Dechlorination of 2,4-dichlorophenoxyacetic acid using biochar-supported nano-palladium/iron: Preparation, characterization, and influencing factors

Article abstract of DOI:10.1002/aoc.6010

In the present study, peanut shell, a green waste raw material, was used to prepare biochar (BC) and to obtain BC-supported nano-palladium/iron (BC-nPd/Fe) composites for removing 2,4-dichlorophenoxyacetic acid (2,4-D) from water. Characterization analysis demonstrated that nPd/Fe particles were well dispersed on the BC surface with weakened magnetic properties. The average particle diameter and specific surface area of nPd/Fe were 101.3 nm and 6.7 m² g^?1, whereas the corresponding values of the BC-nPd/Fe materials were 88.8 nm and 14.8 m² g^?1, respectively. Several factors were found to influence the dechlorination of 2,4-D, including the weight ratio of BC to Fe, Pd loading ratio, initial solution pH, 2,4-D concentration, and reaction temperature. Dechlorination results indicated that the 2,4-D removal and phenoxyacetic acid (PA) generation rates were 44.1% and 20.1%, respectively, in the nPd/Fe system, and 100.0% and 92.1%, respectively, in the BC-nPd/Fe system. The dechlorination of 2,4-D was well described by the pseudo-first-order kinetic model (R² > 0.97), and the observed rate constants k_obs were 0.0042 min (nPd/Fe) and 0.0578 min (BC-nPd/Fe), respectively. The reaction mechanism indicated that the dechlorination hydrogenation was the main process to remove 2,4-D from water in the BC-nPd/Fe system. In addition, BC inhibited the formation of a passivation layer on the particle surface during the reaction, thus maintaining the high reactivity of BC-nPd/Fe. The easy preparation technique, high 2,4-D dechlorination capacity, and mild reaction conditions suggest that BC-nPd/Fe may be a promising alternative composite to remove 2,4-D from water.

Full text of DOI:10.1002/aoc.6010

Received: 14 January 2020
DOI: 10.1002/aoc.6010
Revised: 25 July 2020
Accepted: 23 August 2020
F U L L P A P E R
Dechlorination of 2,4-dichlorophenoxyacetic acid using
biochar-supported nano-palladium/iron: Preparation,
characterization, and influencing factors
1
1
1
2
Hongyi Zhou
| Ning Huang
| Yongkang Zhao | Shams Ali Baig
|
1
Junchao Xiang
1
College of Environment, Zhejiang
University of Technology, Hangzhou,
Zhejiang, 310014, China
In the present study, peanut shell, a green waste raw material, was used to pre-
pare biochar (BC) and to obtain BC-supported nano-palladium/iron
2
Department of Environmental Sciences,
(BC-nPd/Fe) composites for removing 2,4-dichlorophenoxyacetic acid (2,4-D)
Abdul Wali Khan University, Mardan,
from water. Characterization analysis demonstrated that nPd/Fe particles were
well dispersed on the BC surface with weakened magnetic properties. The
average particle diameter and specific surface area of nPd/Fe were 101.3 nm
2
3200, Pakistan
Correspondence
Hongyi Zhou, College of Environment,
Zhejiang University of Technology,
Hangzhou, Zhejiang 310014, China.
Email: zhouhy@zjut.edu.cn
2
−1
and 6.7 m g , whereas the corresponding values of the BC-nPd/Fe materials
2
−1
were 88.8 nm and 14.8 m g , respectively. Several factors were found to
influence the dechlorination of 2,4-D, including the weight ratio of BC to Fe,
Pd loading ratio, initial solution pH, 2,4-D concentration, and reaction
temperature. Dechlorination results indicated that the 2,4-D removal and
phenoxyacetic acid (PA) generation rates were 44.1% and 20.1%, respectively,
in the nPd/Fe system, and 100.0% and 92.1%, respectively, in the BC-nPd/Fe
Funding information
National Natural Science Foundation of
China, Grant/Award Number: 21477108
system. The dechlorination of 2,4-D was well described by the pseudo-
2
first-order kinetic model (R > 0.97), and the observed rate constants k were
obs
0.0042 min (nPd/Fe) and 0.0578 min (BC-nPd/Fe), respectively. The reaction
mechanism indicated that the dechlorination hydrogenation was the main
process to remove 2,4-D from water in the BC-nPd/Fe system. In addition, BC
inhibited the formation of a passivation layer on the particle surface during
the reaction, thus maintaining the high reactivity of BC-nPd/Fe. The easy
preparation technique, high 2,4-D dechlorination capacity, and mild reaction
conditions suggest that BC-nPd/Fe may be a promising alternative composite
to remove 2,4-D from water.
K E Y W O R D S
2
,4-D, biochar, Pd/Fe nanoparticles, reductive dechlorination
1
| INTRODUCTION
controlling broad-leafed weeds and increasing agriculture
[
1]
productivity in recent years. The most common trails
of 2,4-D in the environment result from its direct applica-
tion as herbicide or the wastewater generated during its
The
chlorinated
phenoxy
herbicide
2,4-
dichlorophenoxyacetic acid (2,4-D), a common plant
growth regulator, has been extensively employed for
[
2]
manufacturing and transportation.
Research reports
Appl Organomet Chem. 2020;e6010.
wileyonlinelibrary.com/journal/aoc
© 2020 John Wiley & Sons, Ltd.
1 of 15
https://doi.org/10.1002/aoc.6010

2
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ZHOU ET AL.
have shown that 2,4-D occurrence in the environment
has different degrees of impact on human health and eco-
experiments on 2,3,4-trichlorobiphenyl (PCB 21) using
nZVI (1 g) and nZVI/Pd (0.1 g with 0.25% Pd) for
20 and 70 days, respectively. The results indicated that
palladization caused a dechlorination rate for PCB
21 500 times higher than before. In addition, Lien
[2,3]
systems.
Generally, 2,4-D is both persistent and toxic,
classified as an environmental endocrine disruptor and a
[4–7]
carcinogenic substance.
In addition, it has been
[
30]
reported that 2,4-D has been frequently detected in
surface or groundwater worldwide because of its poor
et al.
pointed out that in the dechlorination experi-
ments of trichloroethylene (TCE), the reactivity of the
Pd/Fe nanoparticles (with 1% Pd) was approximately
70 times greater than that of the unpalladized nano-
scale Fe particles.
[8]
biodegradability and volatility.
Therefore, effective
treatments are required to efficiently remove 2,4-D
from water.
At present, there are many technologies available
for the remediation of chlorinated pollutants from
The newly prepared Pd/Fe nanoparticles tended to
quickly form large flocs or aggregates because of the mag-
netic attractions and van der Waals forces. Hence, their
[
9]
aqueous systems, such as biological degradation,
[10]
[11]
electroreductive dechlorination,
advanced oxidation,
and nano-zero-valent iron (nZVI) reduc-
Advanced oxidation technology and photo-
stability and reactivity in contaminated environments
[12,13]
[31,32]
adsorption,
decreased as the reaction proceeded.
In addition,
[14]
tion.
the high surface energy of Pd/Fe nanoparticles leads to
[33]
catalysis possess the advantages of short reaction time
and high degradation efficiency, but the disadvantage
severe aggregation, which may reduce the reactivity.
In order to solve the problem of easy agglomeration,
many researchers have loaded nZVI or Pd/Fe
nanoparticles on various functional materials to form
composite materials. The common loaded matrices are
[10,11]
associated with high-energy intensity.
The energy
consumption of any biodegradation technology is
generally lower, but the selection and the acclimation
[
34]
[35]
[36]
of microbial communities are
a
long-term and
activated carbon,
silica,
bentonite,
multiwalled
and biochar
[
37]
[38]
complex work because of the toxicity of chlorinated con-
carbon nanotubes,
metal oxides,
[9]
[39]
taminants.
Adsorption and nZVI technology have
(BC).
Among them, the BC produced by pyrolysis of
become common methods for removing chlorinated
pollutants from water environments owing to their
environmental friendliness, ease of operation, and high
biomass under oxygen-limited or oxygen-free conditions
has attracted research interest because it is considered to
be a promising environmental remediation material.
[40]
[12,14]
efficiency.
In practical applications, these two tech-
Plant materials, woodchips, wood biomass, crop
waste, carbon-rich organic biomass, and solid sludge
waste are often considered as excellent raw materials for
nologies are usually combined to achieve better
processing results.
[
41]
In recent years, nZVI has attracted wide attention
because of its large surface area and high reaction prop-
BC.
BC has a porous structure and possesses a large
[42,43]
surface area rich in diverse functional groups.
BC
[15]
erties.
Generally, nZVI has a low redox potential of
can be successfully used to disperse and stabilize
nanoparticles, thereby enhancing their performance in
−
0.44 V, which is usually used as an electron donor
[15,16]
[39]
[42]
during the reduction process.
Therefore, nZVI is
environmental applications.
Wang et al.
reviewed
extensively used to treat pollutants in water environ-
the preparation of BC from different raw materials and
[16]
ments that contain chlorinated organics,
heavy
its capacity to remove various contaminants after loading
[17]
[18]
[19]
[44]
metals,
organic dyes,
and nitrates.
However, as
nZVI. Qian et al.
produced BC by pyrolyzing rice
the reaction progresses, their reactivity for pollutants
could be substantially reduced by the passivation layers
straw at different temperatures and loaded nZVI on the
obtained composite to remove hexavalent chromium
[
45]
(
e.g., iron oxides and hydroxides) on the surface of
(Cr(VI)) in a water solution. Devi and Saroha
synthe-
[20]
nZVI.
In addition, the passivation layer suppresses
sized BC (prepared from paper mill sludge) Ni-ZVI-
combined magnetic materials (Ni-ZVI-MBCs), which
were utilized as an adsorbent to remove pentachlorophe-
nol (PCP) from wastewater. These studies focused on the
simultaneous adsorption and reduction of pollutants by
BC-supported nZVI or bimetals. Nevertheless, few studies
have reported the removal of chlorinated organic pollut-
the electron transfer efficiency between zero-valent iron
0
[21]
(
ZVI, Fe ) and pollutants.
The reactivity of nZVI
could be further enhanced by loading another noble
[22,23]
[20,24]
[25]
[26]
metal (such as Pd,
Ni,
Cu,
and Ag ) on
the nZVI surface. These noble metals can contribute to
the formation of Fe-based bimetallic nanoparticles, of
which the Pd/Fe nanoparticles exhibit the greatest reac-
tivity. Thus, Pd/Fe nanoparticles have been widely
utilized for the catalytic reduction of compounds, which
[46,47]
ants by BC-supported iron-based bimetal particles.
Therefore, it is of great significance to investigate BC
composites that can efficiently reduce chlorinated
organic contaminants from water under different envi-
[
27–29]
typically have poor reactivity with nZVI.
instance, Zhuang et al.
For
[22]
[48,49]
performed dechlorination
ronmental conditions.

ZHOU ET AL.
3 of 15
Peanut shells are usually regarded as food by-
products that are normally discarded in the food
BC-supported nZVI (BC-nZVI) was formed, and the reac-
tion can be depicted as Equation 1:
[50]
industry and agriculture.
Owing to its porous struc-
Fe2 + + 2BH + 6H O ! Fe + 2BðOHÞ + 7H " : ð1Þ
−
0
ture, the peanut shell could become an excellent
supporting material for nZVI or bimetal. In the present
study, a peanut shell was used as the raw material of
BC to synthesize BC-supported nano-palladium/iron
4
2
3
2
A 10.0 ml of K PdCl solution was added to the BC-nZVI
2
6
2+
0
(BC-nPd/Fe),
a
new composite material for the
solution after 30.0 min. Then, Pd was reduced by Fe ,
effective dechlorination of 2,4-D in water. In addition,
the effects of reaction parameters on 2,4-D dechlorina-
tion by BC-nPd/Fe were explored, and the reaction
mechanism of 2,4-D dechlorination by BC-nPd/Fe was
further discussed by characterizing the material
properties and investigating the reductive dechlorina-
tion process.
resulting in
Equation 2:
a
BC-nPd/Fe solution according to
PdCl2 + 2Fe ! 2Fe + Pd + 6Cl :
−
0
2 +
0
−
ð2Þ
6
After 60.0 min of continuous reaction, BC-nPd/Fe
particles were separated by an externally enhanced mag-
netic field, washed three times with oxygen-free DI water
and ethanol, and then collected by drying in a vacuum
2
2
| MATERIALS AND METHODS
.1 | Materials
ꢁ
oven at 60 C for 6.0 h. The synthesis steps of nPd/Fe are
consistent with the aforementioned methods. In this
study, the Pd loading (weight ratio of Pd to Fe) was calcu-
lated using Equation 3:
The 2,4-D (C H Cl O , 98%), 2-chlorophenoxyacetic acid
8
6
2 3
(2-CPA, C H ClO , >98%), 4-chlorophenoxyacetic acid
8 7 3
(
4-CPA, C H ClO > 98%), and phenoxyacetic acid (PA,
MPd
MK PdCl × 26:78%
8
7
3
2
6
Pd=Feð%Þ =
=
:
ð3Þ
C H O , 98%) were purchased from Alfa Aesar (Tianjin,
MFe
MFe
8
8 3
China). Sodium borohydride (NaBH , ≥97%), ferrous sul-
4
fate (FeSO ꢀ7H O, 99–101%), ethanol (CH CH OH,
4
2
3
2
>
99.7%), and methanol (CH OH, 99.5%) were obtained
3
from Sinopharm Chemical Regent Co., Ltd. (Shanghai,
China). Potassium hexachloropalladate (K PdCl , 99%)
2.3 | Batch experiments
2
6
was purchased from Macklin Biochemical Co., Ltd.
Shanghai, China). Methanol (CH OH, high-performance
liquid chromatography [HPLC] grade) was purchased
from Tedia Company Inc. (Shanghai, China). Deionized
Batch experiments of 2,4-D dechlorination were per-
formed in a three-necked flask immersed in a constant-
temperature water bath. The 2,4-D solutions and a
certain amount of DI water were added to the flask con-
taining freshly prepared nPd/Fe or BC-nPd/Fe particles
to 500.0 ml of total reaction volume. The reaction solu-
(
3
(
DI) water, previously deoxygenated by bubbling N gas
2
for 1.0 h, was used in all the experiments.
−
1
tion was continuously stirred (n = 200 r min ) under an
N atmosphere. The effects of 2,4-D dechlorination by
2
2.2 | Preparation of BC-nPd/Fe
BC-nPd/Fe under different experimental conditions
(
MBC:MFe [weight ratio of BC to Fe], Pd loading, initial
The BC precursor material, peanut shells, was obtained
from a local market in Hangzhou City, China. Briefly,
peanut shells were smashed and passed through a
pH, initial 2,4-D concentration, and reaction tempera-
ture) were investigated. The reaction solution was sam-
pled at different reaction time intervals, passed through a
0.22-μm microporous membrane and stored in a sample
vial for HPLC analysis. Samples were processed in tripli-
cate, and the mean values were used for the analysis.
ꢁ
1
50-mesh screen and then calcined at 500 C for 3.0 h in a
ꢁ
−1
tube furnace at a heating rate of 10.0 C min under
−
1
nitrogen (N ) flow (100.0 ml min ).
2
The BC-nPd/Fe was synthesized in a three-necked
−
1
flask under mechanical stirring (200 r min ) under N
2
atmosphere. First, BC was added to 400.0 ml of
2.4 | Characterization analyses
FeSO ꢀ7H O solution and sonicated for 30.0 min to
4
2
ensure that the BC was evenly dispersed in the solution.
The Brunauer–Emmett–Teller (BET) surface areas of
nPd/Fe and BC-nPd/Fe were measured using an N2
adsorption method with a surface analyzer (ASAP2020M,
Next, 80.0 ml of NaBH solution was added dropwise
4
−
2+
to the solution (the molar ratio of BH /Fe was 2.0).
4

4
of 15
ZHOU ET AL.
Micromeritics Instrument Corp., USA). The surface mor-
phology of the sample was analyzed by scanning electron
microscopy (SEM, S4700, Hitachi, Japan), and the surface
element distribution of the material was analyzed by
energy dispersive spectroscopy (EDS, NSS7, Thermo
Fisher Scientific, China). X-ray diffraction (XRD) spectra
of the materials were recorded with X'Pert PRO
indicate that the nPd/Fe particles are interconnected by a
chain structure and agglomeration, as shown in
Figure 1a. The average particle diameter was 101.3 nm
(Figure 1b). Such aggregation could be explained by the
severe magnetic attractions and van der Waals forces
interaction between the particles. The agglomeration of
the nPd/Fe particles decreased the reactivity because the
[33]
(
5
PANalytical, Netherlands) using K radiation over the
–80 scan range. Magnetic characterization was per-
surface reaction sites were covered by other particles.
ꢁ
In contrast, more dispersed nPd/Fe particles were clearly
observed on the surface of the BC-nPd/Fe material
(Figure 1c). Moreover, the average diameter of these par-
ticles decreased to 88.8 nm (Figure 1d).
formed using a vibrating sample magnetometer (VSM,
VersaLab, USA) instrument between a magnetic field
from −15,000 to 15,000 Oe. Fourier transform infrared
(
FTIR) spectra were collected using an IRTracer-100
spectrometer in the wavenumber range of
,000–500 cm . Thermogravimetric analyses (TGAs)
were conducted on a TGA 4000 analyzer (PerkinElmer,
Figure 2a presents the N adsorption–desorption iso-
2
therms. The specific surface areas of nPd/Fe and BC-
−
1
2
−1
4
Pd/Fe calculated were 6.7 and 14.8 m g , respectively.
The VSM characterization of nPd/Fe and BC-nPd/Fe is
illustrated in Figure 2b. The hysteresis loops of both
materials were typically S-like curves, and the intensity
of the saturation magnetization values were 2.35 and
0.64 emu mg− for nPd/Fe and BC-nPd/Fe, respectively.
Compared with nPd/Fe, the lower magnetic saturations
of BC-nPd/Fe could be attributed to the introduction of
BC. The SEM, BET, and VSM characterization results
indicated that the presence of BC reduced the intensive
magnetic attraction between nPd/Fe particles, thereby
ꢁ
ꢁ
USA) from 30 C to 800 C at a heating rate of
1
positions of the samples were analyzed by X-ray photo-
electron spectroscopy (XPS, Kratos AXIS Ultra DLD,
Shimadzu Ltd, Japan).
ꢁ
0.0 C min under an N atmosphere. The surface com-
2
−1
1
2
.5 | Analytical methods
The concentration of organic matter (2,4-D, PA, 2-CPA, and
-CPA) in aqueous solution were quantified using HPLC
Agilent Technologies 1200 Series, USA) equipped with an
suppressing
the
agglomeration
between
these
4
(
nanoparticles. The dispersion of nPd/Fe particles with
smaller size on the surface of BC means that BC-Pd/Fe
composites possess more active reaction sites and larger
contact area with contaminants, which is conducive to
the reductive dechlorination of 2,4-D.
Eclipse XDB-C18 HPLC column (150 mm × 4.6 mm;
I.D. 5 μm Agilent, USA), and an ultraviolet detector (UVD)
[51]
operated at a wavelength of 210 nm. Methanol/ultrapure
water (60/40, v/v) was selected as the HPLC mobile phase
The FTIR spectra of BC, nPd/Fe, and BC-nPd/Fe are
shown in Figure 3a. The bands around 1,490 cm repre-
−1
−1
(
pH was adjusted to 3.0 by adding 0.5 mol L H PO ) at a
3
4
−1
flow rate of 1.0 ml min with an injection volume of 20 μl.
The column temperature was set at 30 C, and the total
sented the stretching vibration of aromatic C O and ben-
ꢁ
−1
zene ring C C, and the band at 750 cm was assigned
[
43]
analysis time was 10.0 min. The pseudo-first-order kinetic
to the bending vibration of C H and O H.
The
were
[30]
−1
model
was applied to describe the dechlorination of
absorption peaks at 1,100–900 and 553 cm
[
3,38]
2,4-D, which can be represented by the following equation:
ascribed to the Fe O stretch vibration.
These charac-
teristic peaks can be clearly observed on BC-nPd/Fe, indi-
cating that nPd/Fe was successfully loaded on the
BC. TGAs of BC, nPd/Fe, and BC-nPd/Fe are shown in
dC
dt
=
−kobsC,
ð4Þ
ꢁ
Figure 3b. The weight loss of BC below 250 C was due to
−
1
where C is the 2,4-D concentration (mg L ) in the aque-
ous solution at time t (min) and kobs is the observed
pseudo-first-order kinetic rate constant (min).
the desorption of water and small organic molecules. As
the temperature increased, the organic structure in the
BC decomposed, with a total weight loss of 31.7% at
ꢁ
[40]
8
1
00 C.
The mass of the nPd/Fe material increased to
ꢁ
20.5% of the original sample at 800 C, which may be
3
3
| RESULTS AND DISCUSSION
.1 | Characterization analysis results
attributed to the thermal decomposition of iron oxides
(
which was verified by XPS) under high-temperature con-
[
27]
ditions.
The BC-nPd/Fe composites slowly lost weight
ꢁ
below 450 C, with a maximum mass loss of 2.9%. The
sample gradually gained weight in the temperature range
Figure 1 presents the surface morphology and particle
size distribution of nPd/Fe and BC-nPd/Fe. The results
ꢁ
of 450–650 C; then, its weight almost reached

ZHOU ET AL.
5 of 15
FIGURE 1 SEM images of (a) nPd/Fe and (c) BC-nPd/Fe; (b) and (d) are the particle size distribution histograms to (a) and (c).
BC-nPd/Fe, biochar-supported nano-palladium/iron; SEM, scanning electron microscopy
FIGURE 2 (a) N
2
adsorption–desorption isotherms and (b) magnetization curves of nPd/Fe and BC-nPd/Fe. BC-nPd/Fe, biochar-
supported nano-palladium/iron

6
of 15
ZHOU ET AL.
FIGURE 3 (a) FTIR spectra, (b) thermogravimetric curves, and (c) X-ray diffraction patterns of BC, nPd/Fe, and BC-nPd/Fe.
BC-nPd/Fe, biochar-supported nano-palladium/iron; FTIR, Fourier transform infrared
equilibrium, with a final weight 104.4% of that of the
original sample. The results further indicated that nPd/Fe
was successfully supported on BC, and the BC-nPd/Fe
inhibited the formation of the passivation layer during
the reaction so that the Fe was exposed on the surface of
the particles, thereby maintaining the reactivity of BC-
nPd/Fe. It is difficult to observe the characteristic peaks
of Pd in the XRD patterns because of the extremely low
Pd content in the materials.
The local element information of the BC-nPd/Fe
materials before and after the reaction was determined
by EDS. As shown in Figure 4, the elements on the sur-
face of BC-nPd/Fe are mainly composed of C, Fe, O, and
Pd. The layered image further confirmed that the nPd/Fe
particles were dispersed on the BC surface. In compari-
son, the O weight ratio increased from 9.51% to 19.82%,
whereas the Fe weight ratio decreased from 26.25% to
22.04% on the surface of the BC-nPd/Fe after the reac-
tion. The results indicated that iron was oxidized during
0
ꢁ
ꢁ
composites remained stable between 650 C and 800 C in
an N atmosphere. The XRD patterns of BC, nPd/Fe, and
2
BC-nPd/Fe are shown in Figure 3c. The characteristic
ꢁ
[52]
peak of carbon (2θ = 22.7 )
can be clearly observed in
ꢁ
BC. The peak at 2θ = 44.7 was the typical diffraction
0
[44]
peak of ZVI (Fe ),
which was clearly observed in the
XRD patterns of nPd/Fe and BC-nPd/Fe before the reac-
0
tion. The characteristic peaks for both carbon and Fe
appearing in the BC-nPd/Fe suggest that the composites
were successfully synthesized. However, after the reac-
0
tion, a significant Fe peak was observed again in BC-
0
nPd/Fe, sharply contrasting with the faint Fe peak
intensity of nPd/Fe. A possible explanation is that BC

ZHOU ET AL.
7 of 15
FIGURE 4 Elemental mappings of BC-nPd/Fe (a) before and (c) after reaction; (b) and (d) are the EDS spectrums to (a) and (c).
BC-nPd/Fe, biochar-supported nano-palladium/iron; EDS, energy dispersive spectroscopy
the reaction to form a passivation layer covering the sur-
face of the material.
The XPS analyses for Fe 2p and Pd 3d before and after
the nPd/Fe and BC-nPd/Fe reactions are shown in
Figure 5. The XPS of Fe 2p shows that the two broad
peaks at 710.9 and 724.1 eV represent Fe 2p3/2 and Fe
respectively, and the values were 15.3% and 81.0% in the
fresh BC-nPd/Fe materials, respectively. The BC-nPd/Fe
0
particles with higher content of reducing agent (Fe ) and
0
catalyst (Pd ) displayed greater reactivity at the beginning
0
0
of the reaction, after which, the contents of Fe and Pd
on the surface of the nPd/Fe particles decreased to 4.6%
and 39.0%, respectively. The results suggested that nZVI
was oxidized to form a passivation layer (Equations 5 and
6) covering the surface of the particles during the reac-
2
p1/2, respectively, which imply that Fe oxides or hydrox-
ides (Fe O , Fe(OH) , and FeOOH) are present on the
x
3
[14,37,53]
surface of these particles (Figure 5a,c).
The char-
0
acteristic peak at 706.9 eV corresponds to the binding
tion, and Pd as the catalyst was also quickly consumed,
0
[37]
energy of Fe .
As shown in Figure 5b,d, the Pd 3d5/2
resulting in a decrease in the reaction activity of
nPd/Fe. Compared with nPd/Fe, BC-nPd/Fe particles
and Pd 3d3/2 photoelectron peaks are located at 335.2 and
[51]
0
0
0
0
3
40.3 eV, respectively.
The contents of Fe and Pd in
maintained a higher Fe (11.3%) and Pd (48.3%) content
after the reaction, indicating that the presence of BC
the fresh nPd/Fe particles were 13.0% and 59.9%,

8
of 15
ZHOU ET AL.
FIGURE 5 XPS spectra for (a) Fe2p of nPd/Fe, (b) Pd3d of nPd/Fe, (c) Fe2p of BC-nPd/Fe, and (d) Pd3d of BC-nPd/Fe. BC-nPd/Fe,
biochar-supported nano-palladium/iron; XPS, X-ray photoelectron spectroscopy
0
0
2 + =3 +
−
effectively hindered the passivation layer generated (the
passivation layer present on the surface of the nPd/Fe
particles was thinner), so that more Fe and Pd were
detected on the surface of the materials. In addition, the
thinner passivation layer was beneficial to the reaction of
Fe −HFeOH + H O ! Fe −Fe
+ H + 2OH :
n−1
2
n−1
2
ð6Þ
3.2 | Factors affecting 2,4-D
dechlorination by BC-nPd/Fe
nZVI corrosion and H dissociation. Therefore, the car-
2
rier BC retained the high reduction reactivity of nPd/Fe,
which was conducive to the dechlorination of 2,4-D by
BC-nPd/Fe.
3.2.1 | Effect of M :M
BC
Fe
Figure 6 shows the effect of MBC:M on 2,4-D dechlori-
Fe
1
−
nation. The nZVI content (1.0 g L ) was controlled and
the quantity of BC was changed to obtain BC-nPd/Fe
0
n
0
n−1
Fe + H2O ! Fe
−HFeOH,
ð5Þ

ZHOU ET AL.
9 of 15
FIGURE 6 Effect of weight ratio of BC to Fe on dechlorination of 2,4-D by BC-nPd/Fe (nZVI = 1.0 g L−1, C0 2,4-D = 20.0 mg L−1
Pd/Fe = 0.5%, pH = 7.0, T = 25.0 C, n = 200 r min ). 2,4-D, 2,4-dichlorophenoxyacetic acid; BC-nPd/Fe, biochar-supported nano-
,
ꢁ
−1
palladium/iron; nZVI, nano-zero-valent iron
FIGURE 7 Effect of Pd loading on 2,4-D dechlorination by BC-nPd/Fe (nZVI = 1.0 g L−1, C0 2,4-D = 20.0 mg L−1, pH = 7.0, T = 25.0 C,
ꢁ
BC:MFe = 1:1, n = 200 r min− ). 2,4-D, 2,4-dichlorophenoxyacetic acid; BC-nPd/Fe, biochar-supported nano-palladium/iron; nZVI, nano-
1
M
zero-valent iron
materials with different MBC:MFe ratios (0.25:1, 0.5:1, 1:1,
results, an MBC:MFe ratio of 1:1 was chosen for further
experiments.
2
:1, and 3:1, respectively). At higher MBC:MFe ratios
(
(
(
≥1:1), 2,4-D was completely removed by BC-nPd/Fe
Figure 6a), whereas the phenoxyacetic acid
PA) generation was higher than 88% (Figure 6b) after
3.2.2 | Effect of Pd loading
2
00 min of reaction. The results showed that BC-nPd/Fe
composites with a high MBC:MFe ratio had a significant
dechlorination effect on 2,4-D. In contrast, BC-nPd/Fe
with an MBC:MFe ratio of 0.25:1 had a poor dechlorina-
tion effect on 2,4-D. The possible reason may be that the
nPd/Fe particles cannot be sufficiently dispersed on the
Pd loading had a significant effect on the dechlorination
of 2,4-D (Figure 7). The 2,4-D removal rate increased
from 51.5% to 100.0% with an increase in Pd loading
(from 0.1% to 0.5%) (Figure 7a). The increasing trend of
the PA generation rate (from 26.9% to 92.1%) was more
obvious, and resultantly the higher Pd loading, the higher
material surface with low BC content. When the MBC
:
MFe ratio was greater than 1:1, the PA generation rate
increased slightly, which was 90.5% (MBC:MFe = 2:1) and
dechlorination efficiency. During the reaction, H was
generated due to the corrosion of nZVI (Equations 7 and
2
9
2.1% (MBC:MFe = 3:1). Based on the aforementioned
8). The generated H was quickly adsorbed by Pd (a good
2

10 of 15
ZHOU ET AL.
[
30]
catalyst
)
and converted into strongly reducible
increased to 8.5, the 2,4-D removal and PA generation
rate decreased to 75.1% and 29.9%, respectively. The
remarkable decline (particularly of PA generation, as
shown in Figure 8b) could be explained by the following
reasons. First, the OH− concentration in the solution
increased with increasing initial pH value. A large
hydrogen (H*) (Equation 9), facilitating the
hydrodechlorination of 2,4-D. Therefore, sufficient Pd
loading can produce the reducible H* required for the
reaction in time. In addition, metal Pd as a catalyst
reduces the activation energy,
reaction of 2,4-D is more likely to occur. Moreover, it is
generally believed that metal Pd could form a galvanic
[48]
so the dechlorination
−
amount of OH reacts easily with free iron ions to form a
passivation layer, which covers the surface of the nPd/Fe
particles and hinders the corrosion of nZVI. Second, the
passivation layer also hindered the contact of 2,4-D with
the BC-nPd/Fe surface reaction sites and thus suppressed
[54]
cell with ZVI.
More efficient electron transfer between
BC-nPd/Fe and 2,4-D at high Pd loading ratios may be
one of the reasons for the effective improvement of the
dechlorination efficiency of 2,4-D.
+
the dechlorination reaction. On the contrary, H in the
Neutral or alkaline conditions:
solution could dissolve the passivation layer and make
the corrosion reaction of nZVI to easily occur, which
facilitates the dechlorination of 2,4-D. Third, 2,4-D
0
2 +
−
Fe + 2H O ! Fe + H + 2OH :
ð7Þ
2
2
[23]
carries negative charges in an alkaline solution.
[
42]
Acidic conditions:
Because of the negative surfaces of BC,
the electro-
static repulsion between BC and 2,4-D resulted in a
decreased dechlorination efficiency of BC-nPd/Fe for
0
+
2 +
Fe + 2H ! Fe + H2,
ð8Þ
ð9Þ
2,4-D. In general, 2,4-D removal and PA generation
Pd
(which varied with the initial pH values) did not change
significantly under acidic or neutral conditions (pH ≤ 7.0),
whereas the downward trend gradually became apparent
in an alkaline environment (pH > 7.0). For example, the
PA generation at an initial pH of 7.0 was 82.4%, which
was 2.76 times that at pH 8.5. Hence, it can be considered
that an alkaline reaction environment is more detrimen-
tal to 2,4-D dechlorination.
ꢂ
H2 ! 2H :
3
.2.3 | Effect of initial solution pH
The pH value is a critical factor in the reduction of
[55]
organics.
As shown in Figure 8, 2,4-D was completely
removed with an increase in initial pH from 5.0 to 7.0,
whereas the PA production decreased from 92.7% to
3.2.4 | Effect of initial 2,4-D concentration
8
9.4%. This insignificant change is attributed to the fact
that ZVI could produce more corrosion in acidic solu-
tions, so BC-nPd/Fe could maintain excellent dechlorina-
tion performance for 2,4-D. When the initial pH
The effect of initial 2,4-D concentration on the dechlori-
nation efficiency was evaluated, and the results are
shown in Figure 9. The 2,4-D removal rate after
FIGURE 8 Effect of initial pH on 2,4-D dechlorination by BC-nPd/Fe. (nZVI = 1.0 gꢀL , C0 2,4-D = 20.0 mg L−1, Pd/Fe = 0.5%,
−1
ꢁ
−1
T = 25.0 C, MBC:MFe = 1:1, n = 200 r min ). 2,4-D, 2,4-dichlorophenoxyacetic acid; BC-nPd/Fe, biochar-supported nano-palladium/iron;
nZVI, nano-zero-valent iron

ZHOU ET AL.
11 of 15
FIGURE 9 Effect of initial 2,4-D concentration on 2,4-D dechlorination by BC-nPd/Fe (nZVI = 1.0 g L−1, C0 2,4-D = 20.0 mg L−1
,
pH = 7.0, Pd/Fe = 0.5%, T = 25.0 C, MBC:MFe = 1:1, n = 200 r min−1). 2,4-D, 2,4-dichlorophenoxyacetic acid; BC-nPd/Fe, biochar-supported
ꢁ
nano-palladium/iron; nZVI, nano-zero-valent iron
2
00.0 min was 100.0%, 100.0%, 93.7%, 90.2%, and 85.9%
and PA generation were observed. When the reaction
temperature was increased to 40.0 C, the 2,4-D removal
ꢁ
when the concentrations of 2,4-D were 10.0, 20.0, 30.0,
−1
4
0.0, and 50.0 mg L , respectively (Figure 9a). Addition-
ally, the PA generation rates were 94.5%, 92.1%, 71.5%,
1.0%, and 59.9%, respectively (Figure 9b). The results
rate decreased to 34.4%, and the PA generation rate was
only 10.1% (Figure 10a). 2,4-D molecules can easily move
to the surface of the nanoparticles at higher tempera-
tures, but excessive temperature could accelerate the oxi-
dation of nZVI and form a passivation layer to cover the
6
showed that the dechlorination efficiency of BC-nPd/Fe
for 2,4-D decreased with an increase in the initial 2,4-D
concentration. Interestingly, a lower initial concentration
[
28,29]
reaction sites,
thus reducing the dechlorination abil-
−
1
ꢁ
of 2,4-D (10.0–20.0 mg L ) was rapidly adsorbed by the
surface reaction site and then reduced by BC-nPd/Fe,
resulting in greater 2,4-D removal and PA generation.
Simultaneously, the reaction sites of the composites were
fully utilized. However, the concentration of 2,4-D con-
tinued to increase, and BC-nPd/Fe did not have sufficient
reaction sites for the dechlorination of 2,4-D, leading to a
reduction in dechlorination efficiency. In addition, as the
reaction proceeded, the competitive adsorption between
the dechlorinated product and 2,4-D also reduced the
ity of BC-nPd/Fe to 2,4-D. In summary, 25.0 C was the
optimal reaction temperature recorded in the present
study.
3.3 | Dechlorination process and kinetics
of 2,4-D by nPd/Fe and BC-nPd/Fe
Figure 11 shows the progress of 2,4-D dechlorination by
nPd/Fe and BC-nPd/Fe. In the nPd/Fe system
(Figure 11a), 2,4-D removal rate was 44.1%, and PA gen-
eration rate was 20.1% at a reaction time of 200 min. At
the same reaction time, 2,4-D was completely removed
by BC-nPd/Fe, whereas 92.1% of PA was generated
[41,44]
reaction rate.
3
.2.5 | Effect of reaction temperature
(
Figure 11b). The experimental results showed that BC
The effects of reaction temperature on the dechlorination
of 2,4-D were investigated at 20.0 C, 25.0 C, 30.0 C,
loading significantly promoted the dechlorination of
2,4-D by nPd/Fe. In both systems, the 2,4-D dechlorina-
tion products were 2-CPA and PA, but the intermediate
4-CPA was not detected. During the entire dechlorination
process, almost all 2,4-D was first converted to 2-CPA
(instead of 4-CPA) and then reduced to PA. The main
reason for this may be the steric effect of the oxygen-
containing functional group in the 2,4-D molecule, mak-
ing the para-chlorine atom more easily replaced by
ꢁ
ꢁ
ꢁ
ꢁ ꢁ
3
5.0 C, and 40.0 C, as shown in Figure 10. When the
ꢁ
ꢁ
reaction temperature increased from 20.0 C to 25.0 C,
,4-D complete removal rates (100.0%) were observed,
2
and the PA generation increased from 89.1% to 92.1%.
Increasing the temperature accelerated the molecular
movement, which was conducive to the diffusion of
2
,4-D to the surface of BC-nPd/Fe, but the enhancement
[
22]
of the 2,4-D dechlorination reaction was limited to the
above conditions. As the temperature continued to
increase, significant reductions in both 2,4-D removal
hydrogen.
It can be observed that the concentration of
2-CPA reaches a peak within 10–20 min and then
[
23]
decreases slowly (Figure 11b). Zhou et al.
also pointed

12 of 15
ZHOU ET AL.
FIGURE 10 Effect of reaction temperature on 2,4-D dechlorination by BC-nPd/Fe (nZVI = 1.0 g L− , C0 2,4-D = 20.0 mg L−1, pH = 7.0,
Pd/Fe = 0.5%, MBC:MFe = 1:1, n = 200 r min− ). 2,4-D, 2,4-dichlorophenoxyacetic acid; BC-nPd/Fe, biochar-supported nano-palladium/iron;
nZVI, nano-zero-valent iron
1
1
FIGURE 11 The dechlorination process of 2,4-D by (a) nPd/Fe and (b) BC-nPd/Fe and (c) the dechlorination kinetics of 2,4-D (MBC
M
:
Fe = 1:1, nZVI = 1.0 g L− , C0 2,4-D = 20.0 mg L , T = 25.0 C, pH = 7.0, Pd/Fe = 0.5%, n = 200 r min ). 2,4-D, 2,4-dichlorophenoxyacetic
1
−1
ꢁ
−1
acid; BC-nPd/Fe, biochar-supported nano-palladium/iron; nZVI, nano-zero-valent iron

ZHOU ET AL.
13 of 15
out that 2-CPA was temporarily formed before the final
dechlorination product PA was generated. The entire
dechlorination process can be represented by Equa-
tions 10, 11, and 12:
may be due to the adsorption of 2,4-D and its dechlorina-
tion products by BC or by the nonreactive sites or passiv-
[
52,56]
ation layer on the surface of the particles.
The
carbon loss in the BC-nPd/Fe system, however, did not
increase significantly (compared with nPd/Fe), as the
total carbon loss of contaminants was 2.13% (in the
nPd/Fe system) and 3.56% (in the BC-nPd/Fe system) at
a reaction time of 200 min, respectively. It can be consid-
ered that the reductive dechlorination of 2,4-D was the
main reaction process. In addition, the generation rate of
the intermediate product 2-CPA was extremely low (4.3%
in the BC-nPd/Fe system). Therefore, the concentration
changes of 2,4-D and its final product PA can objectively
reflect the dechlorination effect of BC-nPd/Fe on 2,4-D.
Figure 11c shows the kinetics of 2,4-D dechlorination,
and the data fitted well with the pseudo-first-order model
Pd
ꢂ
−
−
2
,4−D + H + e ! 2−CPA + Cl ,
ð10Þ
ð11Þ
Pd
ꢂ
−
−
2
−CPA + H + e ! PA + Cl :
Overall reaction:
,4−D + 2H + 2e ! PA + 2Cl :
Pd
ꢂ
−
−
2
ð12Þ
2
In these two systems, the total carbon from organic
contaminants showed a similar trend. The total carbon
loss from organic contaminants was approximately 2%–
(R > 0.97). The results demonstrated that the removal
rate of 2,4-D by nPd/Fe and BC-nPd/Fe was controlled
by the chemical reaction on the surface of the
nanoparticles, which was consistent with the results
6
% throughout the reaction. This small amount of loss
FIGURE 12 Mechanism of 2,4-D dechlorination by BC-nPd/Fe. 2,4-D, 2,4-dichlorophenoxyacetic acid; BC-nPd/Fe, biochar-supported
nano-palladium/iron

14 of 15
ZHOU ET AL.
0
discussed above. The kobs values were significantly differ-
ent in the nPd/Fe (0.0042 min) and BC-nPd/Fe
of BC-nPd/Fe particles retained a higher Fe (8.1%) and
0
Pd (48.3%) content after the reaction, and the generation
(
0.0578 min) systems; the latter was 13.8 times that of the
of a passivation layer was significantly suppressed, thus
maintaining the high reduction reactivity of 2,4-D by BC-
nPd/Fe during the reaction. The results also demon-
strated that a higher weight ratio of BC to Fe and Pd
loading, lower initial pH, suitable 2,4-D concentration,
and reaction temperature were advantageous for the
reductive dechlorination of 2,4-D by BC-nPd/Fe. Under
optimal reaction conditions, the 2,4-D removal and PA
generation rates increased from 44.1% and 20.1%
(nPd/Fe) to 100.0% and 92.1% (BC-nPd/Fe), and the
kinetic rate constant increased from 0.0042 to
0.0578 min, respectively. So this method might be a satis-
factory approach if applied to treat the practical wastewa-
ter containing biologically resistant chlorinated-organics.
Furthermore, the mechanism of 2,4-D removal by BC-
former. The results clearly indicated that the presence of
BC efficiently accelerated the rapid dechlorination of
2
,4-D by nPd/Fe.
3
.4 | Mechanism of 2,4-D dechlorination
by BC-nPd/Fe
Figure 12 presents the reaction mechanism of 2,4-D by
BC-nPd/Fe. During the dechlorination of 2,4-D, H pro-
duced by the corrosion of nZVI was adsorbed by Pd and
converted into H* (Equations 7, 8, and 9). Then, the chlo-
rine atoms in the 2,4-D molecule were substituted by H*
to form the intermediate 2-CPA and final product PA
2
2+
0
(Equations 10, 11, and 12). However, the generated Fe
nPd/Fe involved the corrosion of Fe , and the generated
may form a passivation layer covering the surface of the
nPd/Fe particles during the nZVI corrosion process
H , adsorbed by Pd, was converted into H*, followed by a
2
quick and efficient reduction of 2,4-D to PA because of
the increased number of active sites and electron transfer.
Thus, BC-nPd/Fe could be a promising composite mate-
rial for the effective dechlorination of 2,4-D from a water
environment.
(Equations 5 and 6), which reduces the exposure of effec-
tive reaction sites on the particle surface, whereas the
electron transfer between nZVI and contaminants is hin-
dered. Therefore, the dechlorination efficiency of 2,4-D
by nPd/Fe decreased rapidly.
When the supporting material BC was introduced,
the agglomeration phenomenon between the nPd/Fe par-
ticles was significantly suppressed. Characterization anal-
ysis of BC-nPd/Fe composites revealed that nPd/Fe
particles were dispersed and stabilized on the BC surface.
Compared with nPd/Fe, the surface of the BC-nPd/Fe
particles possessed more reactive sites and a larger con-
tact area with the pollutant 2,4-D. In addition, BC
inhibited the formation of a passivation layer on the
nPd/Fe surface during the reaction (proved by XRD and
ACKNOWLEDGMENT
The authors would like to acknowledge for the financial
support from the National Natural Science Foundation of
China (CN) [project no. 21477108].
AUTHOR CONTRIBUTIONS
Hongyi Zhou: Funding acquisition; project administra-
tion. ning huang: Investigation; visualization.
Yongkang Zhao: Validation. Junchao Xiang: Data
curation.
XPS), which was beneficial to nZVI corrosion, H dissoci-
2
ation, and electron transfer, thereby improving the effi-
ciency and capacity to reduce 2,4-D by BC-nPd/Fe.
DATA AVAILABILITY STATEMENT
The data that support the findings of this study are avail-
able on request from the corresponding author. The data
are not publicly available due to privacy or ethical
restrictions.
4
| CONCLUSION
ORCID
In summary, BC prepared from peanut shells was found
to be a good supporting multifunctional material for
removing 2,4-D from water, and the characterization ana-
lyses indicated that the BC-nPd/Fe composites were suc-
cessfully synthesized. SEM analysis revealed that nPd/Fe
particles were well dispersed and stabilized on the sur-
face of BC, and the average diameter of these particles
was 88.8 nm. BET and VSM analyses indicated that the
BC-nPd/Fe composites have a larger specific surface area
and weaker magnetic properties as compared with
nPd/Fe. XRD and XPS analyses showed that the surface
Hongyi Zhou https://orcid.org/0000-0001-5721-725X
Ning Huang https://orcid.org/0000-0002-4639-2041
REFERENCES
[1] Z. M. Yang, X. H. Xu, M. Dai, L. Wang, X. S. Shi, R. B. Guo,
J. Hazard. Mater. 2018, 353, 490.
[
2] F. Islam, J. Wang, M. A. Farooq, M. S. S. Khan, L. Xu,
J. W. Zhu, M. Zhao, S. Muños, Q. X. Li, W. J. Zhou, Environ.
Int. 2018, 111, 332.
[3] N. Jaafarzadeh, F. Ghanbari, M. Ahmadi, Chemosphere 2017,
169, 568.

ZHOU ET AL.
15 of 15
[
4] E. I. Seck, J. M. Doña-Rodríguez, C. Fernández-Rodríguez,
O. M. González-Díaz, J. Araña, J. Pérez-Peña, Appl. Catal. B-
Environ. 2012, 125, 28.
5] M. A. Lemus, T. López, S. Recillas, D. M. Frías, M. Montes,
J. J. Delgado, M. A. Centeno, J. A. Odriozola, J. Mol. Catal. A-
Chem. 2008, 281, 107.
[34] B. Kakavandi, R. R. Kalantary, M. Farzadkia, J. Environ.
Health Sci. 2014, 12, 115.
[35] S. Nivedita, G. Vatsana, R. A. Singh, G. V. Kumar, S. Sarita,
R. P. Kumar, J. Environ. Chem. Eng. 2018, 6, 5196.
[36] Z. Li, S. Q. Luo, Y. Yang, J. W. Chen, Chemosphere 2019,
216, 499.
[
[
6] A. M. Smith, M. T. Smith, M. A. L. Merrill, J. Liaw,
C. Steinmaus, Ann. Epidemiol. 2017, 27, 281.
[37] G. Sheng, A. Alsaedi, W. Shammakh, S. Monaquel, J. Sheng,
X. Wang, H. Li, Y. Huang, Carbon 2016, 99, 123.
[38] X. S. Lv, X. Q. Xue, G. M. Jiang, D. L. Wu, T. T. Sheng,
H. Y. Zhou, X. H. Xu, J. Colloid Interface Sci. 2014, 417, 51.
[39] J. C. Yan, L. Han, W. G. Gao, S. Xue, M. F. Chen, Bioresour.
Technol. 2015, 175, 269.
[
7] D. Loomis, K. Guyton, Y. Grosse, F. E. Ghissasi, V. Bouvard,
L. Benbrahim-Tallaa, N. Guha, H. Mattock, K. Straif, Lancet
Oncol. 2015, 16, 891.
[
[
8] K. Shareef, G. Shaw, Chemosphere 2008, 72, 8.
9] S. González-Cuna, J. Galíndez-Mayer, N. Ruiz-Ordaz,
S. Murugesan, Int. Biodeterior. Biodegrad. 2016, 107, 123.
[40] R. K. Xu, S. C. Xiao, J. H. Yuan, A. Z. Zhao, Bioresour. Technol.
2011, 102, 10293.
[
[
[
10] Z. R. Sun, X. F. Wei, Y. B. Han, S. Tong, X. Hu, J. Hazard.
[41] K. Weber, P. Quicker, Fuel 2018, 217, 240.
[42] S. S. Wang, M. Y. Zhao, M. Zhou, Y. C. C. Li, J. Wang, B. Gao,
S. Sato, K. Feng, W. Q. Yin, A. D. Igalavithana, P. Oleszczuk,
X. Z. Wang, Y. S. Ok, J. Hazard. Mater. 2019, 373, 820.
[43] E. J. Cole, O. R. Zandvakili, B. S. Xing, M. Hashemi,
S. Herbert, H. H. Mashayekhi, Data Brief 2019, 25, 104073.
[44] L. B. Qian, W. Y. Zhang, J. C. Yan, L. Han, Y. Chen,
D. Ouyang, M. F. Chen, Environ. Pollut. 2017, 223, 153.
[45] P. Devi, A. K. Saroha, Chem. Eng. J. 2015, 271, 195.
[46] H. Choi, S. Agarwal, S. R. Al-Abed, Environ. Sci. Technol.
2009, 43, 488.
Mater. 2013, 244-245, 287.
11] H. Chen, Z. L. Zhang, Z. L. Yang, Q. Yang, B. Li, Z. Y. Bai,
Chem. Eng. J. 2015, 273, 481.
12] Y. J. Xiang, Z. Y. Xu, Y. Y. Zhou, Y. Y. Wei, X. Y. Long,
Y. Z. He, D. Zhi, J. Yang, L. Luo, Chemosphere 2019, 237,
1
24464.
[
[
[
13] Y. J. Xiang, X. Yang, Z. Y. Xu, W. Y. Hu, Y. Y. Zhou,
Z. H. Wan, Y. H. Yang, Y. Y. Wei, J. Yang, D. C. W. Tsang, Sci.
Total Environ. 2020, 709, 136079.
14] F. B. Yao, Q. Yang, Y. Zhong, X. Y. Shu, F. Chen, J. Sun,
Y. H. Ma, Z. Y. Fu, D. B. Wang, X. M. Li, Water Res. 2019,
[47] H. Li, Y. F. Qiu, X. L. Wang, J. Yang, Y. J. Yu, Y. Q. Chen,
Y. D. Liu, Chemosphere 2017, 169, 534.
[48] L. Han, J. C. Yan, L. B. Qian, W. Y. Zhang, M. F. Chen,
J. Environ. Manage. 2019, 245, 238.
1
57, 191.
15] X. Q. Li, D. W. Elliott, W. X. Zhang, Crit. Rev. Solid State
Mater. Sci. 2006, 31, 111.
[
[
16] W. Zhang, C. B. Wang, H. L. Lien, Catal. Today 1998, 40, 387.
17] Y. J. Wu, J. H. Zhang, Y. F. Tong, X. H. Xu, J. Hazard. Mater.
[49] X. Shang, L. Yang, D. Ouyang, B. Zhang, W. Y. Zhang,
M. Y. Gu, J. Li, M. F. Chen, L. H. Huang, L. B. Qian, Chemo-
sphere 2020, 249, 126518.
[50] D. Bhaduri, A. Saha, D. Desai, H. N. Meena, Chemosphere
2016, 148, 86.
2
009, 172, 1640.
18] J. Fan, Y. Guo, J. Wang, M. Fan, J. Hazard. Mater. 2009,
66, 904.
[
1
[
[
19] G. C. C. Yang, H. L. Lee, Water Res. 2005, 39, 884.
20] Z. Q. Fang, X. H. Qiu, J. H. Chen, X. Q. Qiu, Desalination
[51] H. Y. Zhou, J. C. Xiang, Y. K. Zhao, Y. Chen, Sep. Purif.
Technol. 2018, 207, 377.
2
011, 267, 34.
[52] P. P. Wang, X. G. Liu, B. C. Yu, X. H. Wu, J. Xu, F. S. Dong,
Y. Q. Zheng, Sci. Total Environ. 2020, 702, 134767.
[53] N. S. Portillo-Vélez, R. Zanella, Chem. Eng. J. 2020, 385,
123848.
[54] J. H. Ramirez, F. J. Maldonado-Hódar, A. F. Pérez-Cadenas,
C. Moreno-Castilla, C. A. Costa, L. M. Madeira, Appl. Catal.
Environ. 2007, 75, 312.
[
[
[
[
[
21] S. Y. Oh, Y. D. Seo, K. S. Ryu, D. J. Park, S. H. Lee, Environ Sci
Process Impacts 2017, 19, 711.
22] Y. Zhuang, S. Ahn, A. L. Seyfferth, Y. Masue-Slowey,
S. Fendorf, R. G. Luthy, Environ. Sci. Technol. 2011, 45, 4896.
23] H. Y. Zhou, J. Han, S. A. Baig, X. H. Xu, J. Hazard. Mater.
2
011, 198, 7.
24] Z. Q. Fang, X. H. Qiu, J. H. Chen, X. Q. Qiu, J. Hazard. Mater.
011, 185, 958.
[55] X. Wang, C. Chao, C. Ying, H. Liu, J. Hazard. Mater. 2009,
161, 815.
2
25] Y. J. Yu, Z. Huang, D. Y. Deng, Y. M. Ju, L. Y. Ren,
M. D. Xiang, L. Z. Li, H. Li, Chem. Eng. J. 2017, 325, 279.
26] Y. Xu, W. X. Zhang, Ind. Eng. Chem. Res. 2000, 39, 2238.
27] D. O'Carroll, B. Sleep, M. Krol, H. Boparai, C. Kocur, Adv.
Water Resour. 2013, 51, 104.
[56] Y. J. Shi, H. D. Hu, H. Q. Ren, Bioresour. Technol. 2019, 297,
122281.
[
[
[
[
[
[
28] B. W. Zhu, T. T. Lim, Environ. Sci. Technol. 2007, 41, 7523.
29] C. B. Wang, W. X. Zhang, Environ. Sci. Technol. 1997, 31, 2154.
30] H. L. Lien, W. X. Zhang, Appl. Catal. Environ. 2007, 77, 110.
31] G. Huang, M. Wang, Y. Hu, S. Lv, C. Li, PLoS ONE 2017, 12,
e0174589.
How to cite this article: Zhou H, Huang N,
Zhao Y, Baig SA, Xiang J. Dechlorination of 2,4-
dichlorophenoxyacetic acid using biochar-
supported nano-palladium/iron: Preparation,
characterization, and influencing factors. Appl
Organomet Chem. 2020;e6010. https://doi.org/10.
[
32] N. Sakulchaicharoen, D. M. O'Carroll, J. E. Herrera,
J. Contam. Hydrol. 2010, 118, 117.
1002/aoc.6010
[
33] T. Wang, J. Su, X. Y. Jin, Z. L. Chen, J. Hazard. Mater. 2013,
2
62, 819.