Chitosan nanoparticle (CS NP)-modified MnO2 nanoflakes were presented as a novel adsorbent for fast adsorption of Pb(II) from aqueous solution. Loading dense CS NPs onto mono-dispersive flower-like MnO2 nanostructures reduces the overlap of CS during adsorption, and thus improves the contact of functional adsorption sites on the surface of MnO2 nanoflakes with heavy metal ions. The results show that the removal efficiency of the nanoadsorbents reaches up to 93% in 3 min for Pb(II). In addition, the maximum adsorption capacity, effects of adsorbent dosage and pH value, and the reusability were investigated. The kinetic process and adsorption isotherm fit well with the pseudo-second-order model and Langmuir model, respectively. These findings provide a potential strategy to address the overlap issue of some common nanoadsorbents.

INTRODUCTION

Heavy metal ions in water are severely harmful to human health (Repo et al. 2013; Carpenter et al. 2015; Chiavola et al. 2015; Gupta et al. 2016). For example, it has been confirmed that Pb(II) increases the risk of cancer and intelligence defects (Clausen et al. 2011; Bhatluri et al. 2015; Du et al. 2016; Jiang et al. 2016). The efficient removal of heavy metal ions from aqueous solution is significant. In recent years, among many other approaches (such as membrane filtration and ion exchange), adsorption of pollutants from water by utilizing adsorbents has attracted broad attention because of its low cost and easy operation, etc. (Chen et al. 2013; Yin et al. 2013; Raychoudhury et al. 2015; Rodriguez-Flores et al. 2015). However, the adsorption rate and capacity of many adsorbents are as yet non-ideal, which limits the applications of adsorbents, especially for the purification of drinking water. Recently, nanomaterials with large surface area and functional groups which can capture pollutants specifically have been considered promising adsorbents (Liu et al. 2015; Sharma et al. 2015; Adeleye et al. 2016).

Chitosan (CS), which is derived from the deacetylation of chitin, possesses many fascinating features including abundance, multifunctionality, and non-toxicity (Sowasod et al. 2013; An et al. 2015; Sun et al. 2016). CS is considered to be a promising material for adsorption of heavy metal ions from water due to its special molecular structure, which contains a lot of amino and hydroxyl groups. However, CS aggregates in solution easily, which limits its potential applications. Despite many significant efforts focused on the cross-linking and modification of CS, reducing aggregation and improving the effective surface area exposed to pollutants remain a great challenge.

In a solution environment without sufficient stirring, CS nanoparticles (NPs) would aggregate and overlap due to the requirement of automatically reducing surface energy (Anson et al. 2004; Michiardi et al. 2007), resulting in the adsorption sites located in deep positions being covered by the top layer. In this condition, many adsorption sites can hardly be effective at contacting and adsorbing surrounding heavy metal ions. Here, we present a promising strategy to address the overlap issue of CS during adsorption. In our investigation, CS NPs are loaded onto a flower-like MnO2 nanostructure which is chemically stable and monodispersive. The flower-like MnO2 with its branched morphology keeps sufficient spaces between adjacent MnO2 materials, providing good conditions to ensure CS NPs located on each petal are well exposed to the surroundings. This adsorbent possesses the following features: (1) the aggregation of CS is effectively reduced due to the support of the three-dimensional (3D) flower-like MnO2, enabling the functional groups of CS to be exposed to heavy metal ions (Pb(II) was employed here); (2) the size of CS NPs is small, which means the CS possesses high surface area and high reactivity for chemically binding with Pb(II); (3) the structure of the adsorbents is stable because of the chemical stability of MnO2, and thus the adsorbents can remove heavy metal ions in neutral drinking water.

METHODS

The preparation procedure for MnO2 nanoflakes is similar to that described in a previous report (Zhang et al. 2013) with some modifications. Typically, 2 mmol of MnSO4·H2O and 2 mmol K2S2O8 were dissolved into 35 mL of de-ionized water under stirring for 30 min. The obtained solution was transferred to a Teflon-lined stainless steel autoclave, which was subsequently sealed and heated in an oven at 150 °C for 30 min. After naturally cooling down, the black precipitates were collected and washed alternately with water and ethanol. At last, the as-prepared samples were dried in a vacuum oven at 60 °C for 5 h.

In order to modify the MnO2 nanoflakes with CS NPs, the following steps were conducted. First, 0.1 mL of CH3COOH was added into 40 mL of de-ionized water to form a homogeneous solution. Then, 0.04 g of CS was added into the solution under stirring until clarification. After that, 0.05 g of MnO2 nanoflakes was put into the solution under stirring for 30 min. Na2CO3 was used for adjusting the pH value to 8 to nucleate the CS particles. Finally, after stirring for an additional 30 min, the samples were collected and washed with water, and dried in a vacuum oven at 60 °C for 3 h.

The samples were characterized by using a Philips X’Pert Pro X-ray diffractometer (XRD), an FEI Sirion 200 field emission scanning electronic microscope (FESEM), and a JEOL JEM-2010 transmission electron microscope (TEM). Fourier transform infrared (FTIR) spectra were recorded on a Shimadzu IR-440 spectrometer. Surface area was measured on a Coulter Omnisorp 100CX Brunauer–Emmett–Teller (BET).

Pb(II) was used as the target heavy metal ion for adsorption. In a typical procedure, 20 mL Pb(II) aqueous solutions, prepared by dissolving Pb(NO3)2 (Sigma-Aldrich Corp. without further purification) into de-ionized water at different concentrations ranging from ∼10 to 400 mg L−1, were dispersed with 20 mg of the as-obtained adsorbents. All adsorption was conducted in a vibration water bath at 25 °C. After a certain period of adsorption time, the solutions were separated and collected by centrifuge. The remaining concentrations of Pb(II) were measured by a Jarrell-Ash ICAP-9000 inductively coupled plasma. For the investigation of the adsorption effect of adsorbent dosage, series amounts of adsorbent ranging from 0.1 to 2 g L−1 were used, while the initial concentration of Pb(II) and adsorption time were 20 mg L−1 and 2 h for each adsorption, respectively. In pH-dependent adsorptions (initial concentration of Pb(II): 10 mg L−1; adsorption time: 3 h), pH values ranged from 2 to 6, and were adjusted by 0.5 M HCl or NaOH solutions.

RESULTS AND DISCUSSION

In FESEM images (Figure 1(a) and 1(b)), the MnO2 nanoflakes exhibit a spherical flower-like morphology. Several nanoflakes assemble around a core, exhibiting a whole architecture in a diameter of ∼2 μm. The thickness of each nanoflake is around 5 to 8 nm. After modifying with CS, the flower-like morphology kept stable, as shown in Figure 1(c). However, as shown in the high-magnification FESEM image (Figure 1(d)), many NPs were loaded on the surface of each nanoflake. It is indicated that some CS NPs were deposited onto the surface of MnO2, which is further supported by the TEM observations (Figure 1(e) and 1(f)). The size of CS NPs ranges from 45 to 65 nm. The composition of the samples is confirmed by the XRD patterns, as shown in Figure 2. All peaks are well indexed to the MnO2 phase (JCPDS card No. 80–1098) without any impurities. The diffraction peaks (2θ) at about 12.6 °, 25.3 °, 37.3 °, 42.6 °, 55.8 °, and 67.1 ° can be assigned to the (001), (002), (111), (112), (113), and (020) crystal planes of birnessite-type MnO2, respectively. However, it should be noted that there is no obvious XRD signal from CS NPs, which can be ascribed to the small amount of CS within the MnO2/CS composites. Because of that, further investigation about the composition was conducted through FTIR characterization (Figure 3).
Figure 1

(a) Low- and (b) high-magnification FESEM images of the MnO2 nanoflakes; (c), (d) FESEM images and (e), (f) TEM images of the MnO2 nanoflakes modified with CS NPs. Insets in (b) and (d) are corresponding schematic illustrations of nanoflakes without and with NP modifications, respectively.

Figure 1

(a) Low- and (b) high-magnification FESEM images of the MnO2 nanoflakes; (c), (d) FESEM images and (e), (f) TEM images of the MnO2 nanoflakes modified with CS NPs. Insets in (b) and (d) are corresponding schematic illustrations of nanoflakes without and with NP modifications, respectively.

Figure 2

XRD patterns of the MnO2 nanoflakes before and after modification with CS NPs.

Figure 2

XRD patterns of the MnO2 nanoflakes before and after modification with CS NPs.

Figure 3

FTIR spectra of the CS NP-modified MnO2 nanoflakes and the pure CS.

Figure 3

FTIR spectra of the CS NP-modified MnO2 nanoflakes and the pure CS.

In Figure 3, the peaks at about 2,870 and 1,410 cm−1 can be indexed to the stretching and bending vibrations of C − H, respectively, which rarely appear in pure MnO2 crystals reported previously (Nand et al. 2016; Ramesh et al. 2016), indicating the existence of organic components in the prepared adsorbents. The peak at ∼1,150 cm−1 is assigned to the characteristic band of C − N stretching vibration, which also supports the modification of MnO2 nanoflakes with CS. The peak located at about 1,600 cm−1 is assigned to the N − H scissor (Colilla et al. 2006), which can be attributed to the functional group of CS molecules, since pure MnO2 lacks this group (Aghazadeh et al. 2016). Impressively, this amine group should be able to play a significant role in the adsorption of positive Pb(II). In additional, as shown in Figure 4, the BET surface area is about 73.5 m2 g−1.
Figure 4

Typical N2 adsorption–desorption isotherm of the CS NPs-modified MnO2 nanoflakes.

Figure 4

Typical N2 adsorption–desorption isotherm of the CS NPs-modified MnO2 nanoflakes.

Figure 5 shows the adsorption performance of the CS NP-modified MnO2 nanoflakes for Pb(II). During the kinetic adsorption process, the initial concentration of Pb(II) is ∼10 mg L−1. In Figure 5(a), we can see that the removal efficiency rapidly reaches ∼93% in 3 min. It goes up to 98% after 10 min, exhibiting a high adsorption rate. The kinetic process is well fitted with a pseudo-second-order model, as shown in Figure 5(b). The pseudo-second-order model is shown in Equation (1) (Vinod & Anirudhan 2003): 
formula
1
where k2 is a pseudo-second-order rate constant. In this model, it is assumed that the difference between the concentration at time t and the saturated adsorption amount (qe) is the driving force of adsorption.
Figure 5

(a) Relationship between the removal efficiency and adsorption time; (b) the kinetic process fitted with a pseudo-second-order model.

Figure 5

(a) Relationship between the removal efficiency and adsorption time; (b) the kinetic process fitted with a pseudo-second-order model.

Figure 6(a) and 6(b) show the saturated adsorption isotherm and the Langmuir model-fitted results, respectively. The maximum adsorption capability of the CS NP-modified MnO2 nanoflakes is ∼102.5 mg g−1, far exceeding the MnO2 nanoflakes without CS NP modification (40.8 mg g−1) and CS bulk (36.6 mg g−1). In addition, both the high adsorption rate and capacity are competitive with some other reports about CS adsorbents (Zhu et al. 2012). The good performance can be ascribed to the special CS NP-modified nanostructure which provides a large surface area and numerous well-exposed functional groups to efficiently adsorb Pb(II). Under the support of the 3D flower-like structure, the adsorption sites of CS can expose and capture target ions in solution. About an adsorption mechanism, it would be a complicated interaction including electrostatic attraction and chemical bonding (complexation and chelation). Within the CS molecules, there are two chemically active sites (amine and hydroxyl groups) that would participate in the reactions with Pb(II). For example, it was demonstrated that the lone pair of electrons present in the nitrogen of the amine group can be supplied to the empty atomic orbital of metal ions, forming a CS–metal complex (Zhang et al. 2016).
Figure 6

(a) The saturated adsorption isotherm of the CS NP-modified MnO2 nanoflakes; (b) the fitting result by a Langmuir isotherm model.

Figure 6

(a) The saturated adsorption isotherm of the CS NP-modified MnO2 nanoflakes; (b) the fitting result by a Langmuir isotherm model.

In addition, the effects of adsorbent dosage and pH value were investigated. In Figure 7(a), the adsorption efficiency increases depending on the increase of dosage from 0.1 to 1.2 g L−1. It can be ascribed to the increase of adsorption sites and surface area. When the amount of adsorbents becomes excessive, the increase of removal efficiency becomes slow. This is perhaps due to the concentration gradient between solute concentration in solution and the one on the surface of the adsorbents (Vadivelan & Kumar 2005). Figure 7(b) shows the pH-dependent adsorption behavior. While increasing pH to 7.0 results in the precipitation of lead hydroxide (Naiya et al. 2008), in our investigations the pH values were lower than 6. In Figure 7(b), low pH value is not favorable for high efficiency. A high pH value increases the deprotonation of the adsorbent surface, forming negatively charged sites, which enhances the attraction between adsorbents and Pb(II) (Nassar 2010). It is indicated the presented adsorbents are more efficient for weak acidic and neutral water purification. Figure 7(c) shows the reusability of the adsorbents. In order to regenerate the CS-modified adsorbents, ethylenediaminetetraacetic acid (EDTA) solution was employed as a strong chelating agent (Yan et al. 2012). The removal efficiency slowly decreases with the increase of cycle number; however, the removal percentage still remains above 91% after four cycles, indicating good stability and reusability for Pb(II) adsorption.
Figure 7

The effects of (a) adsorbent dosage and (b) pH value on Pb(II) adsorption onto the CS NP-modified MnO2 nanoflakes. (c) Adsorbent regeneration cycles (conditions: initial concentration of Pb(II): 20 mg L−1; adsorbent dosage: 1 g L−1; adsorption time: 3 h; desorption time: 3 h; 0.05 mol L−1 EDTA as regeneration agent).

Figure 7

The effects of (a) adsorbent dosage and (b) pH value on Pb(II) adsorption onto the CS NP-modified MnO2 nanoflakes. (c) Adsorbent regeneration cycles (conditions: initial concentration of Pb(II): 20 mg L−1; adsorbent dosage: 1 g L−1; adsorption time: 3 h; desorption time: 3 h; 0.05 mol L−1 EDTA as regeneration agent).

CONCLUSIONS

In summary, we present an effective strategy to utilize CS NPs for fast adsorption of Pb(II) from aqueous solution. Dense CS NPs are modified on monodispersive flower-like MnO2 nanoflakes. The overlap of CS NPs can be significantly reduced, thereby improving the exposure of adsorption sites to heavy metal ions in solution. The adsorbents exhibit a high removal efficiency of about 93% in 3 min for Pb(II), and a maximum adsorption capability around 102.5 mg g−1. The results show that the adsorption efficiency increases depending on the increase of adsorbent dosage; in the range from 2 to 6, a high pH value is favorable for high adsorption efficiency; and the removal percentage remains above 91% after four regeneration cycles. It is expected that the presented nanoadsorbents could be a promising candidate for water purification; and the strategy of loading NP onto flower-like architectures would also stimulate some new studies about high-performance nanoadsorbents.

ACKNOWLEDGEMENTS

This work was financially supported by the State Key Project of Fundamental Research for Nanoscience and Nanotechnology (2011CB933700), and the National Natural Science Foundation of China (51002157, 21277146, 61071054, and 21177131).

REFERENCES

REFERENCES
Adeleye
A. S.
Conway
J. R.
Garner
K.
Huang
Y. X.
Su
Y. M.
Keller
A. A.
2016
Engineered nanomaterials for water treatment and remediation: costs, benefits, and applicability
.
Chemical Engineering Journal
286
,
640
662
.
Aghazadeh
M.
Asadi
M.
Maragheh
M. G.
Ganjali
M. R.
Norouzi
P.
Faridbod
F.
2016
Facile preparation of MnO2 nanorods and evaluation of their supercapacitive characteristics
.
Applied Surface Science
364
,
726
731
.
Anson
A.
Jagiello
J.
Parra
J. B.
Sanjuan
M. L.
Benito
A. M.
Maser
W. K.
Martinez
M. T.
2004
Porosity, surface area, surface energy, and hydrogen adsorption in nanostructured carbons
.
Journal of Physical Chemistry B
108
,
15820
15826
.
Carpenter
A. W.
Lannoy
C. F.
Wiesner
M. R.
2015
Cellulose nanomaterials in water treatment technologies
.
Environmental Science & Technology
49
,
5277
5287
.
Chiavola
A.
D'Amato
E.
Gavasci
R.
Sirini
P.
2015
Arsenic removal from groundwater by ion exchange and adsorption processes: comparison of two different materials
.
Water Science and Technology: Water Supply
15
(
5
),
981
989
.
Clausen
J. L.
Bostick
B.
Korte
N.
2011
Migration of lead in surface water, pore water, and groundwater with a focus on firing ranges
.
Critical Reviews in Environmental Science and Technology
41
,
1397
1448
.
Colilla
M.
Salinas
A. J.
Vallet
M.
2006
Amino-polysiloxane hybrid materials for bone reconstruction
.
Chemistry of Materials
18
,
5676
5683
.
Gupta
V. K.
Moradi
O.
Tyagi
I.
Agarwal
S.
Sadegh
H.
Shahryari-Ghoshekandi
R.
Makhlouf
A. S. H.
Goodarzi
M.
Garshasbi
A.
2016
Study on the removal of heavy metal ions from industry waste by carbon nanotubes: effect of the surface modification: a review
.
Critical Reviews in Environmental Science and Technology
46
,
93
118
.
Jiang
H. Y.
Zhao
Q. X.
Zeng
Y.
2016
Removal of Cd(II) and Pb(II) from aqueous solutions by modified polyvinyl alcohol
.
Desalination and Water Treatment
57
,
6452
6462
.
Liu
P.
Borrell
P. F.
Bozic
M.
Kokol
V.
Oksman
K.
Mathew
A. P.
2015
Nanocelluloses and their phosphorylated derivatives for selective adsorption of Ag+, Cu2+ and Fe3+ from industrial effluents
.
Journal of Hazardous Materials
294
,
177
185
.
Michiardi
A.
Aparicio
C.
Ratner
B. D.
Planell
J. A.
Gil
J.
2007
The influence of surface energy on competitive protein adsorption on oxidized NiTi surfaces
.
Biomaterials
28
,
586
594
.
Naiya
T. K.
Bhattacharya
A. K.
Das
S. K.
2008
Adsorption of Pb(II) by sawdust and neem bark from aqueous solutions
.
Environmental Progress
27
,
313
328
.
Raychoudhury
T.
Schiperski
F.
Scheytt
T.
2015
Distribution of iron in activated carbon composites: assessment of arsenic removal behavior
.
Water Science and Technology: Water Supply
15
(
5
),
990
998
.
Repo
E.
Warchol
J. K.
Bhatnagar
A.
Mudhoo
A.
Sillanpa
M.
2013
Aminopolycarboxylic acid functionalized adsorbents for heavy metals removal from water
.
Water Research
47
,
4812
4832
.
Rodriguez-Flores
L.
Marmolejo-Santillan
Y.
Perez-Moreno
F.
Castaneda-Ovando
A.
Sierra-Zenteno
A.
Flores-Castro
K.
Cadena-Zamudio
J. L.
2015
Adsorption of arsenic in dacitic tuff pretreated with magnesium oxide
.
Water Science and Technology: Water Supply
15
(
1
),
181
187
Sharma
V. K.
Zboril
R.
Varma
R. S.
2015
Ferrates: greener oxidants with multimodal action in water treatment technologies
.
Accounts of Chemical Research
48
,
182
191
.
Sowasod
N.
Nakagawa
K.
Charinpanitkul
T.
Tanthapanichakoon
W.
2013
Encapsulation of curcumin loaded oil droplets with chitosan based cryogel: influence of freezing condition on nanocapsule properties
.
Food Science and Technology Research
19
,
633
640
.
Vinod
V. P.
Anirudhan
T. S.
2003
Adsorption behaviour of basic dyes on the humic acid immobilized pillared clay
.
Water, Air, & Soil Pollution
150
,
193
217
.
Yin
C. Y.
Wei
Y. J.
Wang
F. W.
Chen
Y. H.
Bao
X.
2013
Magnetic hierarchical porous carbon sphere prepared for removal of organic pollutants in water
.
Materials Letters
104
,
64
67
.
Zhang
L.
Zeng
Y. X.
Cheng
Z. J.
2016
Removal of heavy metal ions using chitosan and modified chitosan: a review
.
Journal of Molecular Liquids
214
,
175
191
.