Environmental Pollution 221 (2017) 470e479

Contents lists available at ScienceDirect

Environmental Pollution
journal homepage: www.elsevier.com/locate/envpol

Co-transport of chlordecone and sulfadiazine in the presence offunctionalized multi-walled carbon nanotubes in soils* 
Miaoyue Zhang a, e, *, Irina Engelhardt a, b, Jirka Sim
unek c, Scott A. Bradford d,Daniela Kasel a, Anne E. Berns a, Harry Vereecken a, Erwin Klumpp aaAgrosphere Institute (IBG-3), Forschungszentrum Jülich GmbH, 52425 Jülich, Germany
TU Freiberg, Department of Hydrogeology, 09596 Freiberg, Germany
Department of Environmental Sciences, University of California Riverside, Riverside, CA 92521, USAd
United States Department of Agriculture, Agricultural Research Service, U. S. Salinity Laboratory, Riverside, CA 92507, USAe
Institute for Environmental Research (Biology V), RWTH Aachen University, Worringerweg 1, 52074 Aachen, Germanybca r t i c l e i n f o

a b s t r a c t

Article history:
Received 22 June 2016
Received in revised form
6 November 2016
Accepted 12 December 2016
Available online 21 December 2016

Batch and saturated soil column experiments were conducted to investigate sorption and mobility of two
14
C-labeled contaminants, the hydrophobic chlordecone (CLD) and the sulfadiazine (SDZ), in the absenceor presence of functionalized multi-walled carbon nanotubes (MWCNTs). The transport behaviors of CLD,
SDZ, and MWCNTs were studied at environmentally relevant concentrations (0.1e10 mg L1) and they
were applied in the column studies at different times. The breakthrough curves and retention proﬁleswere simulated using a numerical model that accounted for the advective-dispersive transport of allcompounds, attachment/detachment of MWCNTs, equilibrium and kinetic sorption of contaminants, andco-transport of contaminants with MWCNTs. The experimental results indicated that the presence ofmobile MWCNTs facilitated remobilization of previously deposited CLD and its co-transport into deeper
soil layers, while retained MWCNTs enhanced SDZ deposition in the topsoil layers due to the increased
adsorption capacity of the soil. The modeling results then demonstrated that the mobility of engineered
nanoparticles (ENPs) in the environment and the high afﬁnity and entrapment of contaminants to ENPs
were the main reasons for ENP-facilitated contaminant transport. On the other hand, immobile MWCNTs
had a less signiﬁcant impact on the contaminant transport, even though they were still able to enhance
the adsorption capacity of the soil.
© 2016 Elsevier Ltd. All rights reserved.

Keywords:
Colloid-facilitated contaminant transport
Multi-walled carbon nanotubes
Soil
Retention proﬁleNumerical modeling

1. Introduction
Multi-walled carbon nanotubes (MWCNTs) consist of multiple
rolled layers (concentric tubes) of graphene (Iijima, 1991; Kasel
et al., 2013a). Due to their unique physical and chemical properties, carbon nanotubes (CNTs) have been used in various applications (Gannon et al., 2007; Gohardani et al., 2014; Mattison et al.,2011), which will undoubtedly result in their release into theenvironment. Recent studies have additionally reported thatvarious organic or inorganic contaminants such as pesticides, antibiotics, and heavy metals can sorb to and be transported by CNTs
(Josko et al., 2013; Kim et al., 2014; Luo et al., 2013; Ren et al., 2011).

*

This paper has been recommended for acceptance by Baoshan Xing.
* Corresponding author.
E-mail address: m.zhang@fz-juelich.de (M. Zhang).

http://dx.doi.org/10.1016/j.envpol.2016.12.018
0269-7491/© 2016 Elsevier Ltd. All rights reserved.

An important implication of the CNTs presence in the environmentis that they may signiﬁcantly affect the fate of various contaminants
and that they may potentially be used for their remediation.
To date, only a few studies have investigated the transport
behavior of CNTs in natural soils or in model soil systems. Thesestudies have reported that the transport behavior of CNTs isaffected by ionic strength, ﬂow velocity, structure of porous media,
their input concentrations, and their structure/shape (Jaisi et al.,2008; Kasel et al., 2013a; Mattison et al., 2011; Tian et al., 2012;
Wang et al., 2012). Kasel et al. (2013b) found only limited transport of functionalized MWCNTs in undisturbed and unsaturated
soils. Fang et al. (2013) showed that the mobility of CNTs suspended
in a nonionic surfactant solution varied with the size of soil particles and the soil's sand content, and was negatively correlated tothe soil's clay content. However, current knowledge about the fate
and transport of CNTs in soils is still rather limited. There are evenfewer studies on the effects of CNTs on the mobility of other solutes

M. Zhang et al. / Environmental Pollution 221 (2017) 470e479

and compounds, such as organic and inorganic contaminants, insoils.
In general, colloids (e.g., suspended nanoparticles, clay particles,
and metal oxides) in the subsurface can act as carriers for contaminants because of their mobility and large sorption capacity.
This process is often deﬁned as “colloid-facilitated contaminant transport” (co-transport) (Bradford and Kim, 2010; Sim
unek et al.,2006, 2012). However, only very few studies have investigatedthe co-transport between engineered nanoparticles (ENPs) andcontaminants (Hofmann and Von der Kammer, 2009; Su et al.,2016; Zhang et al., 2011). These studies reported that ENPs affectthe contaminant transport if nanoparticles have a strong mobility
in porous media, a high adsorption capability, or if they are present
in high concentrations (Hofmann and Von der Kammer, 2009;
Zhang et al., 2011). These studies used a mixture of ENPs andcontaminants. However, it is relatively unlikely that ENPs andcontaminants are simultaneously released into the environment. To
the best of our knowledge, there have been no soil columns
experimental and numerical analysis concerning the fate andtransport of CNTs and contaminants that were not released
simultaneously. Such studies could simulate a possible soil remediation scenario when CNTs are released, either on purpose or
accidentally, into an initially contaminated soil.
Numerical models simulating colloid-facilitated contaminanttransport are generally based on the mass balance equations forboth colloids and contaminants and account for various interactions between the colloids, contaminants, and soil. They
consider various equilibrium and kinetic models such as the ﬁrstorder sorption of contaminants to soil, the ﬁrst-order kinetic
attachment of colloids (Jin et al., 1997; Treumann et al., 2014),
irreversible nonlinear kinetic attachment of colloids (Bradford
et al., 2011), and competitive Langmuir kinetic sorption of contaminants on colloids (van de Weerd and Leijnse, 1997). In thisstudy, additional processes are considered in the numerical analysisof the experimental data, including blocking and depth-dependent
retention of MWCNTs during transport (Kasel et al., 2013a) and anon-simultaneous release of MWCNTs and contaminants (Pang and
 
 
Sim
unek, 2006; Sim
unek et al., 2006).
The aim of this study is to investigate the sorption and mobilityof two different contaminants chlordecone (CLD, C10Cl10O), a highly
chlorinated pesticide, and sulfadiazine (SDZ, 4-amino-N-pyrimidin-2-yl-benzenesulfonamide), a widely used sulfonamide antibiotics, in the absence and presence of functionalized MWCNTs.
CLD is a persistent organic pollutant that remains present in soils
almost 20 years after its prohibition due to its long-term application (Cabidoche and Lesueur-Jannoyer, 2012; Fernandez-Bayo et al.,2013b). The intense usage of antibiotics such as SDZ led to theirwide distribution in the environment (McArdell et al., 2003; Pailler
et al., 2009; Tamtam et al., 2008). In this study, the transport
behavior of MWCNTs in combination with the contaminants SDZ
and CLD is investigated by analyzing adsorption isotherms, as wellas breakthrough curves and retention proﬁles of column experiments. The column data are analyzed using the C-Ride module
 
 
(Sim
unek et al., 2012) of HYDRUS-1D (Simunek et al., 2016), which
accounts for advective-dispersive transport of both colloids andcontaminants, attachment, detachment and straining of colloids,
and kinetic sorption of contaminants to both soil and colloids.
2. Materials and methods
2.1. Chemicals
Radioactively (14C) labeled and unlabeled MWCNTs wereapplied for the batch and column experiments (Bayer Technology
Services GmbH, 51368 Leverkusen, Germany). MWCNTs were471

boiled with 70% nitric acid (Sigma-Aldrich Chemie GmbH, 89555
Steinheim, Germany) for 4 h, resulting in additional oxygencontaining functional groups (e.g., carboxylic groups) on theirsurfaces. Kasel et al. (2013a, 2013b) found limited transport of
MWCNT in soil and (small grain size) quartz sand even under unfavorable attachment conditions (negative charge of both MWCNTs
and porous media at low ionic strength) and suggested that thiswas due to particle straining and aggregation, implying that thecolloidal stability of the MWCNT suspension was a signiﬁcant factor. Therefore, the MWCNT suspensions were prepared with 1 mM
KCl and ultrasonicated for 15 min at 65 W by a cup horn sonicator as
stock suspension, and then ultrasonicated again for 10 min before
injection, leading to better-dispersed MWCNTs during the experiments. This MWCNTs suspension (1 mg L1, 1 mM KCl) was found to
be stable for at least 24 h after preparation (Kasel et al., 2013a). The
characterization and aggregation behavior of functionalized
MWCNTs is given in Kasel et al. (2013a, 2013b).
Radioactively (14C) labeled CLD dissolved in acetone was purchased from Moravek Biochemicals (Brea, CA, USA). The speciﬁcradioactivity of 14C-labeled CLD was 2.94 MBq mg1. For preparation of the desired CLD solution concentration, non-labeled CLD(Sigma-Aldrich Chemie GmbH, Steinheim, Germany) was mixed
with the 14C-labeled CLD. The total mass of non-labeled CLD and14
C-labeled CLD was approximately 1 mg. After the acetone had
evaporated, the ﬂasks were ﬁlled with 1 L of 1 mM KCl.
Radioactively (14C) pyrimidine-ring-labeled SDZ was purchasedfrom Bayer-Health Care AG (Wuppertal, Germany) and dissolved in
the 1 mM KCl solution for the desired concentration. The speciﬁcradioactivity was 0.43 MBq mg1. All 14C-labeled samples wereadded to 5 ml of scintillation cocktail (Insta-Gel Plus, Perkin Elmer,
USA) and measured using a liquid scintillation counter (LSC, Perkin
Elmer, USA).

2.2. Batch experiments
In accordance with OECD guideline 106 (OECD., 2000), theadsorption kinetics and the adsorption isotherms of CLD on the
loamy sand soil and on MWCNTs were determined in batch trials€llrigl-Rosta et al., 2003),
and by using the dialysis technique (Horespectively. The soil adsorption isotherms were measured using
mixtures of 10 ml 14C-labeled CLD solutions of different concentrations and 1 g soil (in 1 mM KCl), which were equilibrated for 24 hin an overhead shaker. After centrifugation, the supernatant was€ llrigl-Rosta
taken and measured by LSC. The dialysis technique (Hoet al., 2003) was applied to determine the adsorption of CLD and
SDZ on MWCNTs. Two dialysis half-cells were therefore separated
by a 1 kDa cut-off cellulose membrane and inserted into a special
frame. Then, the bottom part of the half-cell was ﬁlled withapproximately 5 ml of MWCNT suspension (non-labeled,
10 mg L1) and the top part of the half-cell was ﬁlled with 5 ml of
CLD (0e10 mg g1) or SDZ (0e1 mg L1) at selected concentrations.
The ﬁlled half-cell was then rotated at 10 rpm for 48 h. Afterequilibration, the solution concentration of 14C-labeled CLD or SDZ
in the top part of the half-cell was measured by LSC.
The sorption kinetics of CLD to MWCNTs were determined byadding 10 ml of 10 mg g1 of 14C-labeled CLD solution to 10 mg of
unlabeled MWCNTs, and then mixing. This suspension wascentrifuged at selected time intervals and a sample of the supernatant was measured by LSC. The kinetic analysis of the batch
experimental data was carried out using the following equation(Schijven and Hassanizadeh, 2000):

472

M. Zhang et al. / Environmental Pollution 221 (2017) 470e479

Ckþ kamM exp½  ðkamM þ kdmM Þt
¼ dmM
C0
kamM þ kdmM

(1)where C is the contaminant concentration in the liquid phase[ML3]; C0 is the initial contaminant concentration in the liquid
phase [ML3]; kamM and kdmM are the CLD adsorption and desorption rate coefﬁcients to the mobile MWCNTs [T1], respectively. The
calculated adsorption and desorption rate coefﬁcients (kamM andkdmM ) are given in Table 1. Blank experiments were conducted todetermine the sorption of CLD, SDZ, and MWCNTs to wall of thecentrifuge tubes or the dialysis membrane in the batch experiments. Batch concentrations in the presence of soil or MWCNTs
were corrected for sorption losses in blank experiments.
2.3. Transport experiments
Stainless steel columns (3 cm inner diameter and 12 cm length)were used for transport experiments. Samples of the loamy sandsoil were taken from the upper 30 cm in Kaldenkirchen-Hülst,Germany (sieved to a fraction < 2 mm and air dried, and the mediangrain size, d50, equaled 120 mm). This soil had a total organic carbonof 1.1% mass, a cationic exchange capacity of 7.8 cmolc kg1, and pH
value of 5.9 (Kasel et al., 2013b). The soil was composed of 4.9% clay(<2 mm), 26.7% silt (2e63 mm), and 68.5% sand (>2 mm). Soil anddeionized water were alternately and incrementally ﬁlled into thecolumns. The columns were then connected to a pump (MCP V 5.10,
Ismatec SA, Glattbrugg, Switzerland), with the ﬂow direction fromthe bottom to the top of the column. Approximately 30 pore volumes (PVs) of background electrolyte solution (1 mM KCl) wereapplied to the column before the transport experiments were carried out. The bulk density of the packed columns was approximately 1.48 g cm3.
A tracer experiment was conducted ﬁrst to characterize thehydraulic conditions and conservative transport parameters, suchas porosity and dispersivity, of the soil. A pulse of approximately
2.1 PV (about 90 mL) of the tracer (1 mM KBr) was applied for that
purpose. Efﬂuent solutions were collected by a fraction collector
every 30 s (e.g., approximately 2.5 mL per vial) and analyzed todetermine the breakthrough curves (BTCs). The efﬂuent concentrations of bromide were determined using a high-performanceliquid chromatograph (STH 585, Dionex, Sunnyvale, CA, USA)equipped with a UV detector (UV2075, Jasco, Essex, UK).
The same procedure was repeated when CLD, SDZ, and MWCNTs
were applied. The single-species transport of CLD (14C-labeled,experiment I), SDZ (14C-labeled, experiment II and III), or MWCNTs(14C-labeled, experiment IV) in the saturated soil columns were
investigated ﬁrst. Based on information collected during this ﬁrst
step, the second set of column experiments was designed toinvestigate the effect of MWCNTs on the transport of CLD and SDZ.
During the CLD co-transport experiment, a pulse (around 2.1 PV) of
CLD (14C-labeled) was injected ﬁrst, followed by a MWCNT suspension pulse (non-labeled, around 2.1 PV) applied at the same

CLD
CLD
SDZ
SDZa

Absorbents

MWCNT
soil
MWCNT
soil

2.4. Numerical modeling 
The C-Ride module (Simunek et al., 2012) of HYDRUS-1D (Simunek et al., 2016) was used to simulate the one-dimensionaltransport and co-transport of CLD, SDZ, and MWCNTs. The transport of MWCNTs can be described using the advection-dispersion
equation coupled with one-site kinetic retention as follows(Bradford et al., 2003; Gargiulo et al., 2007):

vCM
vS
v2 C MvC
þ r M ¼ qDM
q M
vt
vt
vx
vx2

(2)

vSM
¼ qjkac CM  rkdc SM
vt

(3)

q
r

where q is the volumetric water content [L3L3], CM is the MWCNT
concentration in the liquid phase [nL3], t is time [T], r is the bulk
density of the soil [ML3], SM is the MWCNT concentration associated with (attached or strained) the soil [nM1], x is the spatial
coordinate [L], DM is the hydrodynamic dispersion coefﬁcient for
MWCNTs [L2T1], q is the Darcy velocity [LT1], kac and kdc are theﬁrst-order attachment and detachment coefﬁcients for MWCNTs[T1], respectively, and j [-] is the dimensionless MWCNT retentionfunction to account for time- and depth-dependent retention,
which is deﬁned as:Table 1
Optimized results from batch experiments.
Chemicals

ionic strength and ﬂow velocity (experiment V and VI). In contrast,
during the SDZ co-transport experiment, the MWCNT suspension(non-labeled, around 2.1 PV) was injected ﬁrst, followed by an injection of SDZ (14C-labeled, around 2.1 PV, experiments VII and
VIII). For both CLD and SDZ co-transport experiments, the non-14Clabeled MWCNTs, prepared in the same way as the 14C-labeled
MWCNTs, were applied. We assumed that the non-labeled
MWCNTs in the co-transport experiments showed the sametransport behavior as the 14C-labeled MWCNTs in the singlespecies transport experiments. A summary of the experimental
conditions is provided in Table 2. The retardation factor (R) can becalculated from the mean breakthrough time (tb), the column
length (L), and the pore-water velocity (v) as R¼tbv/L. The value of R
is a measure of the mean breakthrough time relative to a conservative solute tracer.
At the end of the experiments, soil samples were excavated fromthe columns in approximately 0.5e1 cm thick increments andanalyzed to determine the retention proﬁles (RPs) of CLD, SDZ, and
MWCNTs. The soil samples of each layer were dried, crushed andcombusted using a biological oxidizer at 900  C (OX 500, R.J. Harvey
Instrumentation Corporation, Tappan, NY, USA), followed by the
LSC measurement. Before the column experiments, empty columns
without soil were used to determine losses of MWCNTs, CLD, andSDZ to ensure the appropriateness of the column setup. The loss
percentage to the empty column was approximately 3%.

SSMM
j ¼ 1  max

kamM

kdmM

Kd

Koc

[min1][min1][cm3g1][cm3g1]

0.10

5.32E-03

60,000
56
6000
0.56

NF
5090
NF
5.09

NF - not ﬁtted.
a
The adsorption of SDZ in the soil was described by Zarﬂ (2008).



d50 þ x
d50

b
(4)

where d50 is the median grain size of the soil [L], b [-] is a parameter
which is controlled by the shape of the retention proﬁle, Smax
M is themaximum solid phase MWCNT concentration [nM1]. In this study,a b value of 0.765 was employed for non-spherical MWCNTs (Kasel
et al., 2013a, 2013b).
Contaminant (CLD or SDZ) transport in soil was described using
the advection-dispersion equation:

M. Zhang et al. / Environmental Pollution 221 (2017) 470e479

473

Table 2
Experimental conditions, hydraulic parameters and mass balance information for all column experiments. The ionic strength for all column experiments was 1 mM KCl and d50
of soil was 120 mm.
Single-species transport experiments
Chemical
CLD
SDZ
SDZ
MWCNTs

No.
I
II
III
IV

Co
[mg L1 ]q[cm min1]

Porosity

1
0.1
1
1

0.71
0.70
0.72
0.72

0.5
0.52
0.49
0.50

l

Meff

Msoil

Mtotal

0.007
0.937
0.981
0.458

1.002
0.118
0.030
0.488

1.009
1.056
1.011
0.946

[cm]
0.16
0.18
0.31
0.60

Co-transport experiments
Chemical

CLD/MWCNTs
CLD/MWCNTs
MWCNTs/SDZ
MWCNTs/SDZ

No.

V
VI
VII
VIII

Co
[mg L1 ]MWCNTs

CLD/SDZ

1
10
1
1

1
1
1
0.1

q
[cm min1]

Porosity

0.71
0.71
0.70
0.69

0.51
0.54
0.50
0.52

l

Meff

Msoil

Mtotal

0.010
0.010
0.976
0.884

1.008
0.997
0.073
0.280

1.017
1.007
1.049
1.164

[cm]

0.16
NF
0.09
0.21

Co is the input concentration; l is the hydrodynamic longitudinal dispersivity estimated from Br; Meff is the fraction of MWCNT, CLD, or SDZ mass in the efﬂuent; Msoil is thefraction of MWCNT, CLD, or SDZ mass in the solid phase; Mtotal is the total relative mass for the column experiment; NF - not ﬁtted.

q

vC
vS
v2 C
vC
þ r ¼ qD 2  q
vt
vt
vx
vx

(5)

where C is the contaminant concentration in the liquid phase[ML3], D is the hydrodynamic dispersion coefﬁcient for CLD or SDZ[L2T1], S is the contaminant concentration sorbed to the solid
phase [MM1]. Sorption of CLD or SDZ was described using a twosite adsorption-desorption model as follows (Van Genuchten andWagenet, 1989):

S ¼ Se þ Sk

(6)

vSe
vC
¼ fKd
vt
vt

(7)

vSk
¼ u½ð1  f ÞKd C  Sk 
vt

(8)where Se [MM1] is the contaminant concentration on instantaneous equilibrium sorption site, Sk [MM1] is the contaminantconcentration on the remaining ﬁrst-order kinetic sorption site, f [-]is the fraction of exchange sites assumed to be in equilibrium withthe solution phase, Kd is the partition coefﬁcient for linear
adsorption [L3M1], u is the ﬁrst-order rate constant [T1]. The
diffusion coefﬁcients of CLD and SDZ are from Pritchard et al. (1986)
and Chen et al. (2013), respectively. Due to the short duration of theexperiments, degradation and transformation of contaminants canbe neglected.
The interactions between the contaminant (CLD or SDZ) andmobile or immobile MWCNTs during coupled nanotubes and 
contaminant transport are described as follows (Pang and Sim
unek,
 
2006; Simunek et al., 2006, 2012):

q

vC
vSe
vS
vC S
vS S
þr
þ r k þ q M mM þ r M iM
vt
vt
vt
vt
vt
v2 C
vC
v2 CM SmM
vC S
þ þqD
¼ qD 2  q
 q M mM
vx
vx
vx
vx2

(9)

where SmM and SiM [MM1] are the CLD or SDZ concentrations
sorbed to mobile and immobile MWCNTs, respectively. Concentrations of CLD or SDZ sorbed to immobile and mobile colloids canbe written as follows:qr

vCM SmM
v2 SmM CM
vC S¼ qD
 q M mM þ qjm kamM C
vt
vx
vx2
 qkdmM CM SmM  qkac CM SmM þ rkdc SM SiM(10)vSM SiM¼ qjim kaiM C  rkdiM SM SiM þ qkac CM SmM  rkdc SM SiM
vt(11)

where kaiM and kdiM are the rate coefﬁcients for CLD or SDZ sorption
to and desorption from immobile MWCNTs [T1], respectively; andjm and jim are dimensionless variables that adjust the sorption rate
to a number of mobile and immobile MWCNTs present, respectively. Detailed list of all transport and reaction equations involvedin colloid-facilitated solute transport are described in detail in the
manual of C-Ride.

3. Results and discussion
3.1. Adsorption of CLD and SDZ on MWCNTs and soil
All the optimized results from batch experiments are summarized in Table 1. A linear adsorption isotherm provided a good
description of CLD adsorption on the soil (r2 ¼ 0.99, Table 1, Fig. 1a).
The sorption coefﬁcient (Kd) equaled 56 cm3 g1. This value of Kd
can be used in conjunction with the soil organic carbon fraction(foc ¼ 0.011 g g1) (Kasel et al., 2013b) to calculate the organiccarbon normalized adsorption coefﬁcient (Koc ¼ Kd/foc) (OECD.,
2000) that is equal to cm3 g-1. This Koc value is consistent withthe reported range in the literature of 2500 to 20,000 cm3 g-1(Woignier et al., 2012) and indicates that CLD sorption on the soil ismainly controlled by hydrophobic interactions with soil organiccarbon (Cabidoche et al., 2009; Levillain et al., 2012; Li et al., 2012),
while inorganic soil components like clay minerals play a relatively
minor role, which produces a limited mobility of CLD in soils(Fernandez-Bayo et al., 2013a; 2013b). The sorption of SDZ in thesame (Kaldenkirchen-Hülst) soil was described in an earlier study(Zarﬂ, 2008). Reported values of Kd ¼ 0.56 cm3 g-1 andKoc ¼ 5.09 cm3 g-1 indicate that organic carbon and inorganic soil
components may both play an important role for SDZ sorption on
the soil (Kasteel et al., 2010). However, SDZ has a much lower Koc
value than CLD, its mobility in soils is therefore expected to be474

M. Zhang et al. / Environmental Pollution 221 (2017) 470e479

Fig. 1. (a) Adsorption isotherm of CLD on soil. (b) Adsorption isotherm of CLD on MWCNTs. (c) Adsorption isotherm of SDZ on MWCNTs. (d) Kinetic sorption data for CLD on
MWCNTs.

Table 3
Parameters optimized by using different models for the CLD, SDZ, and MWCNT transport, co-transport and retention in soil.
Single-species transport experiments
Experiment
No.

r2

MWCNT parameters
/C0
Smax
M
3 1
[cm g

I
II
III
IV

0.827
0.963
0.997
0.975

Contaminant (CLD or SDZ) parameterskac[min1]]kdc[min1]

Frac. (f)5.46E-04
0
0

5.45

10.3

Kd
[cm3g1]u282.600.17
0.17

8.28E-04
1.22E-02
2.42[min1]

6.80E-03

Co-transport experiments
Experiment
No.

r2

V
VII
VIII

0.872
0.982
0.982

r correlation of observed and ﬁtted data.
a- obtained from the batch experiments.

Contaminant (CLD or SDZ) parameterskaiM[min1]kdiM[min1]

0.10a
1.31E-02
0.38

5.32E-03a
1.02E-02
0.49

M. Zhang et al. / Environmental Pollution 221 (2017) 470e479

475

the loamy sand soil, respectively. This strong adsorption capacity oflarge surface area MWCNTs for CLD and SDZ has been mainly been
attributed to p-p dispersion and hydrogen-bonding interactions(Liao et al., 2008; Upadhyayula et al., 2009). The presence of
MWCNTs in soils is therefore expected to strongly inﬂuence themobility of CLD and SDZ.
Equation (1) provided a good description for the adsorption
kinetics of CLD on MWCNTs (r2 ¼ 0.99, Fig. 1d). The adsorption(kamM) and desorption (kdmM) rate coefﬁcients of CLD to/from the
MWCNTs were determined to be 6.26 h1 and 0.32 h1, respectively. The adsorption rate was much faster than the desorption
rate.
3.2. Single-species transport of CLD, SDZ, and MWCNTs in soil

Fig. 2. Observed and estimated BTCs (a and b) and RPs (c) for a single-species transport of CLD (experiment I), SDZ (experiment II), and MWCNTs (experiment IV) insaturated soil columns. Figure (2a) has two vertical axes. The left axis shows therelative CLD concentrations (C/C0), while the right axis shows the relative SDZ and
MWCNTs concentrations. The breakthrough curves in Fig. b are plotted as log scale
efﬂuent concentrations (log10C/Co) versus pore volumes.

much higher than that of CLD.
The Freundlich equation provided a good description ofadsorption isotherms of CLD on MWCNTs (r2 ¼ 0.95), with theFreundlich coefﬁcient (KF) and exponent (n) equal to 229 mg(11/n)
L1/n g1 and 1.49, respectively (Fig. 1b). The sorption coefﬁcient Kd
was calculated from a linear ﬁt of the ﬁrst part of the isotherm to be
60,000 cm3 g-1, which demonstrates strong adsorption. A linear
adsorption isotherm provided a good description of SDZ adsorption
on the MWCNTs (r2 ¼ 0.99, Fig. 1c), with a sorption coefﬁcient Kd of
6000 cm3 g-1, which is in agreement with reported Kd values(103e104 cm3g1) (Ji et al., 2009). The Kd values of CLD and SDZ on
MWCNTs were three and ﬁve orders of magnitude higher than on

The ﬁrst three column experiments (I, II, and IV) were performed to improve our understanding of the single-species transport behavior of CLD, SDZ, and MWCNTs in soils. The experimental
conditions are presented in Table 2 and optimized model parameters in Table 3. The BTCs (Fig. 2a) are plotted as normalized efﬂuent
concentrations (C/C0) versus pore volumes, where C0 is the inﬂuent
concentration. Efﬂuent concentrations for CLD, SDZ, and MWCNTs
were plotted on a log scale (log10 C/Co) in Fig. 2b versus pore volumes. The RPs (Fig. 2c) are plotted as normalized solid phase
concentrations (S/C0) versus column depth. The total recovered
mass from the efﬂuent and retained in the column ranged from0.95 to 1.05 for the three compounds. The retardation factors for
CLD, SDZ, and MWCNTs that were calculated from respective BTCs
relative to the conservative tracer transport (Ptak et al., 2004)
indicated that CLD, SDZ, and MWCNTs reached the column outlet
with only a slight retardation compared to the conservative tracer(data not shown). Thus, pore size exclusion was not observed(Bradford et al., 2003).
Mass balance information in Table 2 shows that about half of theinjected MWCNTs were retained in the soil column (Fig. 2, Table 2).
Equations (2)e(5) provided a good description of the BTC and RP(r2 ¼ 0.975). The ﬁtted value of kac was considerably (several ordersof magnitude) higher than the kdc (Table 3), indicating only a slow
rate of release. The BTC exhibited blocking behavior (a decreasingrate of retention with continued MWCNT injection). The ﬁtted
value of Smax/Co ¼ 5.45 cm3 g-1 was used to account for this blocking
process, and its low value indicates that only a small fraction of thesolid surface area contributed to MWCNT retention (Bradford et al.,2009; Kasel et al., 2013a). This result is expected because the soil
and functionalized MWCNTs exhibit a net negative charge under
low ionic strength conditions that produces an energy barrier toattachment (Bradford et al. 2006, 2009). The RP shape was hyperexponential (e.g., a greater rate of retention near the column inlet
than the outlet). A number of potential explanations for hyperexponential RPs have appeared in the literature, including: straining (Bradford et al. 2002, 2003), heterogeneity in colloid size andcharge (Bolster et al., 1999; Tong and Johnson, 2007; Tufenkji andElimelech, 2005), and system hydrodynamics (Bradford et al.,2009; Wang et al., 2011). The exact reason of the hyperexponential RP cannot be deduced from the experimental data,but other literature information for the retention of MWCNTs
suggests that straining was the domain process under our experimental conditions (Jaisi et al., 2008; Kasel et al., 2013a; Wang et al.,2012).
CLD and SDZ exhibited very different mobilities in the soil.
Almost 100% of the CLD was retained in the soil, mostly in theshallow layers (0e2 cm). On the other hand, SDZ easily passed
through the column and, only 12% was retained in the soil (Table 2).
These results are consistent with the batch experiments, whichfound that the Kd value was much larger for CLD (56 cm3 g-1) than476

M. Zhang et al. / Environmental Pollution 221 (2017) 470e479

Fig. 3. Observed and estimated BTCs (a and c) and RPs (b and d) for CLD in single-species (experiment I) and co-transport (experiment V and VI, respectively) experiments insaturated soil columns. The input concentration for CLD and MWCNTs was 1 mg L1. Figure (3a) has two vertical axes. The left axis shows the relative CLD concentrations, while theright axis shows the relative MWCNT concentrations. The vertical dotted lines indicate applications of CLD and MWCNTs.

for SDZ (0.56 cm3 g-1). The two-site sorption model ﬁtted well theobserved BTCs for both CLD and SDZ (Table 3). Values of Kd thatwere ﬁtted to the column experiments were not the same as thoseobtained in the batch experiments, especially for CLD. One plausible explanation is because continuous advection in the column
systems produces kinetic sorption on the soil surface due to theshorter solute residence time than the equilibrium batch system.
Indeed, the ﬁtted fraction of equilibrium exchange sites (f) was
always very small (Table 3), indicating that sorption of both CLD
and SDZ was mainly a kinetic process. Equilibrium sorption can beneglected for SDZ due to the small value of Kd (0.17 cm3 g-1).
Therefore, only a one-site kinetic sorption process was consideredin the model by setting f ¼ 0. In contrast, equilibrium sorption
cannot be neglected for CLD because of the large value of Kd(282.6 cm3 g-1) that provides retardation of about 1.5. Unold et al.
(2009) and Wehrhan et al. (2007) came to a similar conclusion
for CLD and SDZ, even though they employed different sorption
models. Fig. 2b demonstrates that low amounts of concentration
tailing occurred following recovery of the breakthrough curves.
Detachment/desorption was greatest for MWCNTs, SDZ, and then
CLD.
Although the sorption model described experimental BTCs well,it was not able to correctly ﬁt RPs with their hyper-exponential
shapes (Fig. 2). The same problem was also encountered by Unoldet al. (2009) and Wehrhan et al. (2007). The most plausible

reason for the hyper-exponential RPs for CLD and SDZ is due tosystem hydrodynamics (Li et al., 2005). In particular, the same kinetic retention rate constant will produce greater amounts of CLD
and SDZ removal in lower than higher velocity regions of theporous medium. The transport of CLD and SDZ is therefore expected to be controlled by high velocity regions with less removalwith increasing distance.
3.3. Co-transport
Different injection sequences were employed in the cotransport experiments for CLD and SDZ in the presence of
MWCNTs. Since CLD has a limited mobility in soil (Fig. 2), the CLD
co-transport experiment was conducted by ﬁrst injecting CLD into
the soil column followed by a MWCNT suspension. This sequence
was meant to simulate the process of using MWCNTs to remediatesoil contaminated with CLD. On the other hand, MWCNTs retainedin the soil inﬂuence the soil adsorption capacity and may thus affectthe transport and sorption of highly mobile contaminants such as
SDZ. The SDZ co-transport experiment was thus performed byinjecting the MWCNT suspension ﬁrst, followed by a SDZ pulse. Thissequence was meant to simulate the process of using retained
MWCNTs in the soil to slow down the spread of a contaminant, as iscommonly done using reactive barriers.
The colloid-facilitated contaminant transport model in the C-

M. Zhang et al. / Environmental Pollution 221 (2017) 470e479

477

Fig. 4. Observed and estimated BTCs (a and c) and RPs (b and d) for single species SDZ (experiments II and III) and MWCNTs (experiment IV) or co-transport (experiment VII and
VIII) experiments in saturated soil columns. The input concentration for SDZ and MWCNT was 0.1, 1 and 1 mg L1, respectively. The vertical dotted lines indicate applications of SDZ
and MWCNTs.

Ride module of HYDRUS-1D was used to describe the interactions
between CLD and SDZ with soil, and with mobile and retained
MWCNTs in the four co-transport experiments (V, VI, VII and VIII,
respectively). The interactions between CLD or SDZ (experiment I
and II) and soil (u, f, and Kd) were described using equations (5)through (8). The interactions between MWCNTs (experiment III)and soil (Smax, kac and kdc) were described using equations (2)through (4). The interactions between CLD and mobile orretained MWCNTs were described by parameters kamM, kdmM, kaiM,and kdiM using equations (9) through (11). The interactions between
SDZ and retained MWCNTs were described by parameters kaiM andkdiM using equations (9) through (11). As discussed above, rate
parameters kamM (6.26 h1 ¼ 0.10 min1) and kdmM (0.32h1 ¼ 5.32E-3 min1) were estimated from the sorption kinetic
experiment (equation (1), Fig. 1d). We have assumed that adsorption or desorption rates of CLD to/from mobile and retained
MWCNTs were the same and used these values (kaiM ¼ kamM ¼ 0.10
min1 and kdiM ¼ kdmM ¼ 5.32E-3 min1).
3.3.1. Pulse of CLD followed by pulse of MWCNTs
In experiments I, V and VI the effect of co-transport of CLD with
MWCNTs under saturated conditions was investigated (Fig. 3). It isnoteworthy that even though CLD's mobility is limited, it still
appeared again in the outﬂow after the application of the MWCNTs

suspension (Fig. 3a). The concentrations of retained CLD decreased
approximately 11% in the ﬁrst layer (point 1 and 10, Fig. 3b) andincreased nearly 8% in the fourth layer (point 4 and 40 , Fig. 3b) due
to the MWCNT injection. Overall, nearly 17% of the CLD mass wasremobilized from the top layers and shifted into deeper layers after
injection of the MWCNT suspension (experiment I and V). This shiftdemonstrated that transport of MWCNTs facilitated the remobilization of CLD due to its strong sorption capacity for CLD. A similar
trend was observed when CLD was injected prior to being eluted
with a MWCNTs suspension at a higher input concentration of
10 mg L1 (Fig. 3c and d).
Table 3 shows the optimized parameters of the CLD co-transport
experiment. For both mobile and retained MWCNTs, the adsorption
rate of CLD was much greater than desorption rate, which
demonstrated that the sorption was a slowly reversible process and
MWCNTs had a strong adsorption capacity for CLD.
3.3.2. Pulse of MWCNTs followed by pulse of SDZ
In experiments II, III, VII and VIII the inﬂuence of retained
MWCNTs on SDZ transport and sorption was investigated (Fig. 4).
While SDZ has a high mobility in soil, it was retained to a greater
extent in soil in the presence of immobilized MWCNTs. About 12%of the SDZ was retained in the soil column when it was applied
without MWCNTs (experiment II), whereas about 28% of the SDZ

478

M. Zhang et al. / Environmental Pollution 221 (2017) 470e479

was retained when it was applied after a MWCNT pulse (Fig. 4a andb, Table 2).
Mass balance information in Table 2 indicates that application of
the MWCNT pulse reduced the SDZ mass that was recovered in the
efﬂuent (Meff) by 5.3% and produced a corresponding increase in the
SDZ mass that was retained in the soil (Msoil). These ﬁndings were
further conﬁrmed by conducting similar SDZ co-transport experiments (Fig. 4c and d, experiments III and VII) with a higher input
concentration of SDZ (C0 ¼ 1 instead of 0.1 mg L1). Clearly, more
SDZ was retained in the upper layer in the presence of MWCNTs
when C0 ¼ 1 mg L1 (Fig. 4c and d).
The isotherm data for SDZ on MWCNT (Fig. 1c) and soil (Table 1)suggest a linear increase in the retained SDZ concentration with C0.
Similarly, co-transport experiments conducted with a higher valueof C0 for SDZ also produced greater amounts of SDZ retention(Fig. 4). However, mass balance information in Table 2 indicatesthat the value of Msoil for SDZ was greater when the experiment was
conducted with a lower (C0 ¼ 0.1 mg L1, experiment VIII) thanhigher (C0 ¼ 1 mg L1; experiment VII) input concentration of SDZ.
Since the same amount of MWCNT (1 mg L1) was injected in bothexperiments, this implies that the relative amount of SDZ retention
increased at a lower value of C0 because the kinetic rate coefﬁcientswere much higher (Table 3). All these measures demonstrate thatretained MWCNTs enhanced the sorption of SDZ in the soil column.
Unlike the CLD co-transport experiment that had both mobile
and retained MWCNTs, SDZ only interacted with retained MWCNTs.
Consequently, only parameters kaiM and kdiM needed to be optimized to the SDZ co-transport experiments (Table 3). Fitted valuesof kaiM were similar to kdiM, which demonstrated that SDZ sorption
was a rapidly reversible process. In addition, estimated values ofkaiM and kdiM were strong functions of C0 for SDZ, increasing withlower values of C0. In particular, values of kaiM and kdiM in experiment VIII (C0 ¼ 0.1 mg L1 for SDZ) were about one order ofmagnitude higher than in experiment VII (C0 ¼ 1 mg L1 for SDZ).
4. Conclusions
The hydrophobic pesticide CLD was found to have limited
mobility in the soil because of its strong partitioning to the soilorganic matter fraction. However, injection of MWCNTs facilitatedthe remobilization of a portion of the retained CLD from the soil
because MWCNTs have a large sorption capacity for CLD and agreater mobility than CLD. These results demonstrate that mobile
MWCNTs can facilitate the recovery of hydrophobic contaminants
that are retained in soils. In contrast, the readily-water soluble
antibiotic SDZ was found to be highly mobile in soil and exhibited
small amounts of kinetic sorption to both soil organic and inorganic
fractions. Immobilized MWCNTs in the soil diminished the transport of an injected pulse of SDZ because retained MWCNTs have alarge sorption afﬁnity for SDZ. Furthermore, the kinetic sorption
rate for SDZ by MWCNTs was larger at lower SDZ concentrations,and this produced greater amounts of SDZ retention. Thus, immobilized MWCNTs can reduce the contaminant mass that migrates
through the topsoil. Collectively, this research improves our understanding of how ENPs can enhance or diminish the transport ofcontaminants by considering: (i) non-simultaneous injection of
ENPs and contaminants; (ii) environmentally relevant input concentrations for ENPs and contaminants; (iii) measurement ofbreakthrough curves and retention proﬁles for contaminants and
ENPs; and (iv) numerical modeling of co-transport and interactions
between ENPs and contaminants. It is worth noting that different
environmental factors such as ionic strength, ﬂow velocity, and soil
types are known to have a large inﬂuence on MWCNT transport andretention. It is therefore expected that these same factors will havea large inﬂuence on co-transport scenarios with MWCNTs andcontaminants, and additional research is needed to address theseissues.
Acknowledgments
The ﬁrst author thanks the China Scholarship Council forﬁnancial support. The authors would like to acknowledge Stephan€ ppchen for his analysis of bromide in the liquid samples. The
Ko
technical assistance of Herbert Philipp and Claudia Walraf isgratefully acknowledged. The present work was supported by the
INRA Montpellier and the project CHLORDEXCO, funded by the
French National Research Agency (ANR-08-CES-004-01), throughthe supply of the radioactive chlordecone. Thanks are also due to
Dan Zhou for helping with the numerical investigations.
References
Bolster, C.H., Mills, A., Hornberger, G., Herman, J., 1999. Spatial distribution ofdeposited bacteria following miscible displacement experiments in intact cores.
Water Res. Res. 35 (6), 1797e1807.
Bradford, S.A., Kim, H., 2010. Implications of cation exchange on clay release andcolloid-facilitated transport in porous media. J. Environ. Qual. 39 (6), 2040.
Bradford, S.A., Kim, H.N., Haznedaroglu, B.Z., Torkzaban, S., Walker, S.L., 2009.
Coupled factors inﬂuencing concentration-dependent colloid transport andretention in saturated porous media. Environ. Sci. Technol. 43 (18), 6996e7002.
 
Bradford, S.A., Sim
unek, J., Bettahar, M., Van Genuchten, M.T., Yates, S., 2006. Signiﬁcance of straining in colloid deposition: evidence and implications. Water
Res. Res. 42 (12).
 
Bradford, S.A., Sim
unek, J., Bettahar, M., van Genuchten, M.T., Yates, S.R., 2003.
Modeling colloid attachment, straining, and exclusion in saturated porous
media. Environ. Sci. Technol. 37 (10), 2242e2250.
Bradford, S.A., Torkzaban, S., Wiegmann, A., 2011. Pore-scale simulations to determine the applied hydrodynamic torque and colloid immobilization. Vadose
Zone J. 10 (1), 252e261.
 
Bradford, S.A., Yates, S.R., Bettahar, M., Sim
unek, J., 2002. Physical factors affecting
the transport and fate of colloids in saturated porous media. Water Res. Res. 38(12), 63, 61e63e12.
Cabidoche, Y.M., Achard, R., Cattan, P., Clermont-Dauphin, C., Massat, F., Sansoulet, J.,
2009. Long-term pollution by chlordecone of tropical volcanic soils in the
French west Indies: a simple leaching model accounts for current residue.
Environ. Pollut. 157 (5), 1697e1705.
Cabidoche, Y.M., Lesueur-Jannoyer, M., 2012. Contamination of harvested organs inroot crops grown on chlordecone-polluted soils. Pedosphere 22 (4), 562e571.
Chen, C.-E., Zhang, H., Ying, G.-G., Jones, K.C., 2013. Evidence and recommendations
to support the use of a novel passive water sampler to quantify antibiotics inwastewaters. Environ. Sci. Technol. 47 (23), 13587e13593.
Fang, J., Shan, X.-q., Wen, B., Huang, R.-x., 2013. Mobility of TX100 suspended
multiwalled carbon nanotubes (MWCNTs) and the facilitated transport ofphenanthrene in real soil columns. Geoderma 207e208, 1e7.
Fernandez-Bayo, J.D., Saison, C., Geniez, C., Voltz, M., Vereecken, H., Berns, A.E.,
2013a. Sorption characteristics of chlordecone and cadusafos in tropical agricultural soils. Curr. Org. Chem. 17 (24), 2976e2984.
Fernandez-Bayo, J.D., Saison, C., Voltz, M., Disko, U., Hofmann, D., Berns, A.E., 2013b.
Chlordecone fate and mineralisation in a tropical soil (andosol) microcosm
under aerobic conditions. Sci. Total Environ. 463e464, 395e403.
Gannon, C.J., Cherukuri, P., Yakobson, B.I., Cognet, L., Kanzius, J.S., Kittrell, C.,
Weisman, R.B., Pasquali, M., Schmidt, H.K., Smalley, R.E., Curley, S.A., 2007.
Carbon nanotube-enhanced thermal destruction of cancer cells in a noninvasive
radiofrequency ﬁeld. Cancer 110 (12), 2654e2665.
 Gargiulo, G., Bradford, S., Sim
unek, J., Ustohal, P., Vereecken, H., Klumpp, E., 2007.
Bacteria transport and deposition under unsaturated conditions: the role of thematrix grain size and the bacteria surface protein. J. Contam. Hydrology 92(3e4), 255e273.
Gohardani, O., Elola, M.C., Elizetxea, C., 2014. Potential and prospective implementation of carbon nanotubes on next generation aircraft and space vehicles:
a review of current and expected applications in aerospace sciences. Prog.
Aerosp. Sci. 70, 42e68.
Hofmann, T., Von der Kammer, F., 2009. Estimating the relevance of engineeredcarbonaceous nanoparticle facilitated transport of hydrophobic organic contaminants in porous media. Environ. Pollut. 157 (4), 1117e1126.
€llrigl-Rosta, A., Vinken, R., Lenz, M., Scha€ffer, A., 2003. Sorption and dialysis exHo
periments to assess the binding of phenolic xenobiotics to dissolved organic
matter in soil. Environ. Toxicol. Chem. 22 (4), 743e752.
Iijima, S., 1991. Helical microtubules of graphitic carbon. Nature 354 (6348), 56e58.
Jaisi, D.P., Saleh, N.B., Blake, R.E., Elimelech, M., 2008. Transport of single-walledcarbon nanotubes in porous media: ﬁltration mechanisms and reversibility.
Environ. Sci. Technol. 42 (22), 8317e8323.
Ji, L., Chen, W., Zheng, S., Xu, Z., Zhu, D., 2009. Adsorption of sulfonamide antibiotics
to multiwalled carbon nanotubes. Langmuir 25 (19), 11608e11613.
Jin, Y., Yates, M.V., Thompson, S.S., Jury, W.A., 1997. Sorption of viruses during ﬂow

M. Zhang et al. / Environmental Pollution 221 (2017) 470e479
through saturated sand columns. Environ. Sci. Technol. 31 (2), 548e555.
Josko, I., Oleszczuk, P., Pranagal, J., Lehmann, J., Xing, B., Cornelissen, G., 2013. Effectof biochars, activated carbon and multiwalled carbon nanotubes on phytotoxicity of sediment contaminated by inorganic and organic pollutants. Ecol. Eng.
60, 50e59.
 Kasel, D., Bradford, S.A., Sim
unek, J., Heggen, M., Vereecken, H., Klumpp, E., 2013a.
Transport and retention of multi-walled carbon nanotubes in saturated porous
media: effects of input concentration and grain size. Water Res. 47 (2),933e944.
 
Kasel, D., Bradford, S.A., Sim
unek, J., Pütz, T., Vereecken, H., Klumpp, E., 2013b.
Limited transport of functionalized multi-walled carbon nanotubes in two
natural soils. Environ. Pollut. 180 (0), 152e158.
Kasteel, R., Mboh, C.M., Unold, M., Groeneweg, J., Vanderborght, J., Vereecken, H.,
2010. Transformation and sorption of the veterinary antibiotic sulfadiazine in
two soils: a short-term batch study. Environ. Sci. Technol. 44 (12), 4651e4657.
Kim, H., Hwang, Y.S., Sharma, V.K., 2014. Adsorption of antibiotics and iopromide
onto single-walled and multi-walled carbon nanotubes. Chem. Eng. J. 255,
23e27.
Levillain, J., Cattan, P., Colin, F., Voltz, M., Cabidoche, Y.-M., 2012. Analysis of environmental and farming factors of soil contamination by a persistent organic
pollutant, chlordecone, in a banana production area of French West Indies.
Agriculture. Ecosyst. Environ. 159 (0), 123e132.
Li, C., Ji, R., Schaffer, A., Sequaris, J.M., Amelung, W., Vereecken, H., Klumpp, E., 2012.
Sorption of a branched nonylphenol and perﬂuorooctanoic acid on Yangtze
river sediments and their model components. J. Environ. Monit. 14 (10),2653e2658.
Li, X., Zhang, P., Lin, C., Johnson, W.P., 2005. Role of hydrodynamic drag on microsphere deposition and Re-Entrainment in porous media under unfavorable
conditions. Environ. Sci. Technol. 39 (11), 4012e4020.
Liao, Q., Sun, J., Gao, L., 2008. The adsorption of resorcinol from water using multiwalled carbon nanotubes. Colloids Surfaces A Physicochem. Eng. Aspects 312(2), 160e165.
Luo, C., Wei, R., Guo, D., Zhang, S., Yan, S., 2013. Adsorption behavior of Mno2
functionalized multi-walled carbon nanotubes for the removal of cadmiumfrom aqueous solutions. Chem. Eng. J. 225, 406e415.
Mattison, N.T., O'Carroll, D.M., Kerry Rowe, R., Petersen, E.J., 2011. Impact of porous
media grain size on the transport of multi-walled carbon nanotubes. Environ.
Sci. Technol. 45 (22), 9765e9775.
McArdell, C.S., Molnar, E., Suter, M.J.-F., Giger, W., 2003. Occurrence and fate ofmacrolide antibiotics in wastewater treatment plants and in the glatt valley
watershed, Switzerland. Environ. Sci. Technol. 37 (24), 5479e5486.
Test No. 106 OECD, 2000. Adsorption e Desorption Using a Batch Equilibrium
Method.
Pailler, J.-Y., Krein, A., Pﬁster, L., Hoffmann, L., Guignard, C., 2009. Solid phaseextraction coupled to liquid chromatography-tandem mass spectrometry
analysis of sulfonamides, tetracyclines, analgesics and hormones in surface
water and wastewater in Luxembourg. Sci. Total Environ. 407 (16), 4736e4743.
 Pang, L., Sim
unek, J., 2006. Evaluation of bacteria-facilitated cadmium transport ingravel columns using the Hydrus colloid-facilitated solute transport model.
Water Res. Res. 42 (12).
Pritchard, P.H., Monti, C.A., O'Neill, E.J., Ahearn, D.G., Connolly, J.P., 1986. Movementof Kepone®(chlordecone) across an undisturbed sediment-water interface inlaboratory systems. Environ. Toxicol. Chem. 5 (7), 647e657.
Ptak, T., Piepenbrink, M., Martac, E., 2004. Tracer tests for the investigation ofheterogeneous porous media and stochastic modelling of ﬂow and transportda
review of some recent developments. J. Hydrology 294 (1), 122e163.
Ren, X., Chen, C., Nagatsu, M., Wang, X., 2011. Carbon nanotubes as adsorbents inenvironmental pollution management: a review. Chem. Eng. J. 170 (2e3),
395e410.

479

Schijven, J.F., Hassanizadeh, S.M., 2000. Removal of viruses by soil passage: overview of modeling, processes, and parameters. Crit. Rev. Environ. Sci. Technol. 30(1), 49e127.
 

Sim
unek, J., Genuchten, M.T.v., Sejna,M., 2016. Recent developments and applications of the Hydrus computer software packages. Vadose Zone J. 6 (15).
 Simunek, J., He, C., Pang, L., Bradford, S.A., 2006. Colloid-facilitated solute transportin variably saturated porous media. Vadose Zone J. 5 (3), 1035.
 

Sim
unek, J., Sejna,M., Genuchten, M.T.v., 2012. The C-Ride module for Hydrus (2D/
3D): simulating two-dimensional colloid-facilitated solute transport. In: Variably-saturated Porous Media.
Su, C., Zeng, G.-M., Gong, J.-L., Yang, C.-P., Wan, J., Hu, L., Hua, S.-S., Guo, Y.-Y., 2016.
Impact of carbon nanotubes on the mobility of sulfonamide antibiotics insediments in the Xiangjiang river. RSC Adv. 6 (21), 16941e16951.
ment, M., Chevreuil, M.,
Tamtam, F., Mercier, F., Le Bot, B., Eurin, J., Dinh, Q.T., Cle2008. Occurrence and fate of antibiotics in the seine river in various hydrological conditions. Sci. Total Environ. 393 (1), 84e95.
Tian, Y., Gao, B., Wang, Y., Morales, V.L., Carpena, R.M., Huang, Q., Yang, L., 2012.
Deposition and transport of functionalized carbon nanotubes in watersaturated sand columns. J. Hazard. Mater. 213e214, 265e272.
Tong, M., Johnson, W.P., 2007. Colloid population heterogeneity drives hyperexponential deviation from classic ﬁltration theory. Environ. Sci. Technol. 41 (2),493e499.
Treumann, S., Torkzaban, S., Bradford, S.A., Visalakshan, R.M., Page, D., 2014. An
explanation for differences in the process of colloid adsorption in batch andcolumn studies. J. Contam. Hydrology 164, 219e229.
Tufenkji, N., Elimelech, M., 2005. Spatial distributions of cryptosporidium oocysts inporous media: evidence for dual mode deposition. Environ. Sci. Technol. 39(10), 3620e3629.
Unold, M., Kasteel, R., Groeneweg, J., Vereecken, H., 2009. Transport and transformation of sulfadiazine in soil columns packed with a silty loam and a loamy
sand. J. Contam. Hydrology 103 (1e2), 38e47.
Upadhyayula, V.K., Deng, S., Mitchell, M.C., Smith, G.B., 2009. Application of carbon
nanotube Technology for removal of contaminants in drinking water: a review.
Sci. Total Environ. 408 (1), 1e13.
van de Weerd, H., Leijnse, A., 1997. Assessment of the effect of kinetics on colloid
facilitated radionuclide transport in porous media. J. Contam. Hydrology 26 (1),245e256.
Van Genuchten, M.T., Wagenet, R., 1989. Two-site/two-region models for pesticidetransport and degradation: theoretical development and analytical solutions.
Soil Sci. Soc. Am. J. 53 (5), 1303e1310.
Wang, D., Paradelo, M., Bradford, S.A., Peijnenburg, W.J., Chu, L., Zhou, D., 2011.
Facilitated transport of Cu with hydroxyapatite nanoparticles in saturated sand:effects of solution ionic strength and composition. Water Res. 45 (18),5905e5915.
Wang, Y., Kim, J.H., Baek, J.B., Miller, G.W., Pennell, K.D., 2012. Transport behavior offunctionalized multi-wall carbon nanotubes in water-saturated quartz sand as afunction of tube length. Water Res. 46 (14), 4521e4531.
 
Wehrhan, A., Kasteel, R., Sim
unek, J., Groeneweg, J., Vereecken, H., 2007. Transportof sulfadiazine in soil columns: experiments and modelling approaches.
J. Contam. Hydrology 89 (1e2), 107e135.
Woignier, T., Fernandes, P., Jannoyer-Lesueur, M., Soler, A., 2012. Sequestration ofchlordecone in the porous structure of an andosol and effects of added organic
matter: an alternative to decontamination. Eur. J. Soil Sci. 63 (5), 717e723.
Zarﬂ, C., 2008. Chemical Fate of Sulfadiazine in Soil: Mechanisms and Modelling
Approaches (Shaker).
Zhang, L., Wang, L., Zhang, P., Kan, A.T., Chen, W., Tomson, M.B., 2011. Facilitatedtransport of 2,20 ,5,50 -polychlorinated biphenyl and phenanthrene by fullerene
nanoparticles through sandy soil columns. Environ. Sci. Technol. 45 (4),1341e1348.