations (10–50 mg Lꢁ1) and pH 5. The samples were
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then placed in an orbital shaker for continuous stirring at 250 towards a living system. The cells were seeded into 96-well
rpm for different durations and at selected temperatures (300, plates at a density of 1 ꢃ 104 cells per well and incubated for 24
310, and 320 K). Aer a certain time, the suspension was ltered h in a 5% CO2 environment. Then, 200 mL of different concen-
using 0.22 mm cellulose nitrate membrane, and the ltrates trations of the dispersed suspension of G–Fe3O4 (2.0, 1.0, 0.5,
were immediately examined using atomic adsorption spectro- 0.25, 0.125, 0.625, and 0.3125 mg mLꢁ1) in DMEM growth
photometry (AAS) to measure the ion concentration. The medium were added to the cells and incubated for another 24 h
difference between the initial and the equilibrium ion concen- at 37 ꢂC and 5% CO2 environment. Thereaer, the cells were
tration gave the amount of ion adsorbed to the G–Fe3O4 gently washed with phosphate buffer saline (PBS; pH 7.3) and
surfaces.
processed for SRB assay to determine the viable cell population.
Non-treated cells were used as control for the experiments. The
cells were then xed with 10% trichloroacetic acid solution by
incubating at 4 ꢂC for 1 h. This was followed by gentle washing
of cells with water and then staining with 0.4% SRB dissolved in
1% acetic acid and then incubating in the dark. The cell-bound
dye was then extracted with 200 mL of 10 mM Tris buffer solu-
tion (pH 10.5), and its optical density was measured using a
multiwall plate reader at 560 nm. The viable cell population was
calculated using the following formula:
Batch mode adsorption
The effects of experimental parameters, such as initial
concentration (10–50 mg Lꢁ1), pH (3–8), and temperature, were
studied in a batch mode of adsorption for specic contact times
(0–180 min). To determine the effect of each parameter, the
other parameters were maintained constant. Pb solution was
prepared by dissolving Pb(NO3)2 in double distilled water and
used as the stock solution. Then, the solutions were diluted to
the required concentration for the experimental study. The pH
of the solution was adjusted using 0.1 M HCl or 0.1 M NaOH.
For contact time measurements, 100 mL of Pb ion solution
containing 10 mg Lꢁ1 of an initial ion concentration at pH 5 was
put into a 250 mL conical ask with a xed amount of adsorbent
(25 mg Lꢁ1) and agitated in an orbital shaker at 250 rpm and 310
K. At various intervals, the adsorbent were separated from the
samples by ltering, and the ltrates were analyzed by AAS to
determine the concentration of each ion in the solution.24 The
adsorption percentage of a metal ion was calculated as follows:
Absorbance of treated cells
% cell viablity ¼
ꢃ 100
Absorbance of control cells
Antibacterial activity of G-Fe3O4
E. coli were procured from K.C. College, Mumbai, India. The
cultures were initially isolated from patients infected with
water-borne disease. This culture has been sub-cultured for the
last 20 years and hence is assumed to have very low virulence.
Nutrient broth media was procured from HiMedia. All the
chemicals were purchased from Sigma-Aldrich and used
without further purication.
Ci ꢁ Cf
Adsorption ð%Þ ¼
ꢃ 100
(1)
Ci
where Ci and Cf are the initial and nal metal ion concentration
(aer contact with adsorbents), respectively.
Results and discussion
Structural and morphological analysis
Photodegradation activity of G-Fe3O4
The phase composition and structures of G–Fe3O4 nano-
composites were examined using X-ray powder diffraction and
the corresponding pattern is shown in Fig. S1.† In the pattern of
GO (Fig. S1a†), a diffraction peak at 11.08ꢂ is observed, which
corresponds to the (001) plane of graphene oxide structure. In
Fig. S1b,† natural graphite shows diffraction peaks at 26.4ꢂ,
44.3ꢂ and 54.5ꢂ corresponding to the hexagonal lattice of the
(002), (101) and (004) planes, respectively. The major peak is at
The photodegradation activity of the as-prepared G–Fe3O4 was
evaluated by the photodegradation of methylene blue dye using
60 W of a visible light CFL lamp. For the degradation of dye, 0.2
g Lꢁ1 of photocatalyst was added to 100 mL of methylene blue
dye (10 mg Lꢁ1) aqueous solution. 1 mL of 30% H2O2 was added
to the reaction mixture at the beginning of light irradiation.
About 1 mL of aliquot was withdrawn at a given irradiation time
and then magnetically separated to remove the catalyst. In the
case of bare graphene nanosheets, the suspension was sepa-
rated by centrifuging at 3000 rpm. The concentration of the
remaining dye was determined by measuring the absorbance of
solutions at 664 nm by UV-vis spectroscopy.
26.4ꢂ, which corresponds to a basal spacing of d002 ¼ 3.38 A. The
˚
typical process is the reduction of GO to graphene during the
solvothermal process. Aer reduction, the peak at 11ꢂ
completely disappeared and a weak hump (*) could be observed
in Fig. S1(c).† In Fig. S1c,† G–Fe3O4 composite with diffraction
peaks at 2q values of 30.6ꢂ, 36.1ꢂ, 43.7ꢂ, 57.6ꢂ, 63.1ꢂ and 74.7ꢂ
correspond to (220), (222), (400), (511), (440) and (533) planes of
Biocompatibility of G-Fe3O4
Sulforhodamine-B (SRB) assay was performed to evaluate the Fe3O4, respectively. This matches with the standard JCPDS no.
biocompatibility of G–Fe3O4 with normal mouse broblast cells 65-3107. The results indicate the existence of Fe3O4 in the
(L929). Although iron oxide nanoparticles have been well- composite material and clearly show the reduction of GO peak
established as biocompatible nanomaterials, it is essential to in the composite material. The disappearance of the peak at 2q
evaluate the toxic effect of G–Fe3O4 on cell proliferation and ¼ 11ꢂ clearly indicated that GO was reduced to graphene. The
morphology of L929 cells because of the presence of graphene. absence of a sharp peak at 26ꢂ indicates that it is not graphite
It is important to test the toxic effect of functional material but should be graphene. The peak at 2q ¼ 26ꢂ pertaining to
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graphene is not as intense because it is in the form of a weight loss was observed due to adsorbed water and loss of
composite and the Fe3O4 phase dominates the graphene layers. moisture in the as-prepared material. A major weight loss was
The graphene peak was not predominant because the stacking due to graphene and Fe3O4 composite. Apparently, in DTA
of graphene sheets in G–Fe3O4 was disordered and exfoliated to curve, two main exothermic peaks were obtained at 290 ꢂC and
a large extent.25,26
450 ꢂC, which relate to the thermal decomposition of ferrites
along with carbon.
The N2 adsorption–desorption isotherm was used to deter-
mine the porous capacity of graphene–Fe3O4 (Fig. S4a†).
According to IUPAC (International Union of Pure and Applied
Chemistry) classication, the isotherms show typical type IV
and type H3 hysteresis loop. This behaviour indicates the
predominance of mesopores. The mesoporous nature is also
evident from the sharp increment in adsorption volume above
the relative pressure of 0.8. The specic surface area of the
composite from BET analysis was found to be 27.38 m2 gꢁ1. The
Barrett–Joiner–Halenda (BJH) method was used to calculate the
pore size on the desorption branch of the N2 adsorption–
desorption isotherm. The pore volume versus pore diameter
curve (Fig. S4b†) shows a sharp peak at 3.7 nm along with a
broad distribution of mesopores. This behaviour also conrms
the predominance of mesopores in G–Fe3O4.28
Morphological characterization of G-Fe3O4
Fig. 1 shows the surface morphology of G–Fe3O4 at different
magnications. The TEM images of G–Fe3O4 reveal that the
product consists of a large quantity of Fe3O4 spheres with sizes
ranging from 200 to 250 nm. Aer combination with graphene
to form a G–Fe3O4 composite, the Fe3O4 spheres are decorated
and rmly anchored on the wrinkled graphene layers with a
high density, as shown in Fig. 1. Note that the pleated structure
of graphene may favour the hindrance of the Fe3O4 spheres
from agglomeration and enable their good distribution on
graphene, whereas the Fe3O4 spheres serve as a stabilizer to
separate graphene sheets against aggregation. In addition, the
Fe3O4 spheres are observed to be porous in nature, which will
further help in the adsorption process.
Further, dynamic light scattering measurements (DLS) were
recorded, and the results showed that the nanoparticles are
monodispersed (Fig. S2a†). The hydrodynamic sizes of the
particles were found to be around 500 nm. In general, the
hydrodynamic diameter takes into consideration factors such
as solvation and other effects, and hence will be generally
higher than that measured by other methods.27 The sizes of the
particles were further conrmed by SEM analysis, which
showed the dispersion of Fe3O4 particles on graphene sheets
(Fig. S2b†).
Adsorption parameters
To further explore the adsorption property of the prepared
materials, a batch adsorption experiment was conducted; the
corresponding adsorption parameters of lead ions onto G–
Fe3O4 nanocomposites are shown in Fig. S5.† Fig. S5a† shows
the removal efficiency of lead ions onto G–Fe3O4 surfaces, as a
function of contact time. It was noted that the adsorption of Pb
increased quickly with an increase in contact time, and at a
certain period it reached an equilibrium state. It was seen that
adsorption is rapid due to the availability of numerous active
sites on the adsorbent surface at the initial stage. This fast
adsorption may be due to the special one-atom-thick layered
structure of GO. In addition, because it contacts Pb ions in the
aqueous solution, adsorption occurs immediately due to higher
driving force to enable the faster transfer of Pb ions to the active
sites on the surface of G–Fe3O4. With further increase in time,
the diminishing availability of the remaining active sites and
decrease in the driving force makes the adsorption process
slow, and thus equilibrium is achieved much later. Thus, the
adsorption rate decreases. The equilibrium time increases with
increasing initial Pb ion concentration. It can be concluded
from Fig. S5a† that it took about 100 min to reach adsorption
equilibrium. From 120–180 min, the concentration of the ions
remained unchanged.
TGA–DTA/BET analysis
Fig. S3† shows the TGA curve of the as-prepared G–Fe3O4
composite with a minor weight loss from 50 ꢂC to 100 ꢂC
ꢂ
ꢂ
ꢂ
(ꢀ10%) and major weight loss from 270 C to 450 C (ꢀ40%).
Aer this, no weight loss was obtained up to 1000 C. A minor
Fig. S5b† shows the amount of adsorbed ions as a function of
the initial concentration from aqueous solution. The initial
concentration provides an important driving force to overcome
the mass transfer of Pb ions between the aqueous and solid
phases; hence, a higher initial concentration of Pb ions may
increase adsorption capacity. In this experiment, the following
concentrations (10, 20, 30, 40, 50 and 60 mg Lꢁ1) were chosen at
pH 5, equilibrium time of 120 min and T ¼ 310 K. With an
increase in ion concentration, the percentage of ion adsorption
increased. About 99% of the Pb ions were adsorbed onto G–
Fig. 1 TEM images of G–Fe3O4 nanocomposites.
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Fe3O4 surface aer 120 min for an initial Pb concentration of 50
The Langmuir and Freundlich adsorption isotherm models
mg Lꢁ1. Fig. S5c† shows the variation of adsorption capacity of can be expressed as follows:
Pb ions by G–Fe3O4 at various pH values. The adsorption
capacity increases with an initial pH of 3 and reaches a
Ce
1
1
¼
þ
Ce
(3)
(4)
qe Kdqm qm
maximum at pH 7. The effect of pH on the adsorption
percentage of Pb ions on the adsorbent G–Fe3O4 were studied at
varying pH values over the range of 3–8 using the same
concentration of ions; the results are displayed in Fig. S5c.†
Although a maximum uptake was noted at pH 8, as the pH
increased more than 7, the metal ion started to precipitate.
Therefore, no experiment was conducted at pH > 7. The increase
in adsorption capacity at pH > 7 could be attributed to both the
adsorption of ions onto the surface of adsorbent and precipi-
tation. Therefore, the optimal pH was observed to be 5 for the
adsorption of Pb ions. By further increasing the pH adsorption
decreased, which was probably due to the formation of Pb
hydroxides and chemical precipitation.
Fig. S5d† shows the adsorption% of Pb ions onto G–Fe3O4
matrix as a function of temperature. The percentage-adsorption
experiments were conducted at 300, 310 and 320 K to investigate
the effect of temperature with an initial concentration of Pb
ions of 20 mg Lꢁ1, adsorbents dosage of 25 mg Lꢁ1 and pH 5.
When temperature was increased from 300 to 320 K, the
percentage of adsorption of Pb ions increased from 85.2% to
87.2%. An increase in the amount of equilibrium adsorption of
Pb ions with an increase in temperature may be explained by the
fact that the adsorbent sites were more active at higher
temperatures. This condition shows that adsorption occurs
more as a physical rather than a chemical process.
1
ln qe ¼ ln KF þ ln Ce
n
where Ce (mg Lꢁ1) is the equilibrium concentration of ion in the
solution, qe and qm (mg gꢁ1) are the maximum adsorption
capacity at equilibrium, and Kd (L mgꢁ1) is the effective disso-
ciation constant that relates to the affinity binding site, KF and n
are physical constants representing the adsorption capacity and
intensity of adsorption, respectively.
The slope and intercept of the linear Freundlich equation are
1/n and ln KF, respectively. Fig. 2 shows the adsorption isotherm
of Pb ions onto G–Fe3O4 surfaces at 300 K, 310 K and 320 K.
From the graph, it can be seen that the adsorption capacity
increased with an increase in temperature, indicating an
endothermic reaction. We observed that the adsorption
percentage increased along with an increase in Ce.
Two models, the Langmuir and Freundlich isotherm
models, have been adopted for Pb ion adsorption onto G–Fe3O4
at different temperatures and at different pH values in the
experiment. The relative parameters were calculated from the
equation and are listed in Table 1. It can be seen from the R2
values that the adsorption isotherms can be simulated well by
the two models. The Table suggests that the adsorption of Pb on
G–Fe3O4 is mainly a monolayer type.
It can be inferred from the R2 values given in Table 1 that the
Langmuir isotherm ts the experimental data better than the
Freundlich isotherm. It was also noted that adsorption is a
Adsorption isotherm study
Developing an appropriate isotherm model for adsorption is
essential to design and optimize an adsorption process. Several
isotherm models have been developed for evaluating the equi-
librium adsorption of compounds from solutions such as
Langmuir,29,30 Fruendlich and Dubinin-Radushkevich. The
experimental results of this study were tted with two models.
The equilibrium adsorption isotherms are important for
determining the adsorption capacity of Pb ions and diagnose
the nature of adsorption onto G–Fe3O4 surface.31
The equilibrium concentration adsorption capacity of
adsorbent was calculated as:
Fig. 2 Percentage of adsorption isotherm of Pb ions onto G–Fe3O4
surfaces: (a) Langmuir isotherm; (b) Freundlich isotherm. Initial
concentration of the ion was 10, 20, 30, 40, 50, and 60 mg Lꢁ1; pH 5;
vðC0 ꢁ CeÞ
qe ¼
(2)
w
contact time, 120 min and adsorbent dosage 25 mg Lꢁ1
.
where qe is the equilibrium adsorption capacity of adsorbent in
mg metal per g adsorbent, C0 is the initial concentration of the
metal ions in mg Lꢁ1, Ce is the equilibrium concentration of the
metal ions in mg Lꢁ1, v is the volume of metal ion solution in L,
and w is the weight of adsorbent in g.
Table 1 Isotherm models for the adsorption of Pb ions onto G–Fe3O4
Langmuir Freundlich
The adsorption data can then be correlated with Langmuir
and Freundlich isotherm model equations. In this study, two
classical adsorption models were employed to describe the ion
adsorption equilibrium. The Langmuir isotherm is valid for
monolayer adsorption onto a surface with a nite number of
identical sites.
Temp (K)
Kd
qm (mg gꢁ1
)
R2
KF
n
R2
300
310
320
0.12
0.1
0.1
69
69
67
0.998
0.992
0.9818
4.2
26.96
10.86
0.23
0.24
0.27
0.9328
0.9248
0.9686
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monolayer process and that the adsorption of all species G–Fe3O4-based adsorbent has the maximum Pb ion adsorption
requires equal activation energy. From the table it is clear that at capability.
different temperatures there is change in qm and Kd values; the
maximum adsorption capacity qm obtained was 69 mg gꢁ1
.
Photo-Fenton degradation of dye
Further work was carried out for probing the role of this
material towards dye degradation applications. Dye adsorption
on the catalyst surface is a prerequisite for efficient dye removal.
The adsorption of methylene blue dye was studied within a 30
min time interval using UV-vis spectroscopy. Fig. S6† shows that
the adsorption of dye is quite rapid in the rst 30 min, and then
gradually rises with increase in adsorption time. Aer 180 min,
the amount of dye removed from the suspension is calculated
using the following equation:
Thermodynamic parameters
Thermodynamic parameters provide additional in-depth infor-
mation regarding the inherent energetic changes involved
during adsorption. To assess thermodynamic parameters, the
adsorption isotherm of Pb ions onto G–Fe3O4 surfaces were
measured at 300, 310 and 320 K, and the changes in thermo-
dynamic parameters of standard Gibbs free energy of adsorp-
tion (DGꢂ), standard enthalpy (DHꢂ) and standard entropy (DSꢂ)
were calculated from the variation of thermodynamic equilib-
rium constant, Ko, with a change in temperature.
1 ꢁ Ct
Removal ð%Þ ¼
ꢃ 100
(7)
C0
where C0 (mg Lꢁ1) is the initial dye concentration before
removal and Ct (mg Lꢁ1) is the concentration of dye remaining
in the solution aer stirring for 180 min in the dark. The
removal efficiency of the composite at different concentrations
(0.1 to 0.3 g Lꢁ1) and graphene is 49.6, 79.6, 98.5 and 99%.
Graphene shows much better removal efficiency compared
with the composite of the same amount (0.2 g Lꢁ1), which can
be attributed to the large surface area of graphene sheets. In
addition, p–p stacking between the dyes and p-conjugated
regions of the sheets can offer more active adsorption sites and
catalytic reaction centers.36 The photocatalytic property of the
composite is measured by the degradation of the dye in the
presence of visible light (Fig. 3). Before light irradiation, the
suspension of the catalyst and the dye was stirred for 180 min in
the dark to reach the adsorption–desorption equilibrium
between the catalyst and the dye. It has been observed from
Fig. 3(a and c) that upon the addition of H2O2 in the presence of
visible light, the degradation rate increases rapidly. For
comparison, the effect of H2O2 in the dark on G–Fe3O4
composite is also shown in a graph (Fig. 3(b)); an enlarged view
of dye degradation is also shown as an inset in Fig. 3. The
enhancement in dye degradation can be attributed to the
synergistic interaction between GO and Fe3O4 nanoparticles.
The electron transfer between NPs and GO sheet facilitates the
The standard Gibbs free energy of adsorption, DGꢂ, is
DGꢂ ¼ ꢁRTln Ko
where R ¼ gas constant, T ¼ absolute temperature
RTln Ko ¼ TDSꢂ ꢁ DHꢂ
(5)
(6)
From Table 2, it is clear that the value of change in standard
enthalpy is positive and Gibbs free energy is negative, showing
that the adsorption of Pb ions onto G–Fe3O4 adsorbents is
endothermic. Consequently, it can be seen that there is a
change in the adsorption as temperature increases, and here it
is seen that the adsorption is same at 300 and 310 K.
A positive standard enthalpy change suggests that the
interaction of Pb ions with G–Fe3O4 is endothermic, which is
supported by the increasing adsorption of Pb with increase in
temperature. A comparison has been made of the adsorption
capacity for various adsorbents cited in the literature for the
removal of Pb ions (Table 3). It is clear from the table that our
Table 2 Thermodynamics parameters of Pb ion onto G–Fe3O4
Temp (K)
DGꢂ
DHꢂ (kJ molꢁ1
)
DSꢂ (kJ molꢁ1
)
300
310
320
ꢁ1643
ꢁ4796
ꢁ9686
255.8
15.908
Table 3 Comparison of adsorption capacities of various adsorbents
Adsorbent
qm (mg gꢁ1
)
pH
Ref.
CNTs (HNO3)
35.60
15.58
26.90
16.4
1.21
69
5
5
5
5
5
5
32
33
33
34
35
Granular activated carbon
Powdered activated carbon
Microbead
Iron coated sand
G–Fe3O4
Fig. 3 Photo-Fenton degradation spectra of G–Fe3O4 composite at a
concentration of 0.2 g Lꢁ1 in the presence of (a) visible light, (b) H2O2
and (c) visible light and H2O2. (Inset shows the expanded view).
Present work
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reduction of Fe3+ to Fe2+ ion. Subsequently, Fe2+ would react
with H2O2 to produce hydroxyl radicals (OHc). The hydroxyl
radicals produced as a result of Fenton/photo-Fenton processes
attacks the dye adsorbed on the surface of the catalyst and
degrade it to CO2 and H2O.37,38
For the recycling of G–Fe3O4 composite, the absorbent was
separated, washed and then reused. The recycled adsorbed
behaviour is shown in Fig. S7.† It was observed that the removal
efficiency of the dye was over 95% up to 5 cycles. The photo-
catalytic activity of the recycled catalyst was also investigated,
and the results are shown in Fig. S7.† There was no noticeable
change in the photocatalytic activity of the recycled catalyst aer
ve cycles under visible light irradiation, indicating that the
magnetically separable photocatalyst is stable and effective for
the degradation of organic pollutants in water.
Fig. 5 (a) FE-SEM images of E. coli cells control and (b) rupture by G–
Fe3O4.
Destruction of bacterial membrane
FE-SEM was used to illustrate the interaction between G–Fe3O4
and E. coli. In Fig. 5(a), E. coli cells as a control show the entire
cell wall without any rupture, whereas Fig. 5(b) shows complete
disintegration of the cell wall. Spikes created on the E. coli cells
are a clear indication that G–Fe3O4 interacted with the cell wall.
This kind of cell wall rupture is irreversible damage induced by
direct contact with G–Fe3O4.
Antibacterial activity of G -Fe3O4
Fig. 4 shows that about 22.66% of the cells were only viable
when treated with G–Fe3O4, whereas with graphene about
65.33% of the cells were viable. The two activities were further
examined, i.e., concentration and time-dependent behaviour of
G–Fe3O4 on E. coli cells. Different concentrations of G–Fe3O4 (0,
25, 50, 100, 200 ppm) dispersions were incubated with E. coli
cells for 60 min and then plated for investigating cell viability.
Only E. coli cells and graphene were controls.
It was observed that 25 ppm G–Fe3O4 could inhibit 43% of
the cells because the recorded cell viability according to
Fig. S8(a)† is 57%. In addition, 100 ppm was sufficient to inhibit
bacteria for cell viability at 26.84% E. coli cells.
On the other hand, in a time-dependent study, we observed
that when 100 ppm of G–Fe3O4 dispersion was treated with E.
coli cells for various time intervals from 0 to 90 min, 60 min was
sufficient to kill bacteria, showing 23.74% cell viability
(Fig. S8(b)†). The loss of E. coli viability progressively decreased
with an increase in G–Fe3O4 concentration. E. coli viability
almost decreased from 57% by 25 ppm G–Fe3O4 to 21% with 200
ppm.
The majority of E. coli cells were killed aer incubation with
G–Fe3O4 at a concentration of 100 ppm. In a similar manner,
cell viability w.r.t. time from 0–90 min drastically reduced from
64.66% to 19.66%.
Fig. S9† shows the biocompatibility study. It was essential to
understand whether G–Fe3O4, if leached during water puri-
cation, would cause harm to normal cells. As shown, graphene
has less cell compatibility of about 54.11% in 2 mg mLꢁ1
,
whereas at the same concentration compatibility for G–Fe3O4 is
about 92.99%. Presence of Fe3O4, which supposedly is
biocompatible, has been a good support with graphene as a
biocompatible material.
Biochemical assay of total cellular protein degradation due to
G -Fe3O4
Degradation of protein due to the effect of graphene and G–
Fe3O4 was performed by Folin-Lowry test. Table 4 (in ESI†) also
indicates a reduction in protein content when E. coli was treated
with graphene and G–Fe3O4 individually. The reduction in total
protein content was found to be approximately 86.66% in the
case of G–Fe3O4. The total cellular protein degraded drastically
to 86.66 when treated with G–Fe3O4, as observed from Table 2,
whereas the control showed about 236% of protein. Perhaps
this is why only 22.66% of the E. coli could survive (i.e., 78%
mortality) when exposed to 100 ppm G–Fe3O4 nanocomposite,
suggesting that microbial death is due to the degradation of
total cellular protein.
Total protein assay (by SDS-PAGE methods) of E. coli due to G-
Fe3O4
Fig. 6 depicts the degradation of protein as analyzed by SDS-
PAGE electrophoresis upon treating E. coli cells with G–Fe3O4. It
is clear from the protein degradation results that stress-medi-
ated protein degradation of cell membrane le only one higher
density protein intact, whereas the rest of the proteins were
denatured, unlike proteins treated with graphene alone or
control.
Fig. 4 Cell viability of E. coli when treated with graphene and G–
Fe3O4.
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Conclusions
G–Fe3O4 was synthesized by a simple one-step solvothermal
route from graphene oxide and iron oxide nanoparticles. This
nanostructure serves as an efficient adsorbent for the removal
of heavy metal ions from waste water. The maximum adsorption
capacity of G–Fe3O4 was 17 mg gꢁ1 at an initial Pb concentration
of 20 mg Lꢁ1 and temperature of 310 K, indicating that G–Fe3O4
is effective for the adsorption of Pb(II) ions. The experimental
data tted well with the Langmuir isotherm model. The
monolayer adsorption capacity of Pb by G–Fe3O4 was found to
be 69 mg gꢁ1 at 310 K at pH 5. Thermodynamic investigations
indicated that the adsorption reaction was spontaneous and
endothermic. Results indicate that such materials could be
used as adsorbents for the removal of heavy metal ions. Pho-
todegradation of methylene blue dye was successfully achieved
up to 5 cycles. Antibacterial activity depicts that only 22.33% of
the cells were viable when treated with G–Fe3O4. Further protein
degradation conrmed that a higher molecular weight protein
was not denatured, whereas the rest of the proteins as
compared with E. coli had disintegrated. This was due to the
release of some ROS and disintegration of disulphide bonds.
Fig. 6 Protein degradation by G–Fe3O4.
Acknowledgements
Dr Pratap Kollu acknowledges DST-INSPIRE Faculty grant and
authors acknowledge SAIF/CRNTS and Central surface analyt-
ical facility of IIT Bombay. Also the authors gratefully
acknowledge VIT University, Vellore for supporting this work
under the research associate fellowship.
Fig. 7 Oxidative stress release by E. coli due to G–Fe3O4.
Glutathione oxidation due to G -Fe3O4 treatment
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