Nanoscale Metallic Iron for Environmental Remediation:
Prospects and Limitations
Chicgoua Noubactep & Sabine Caré &
Richard Crane
Received: 23 June 2011 /Accepted: 6 September 2011 /Published online: 22 September 2011
Abstract The amendment of the subsurface with
nanoscale metallic iron particles (nano-Fe0) has been
discussed in the literature as an efficient in situ
technology for groundwater remediation. However,
the introduction of this technology was controversial
and its efficiency has never been univocally estab-
lished. This unsatisfying situation has motivated this
communication whose objective was a comprehensive
discussion of the intrinsic reactivity of nano-Fe0 based
on the contemporary knowledge on the mechanism of
contaminant removal by Fe0 and a mathematical model.
It is showed that due to limitations of the mass transfer
of nano-Fe0 to contaminants, available concepts cannot
explain the success of nano-Fe0 injection for in situ
groundwater remediation. It is recommended to test the
possibility of introducing nano-Fe0 to initiate the
formation of roll-fronts which propagation would
induce the reductive transformation of both dissolved
and adsorbed contaminants. Within a roll-front, FeII
from nano-Fe0 is the reducing agent for contaminants.
FeII is recycled by biotic or abiotic FeIII reduction.
While the roll-front concept could explain the success
of already implemented reaction zones, more research
is needed for a science-based recommendation of nano-
Fe0 for subsurface treatment by roll-fronts.
Keywords Environmental remediation .Material
reactivity . Nanoscale iron . Roll-front . Zerovalent iron
1 Introduction
The development of new methods and materials for
environmental remediation is a real challenge for the
scientific community. Such technologies will only be
adopted by industry if they can exhibit marked
improvements in efficiency, affordability or eco-
compatibility compared with conventional techniques.
The use of metallic iron (Fe0) in subsurface reactive
permeable barriers has been identified as such a
technology (Gillham and O’Hannesin 1991, 1994;
Matheson and Tratnyek 1994; Blowes et al. 1997;
Water Air Soil Pollut (2012) 223:1363–1382
DOI 10.1007/s11270-011-0951-1
C. Noubactep (*)
Angewandte Geologie, Universität Göttingen,
Goldschmidtstraße 3,
37077 Göttingen, Germany
e-mail: cnoubac@gwdg.de
S. Caré
Université Paris-Est, Laboratoire Navier,
(ENPC/IFSTTAR/CNRS),
2 allée Kepler,
77420 Champs sur Marne, France
e-mail: sabine.care@ifsttar.fr
R. Crane
Interface Analysis Centre, University of Bristol,
121 St. Michael’s Hill,
Bristol BS2 8BS UK
e-mail: Richard.Crane@bristol.ac.uk
C. Noubactep
Kultur und Nachhaltige Entwicklung CDD e.V.,
Postfach 1502,
37005 Göttingen, Germany
# The Author(s) 2011. This article is published with open access at Springerlink.com
O’Hannesin and Gillham 1998). Since this discovery
almost 20 years ago, extensive research of the Fe0/
H2O system has been performed in an attempt to
understand the controlling mechanisms behind the
remediation of redox-amenable contaminant species
using Fe0-based materials (Scherer et al. 2000;
Henderson and Demond 2007; Bartzas and Komnitsas
2010; Gillham 2010; Li and Benson 2010; Noubactep
2010a, b; Noubactep and Caré 2010a; Kim et al. 2010;
Comba et al. 2011; Courcelles et al. 2011; Gheju
2011; Jeen et al. 2011; Noubactep 2011a). Two
different tools are commonly used to optimise the
efficiency of Fe0 for aqueous contaminant removal:
(1) reducing the particle size of Fe0 down to the
nanoscale (nano-Fe0) (Wang and Zhang 1997;
Ponder et al. 2000) and (2) using bimetallic systems
(Muftikian et al. 1995; Korte et al. 2000). In recent
years, there has been considerable interest into
combining the two methods (Schrick et al. 2002;
Karn et al. 2009; Nagpal et al. 2010; Chen et al.
2011; Mossa et al. 2011).
Since the original proof of concept study into the
application of nano-Fe0 for water treatment at Lehigh
University, USA (Wang and Zhang 1997), research
within this field has boomed. On August 20th 2011, a
search at “ACS publications”, “Science Direct”
(Elsevier journals), and “Springer journals” using
key words “nanoscale” and “zerovalent iron” yielded
740 peer-reviewed articles (Table 1). As shown in
Table 1, 234 articles have already been published in
the first 7 months of 2011 by the three selected
publishers. This clearly demonstrates the interest
within academia for this technology.
In recent years, several review articles and critical
views on nano-Fe0 for environmental remediation
have been published (Zhang et al. 1998; Gillham
2003; Zhang 2003; Li et al. 2006; Macé 2006; Macé
et al. 2006; Tratnyek and Johnson 2006; Karn et al.
2009; Pradeep and Anshup 2009; Agarwal and Joshi
2010; Gillham 2010; Müller and Nowack 2010;
Noubactep and Caré 2010b; Cullen et al. 2011;
Noubactep 2011b; Peralta-Videa et al. 2011; Shi et
al. 2011; Truex et al. 2011a, b). However, the original
discussion on the suitability of nano-Fe0 for in situ
field applications (Gillham 2003, 2010) has not been
satisfactorily addressed (Macé 2006; Noubactep
2011b). Moreover, a recent comparison between field
applications of Fe0 of different particle sizes (nano-
metres, micrometres, and millimetres) for field applica-
tions has clearly demonstrated the superiority of mm-
Fe0 (average efficacy 97%) (Comba et al. 2011). The
decreasing order of reactivity was mm-Fe0 (97%)>μm-
Fe0 (91%)>nano-Fe0 (65%). Expectably, the lower
efficiency of nano-Fe0 is due to their high reactivity
(Gillham 2003, 2010; Noubactep and Caré 2010b).
Therefore, the question arises on the fundamental
necessity to further increase the reactivity of nano-Fe0
by using a noble metal combination.
1.1 The Problem
Nano-Fe0 technology for environmental remediation
was introduced as an alternative to the conven-
tional Fe0 walls mostly for inaccessible aquifers
(Masciangioli and Zhang 2003; Li et al. 2006). The
very small particle size of nano-Fe0 (1–100 nm)
would allow the material to penetrate deep into soil
networks (Ghauch et al. 2009; Karn et al. 2009; Comba
et al. 2011; Gheju 2011; Truex et al. 2011a, b).
Due to the exponential relationship between
specific surface area (SSA) and radius (R=d/2) of
a perfectly spherical object (SSA=4πR2), as a rule, a
decrease in Fe0 particle size increases the surface
area per gram by up to three orders of magnitude
(Macé et al. 2006; Chen et al. 2011). In other words,
the inverse relationship between Fe0 particle size
and reactivity is due to a greater density of reactive
sites on the particle surface at smaller scale. The
following three claims have been made with regard
to the use of nano-Fe0 for aqueous contaminant
removal (Comba et al. 2011): (1) some aqueous
contaminant species that have been proven as
unsuccessful for remediation using μm-Fe0 and
mm-Fe0 can be effectively removed using nano-
Fe0, (2) nano-Fe0 can be used for more rapid
Table 1 Results of a web-search for three relevant publishers
demonstrating the current interest within academia for the
nano-Fe0 technology (search: 20 August 2011)
Publisher Search’s results
Period Total 2011
ACS publications 2004 to 2011 263 96
Elsevier journals 1993 to 2011 392 112
Springer journals 2003 to 2011 85 26
Total 740 234
1364 Water Air Soil Pollut (2012) 223:1363–1382
degradation of contaminants, and (3) the formation
of some undesirable by-products during remediation
using μm- and mm-Fe0 can be avoided by using
nano-Fe0. Such processes whilst correct are all
linked to the greater reactivity that nano-Fe0
possesses due to its size (reactive surface area).
When performed in conditions without a large nano-
Fe0 stoichiometric excess, e.g. a system analogous
to the environment, it may prove that such claims
will be unfounded (Nagpal et al. 2010; Noubactep
and Caré 2010b; Noubactep 2011b; Sakulchaicharoen
et al. 2010; Nagpal et al. 2011).
An undisputed drawback with regards to the
use of nano-Fe0 for environmental applications is
their strong tendency to aggregate and adhere to
solid surfaces (Li et al. 2006; Tratnyek and Johnson
2006; Karn et al. 2009; Sakulchaicharoen et al.
2010; Comba et al. 2011; Gheju 2011; Truex et al.
2011a, b). Karn et al. (2009) listed some parameters
that influence nano-Fe0 adsorption onto soil and
aquifer materials: (1) the surface chemistry of soil
and Fe0 particles, (2) the groundwater chemistry
(e.g. ionic strength, pH and presence of natural
organic matter), and (3) the hydrodynamic condi-
tions (pore size, porosity, flow velocity and degree
of mixing or turbulence). Several methods have
been developed for the stabilization of nano-Fe0
particles over the past decade and proven efficient to
sustain the reactivity of nano-Fe0 (Karn et al. 2009;
Sakulchaicharoen et al. 2010; Comba et al. 2011;
Gheju 2011). One factor that has been overlooked,
however, is the impact volumetric expansion has on
the mobility of (1) residual Fe0, (2) primary
corrosion products (FeII and H2) and contaminants.
The volume of any corrosion product (Fe hydroxide
or oxide) is higher than that of the original metal
(Fe0). The ratio between the volume of expansive
corrosion product and the volume of iron consumed
in the corrosion process is called “rust expansion
coefficient” (η) (Anstice et al. 1993; Caré et al.
2008; Zhao et al. 2011). Volumetric corrosion
products are likely to: (1) contribute to porosity
loss, (2) impact the retention of contaminants and
transformation products, and (3) increase the parti-
cle agglomeration.
Another area of heightened research is with regard
to the determining the toxicity of nano-Fe0, with
mixed results reported (Grieger et al. 2010; Gheju
2011). For example, Barnes et al. (2010) reported
minimal change to the structure of a river water
community due to the addition of nano-Fe0 while
Diao and Yao (2009) reported nano-Fe0 particles as
highly cytotoxic towards both gram-positive and
gram-negative bacteria species.
While taking into account all known influencing
parameters, the following seven features have to be
systematically studied in order to optimise the general
applicability of this technique (Karn et al. 2009;
Tervonen et al. 2009; Gheju 2011): (1) mobility
changes due to nano-Fe0 volumetric expansion during
corrosion, (2) the bioavailability of Fe0 and corrosion
products (FeII/FeIII species, H/H2), (3) the ecotoxicity
of Fe0 and its corrosion products, (4) the bioaccumu-
lation of Fe0 and its corrosion products, (5) the
translocation potential of nano-Fe0, (6) the long-term
reactivity of nano-Fe0 particles and (7) the speciation,
persistence and fate of contaminants and their
transformation products. A major contributing factor
to the latter point is that little is known (compared to
permeable reactive barrier technologies) about the
extent contaminants are removed via size exclusion
using nano-Fe0.
Only when all seven “operational drivers” have
been determined can the global community have full
faith in the technology.
1.2 Objectives of the Study
The present communication is focused on the “field
persistence” or reactive “life span” of nano-Fe0
particles. For in situ applications, a keen understand-
ing of nano-Fe0 reactive fate is essential for effective
and prudent site clean-up. The knowledge of which is
likely to largely underpin decisions as to (1) the
choice of material selected, (2) the mechanism of
application and (3) the strategy (if any) for repeated
treatments.
In the current work, a multidisciplinary approach is
used to analyse the relationship between nano-Fe0
reactivity and its performance for in situ field
applications. The discussion is based on the contem-
porary knowledge of the mechanism of aqueous
contaminant removal by Fe0 (Noubactep 2007,
2008, 2010c, 2011d, e). Much of the impetus for this
work has come from the work of Noubactep and Caré
(2010b) who have challenged the concept that nano-
Fe0 is a strong reducing agent for contaminant
reductive transformation.
Water Air Soil Pollut (2012) 223:1363–1382 1365
2 Nanoscale Metallic Iron for Environmental
Remediation
To date, nano-Fe0 particles have been reported as largely
successful for water and soil treatment (Pradeep and
Anshup 2009; Agarwal and Joshi 2010; Litter et al. 2010;
Comba et al. 2011; Crane et al. 2011). A wide variety of
redox-amenable organic and inorganic species and non-
reducible species (e.g. Cd and Zn) have been efficiently
treated. Similar to μm- and mm-Fe0, adsorption is
considered important only for non-reducible species
(Celebi et al. 2007; Li and Zhang 2007; Noubactep
2007, 2008, 2010c; Boparai et al. 2011; Noubactep
2011c, d; Xiao et al. 2011). For example, Boparai et al.
(2011) reported that heavy metals are either reduced (e.g.
Cu2+ and Ag2+) at, or directly adsorbed (e.g. Zn2+ and
Cd2+) onto the Fe0 surface. They further argued that “the
controlling mechanism is a function of the standard redox
potential of the contaminant”. Recent work has however
challenged this concept (Noubactep 2010c, 2011b, d, e),
which is explained below.
2.1 Contaminant Reduction by Nano-Fe0
The chemical reaction between Fe0 and redox-
amenable aqueous species is considered to involve
three steps: (1) direct electron transfer from Fe0 at the
metal surface or through a conductive oxide film on
Fe0 (direct reduction), (2) catalyzed hydrogenolysis
by the H/H2 (indirect reduction mechanism 1) and (3)
reduction by FeII species resulting from Fe0 corrosion
(indirect reduction mechanism 2). In this constella-
tion, H2 is supposed to result from H2O reduction
during anoxic iron corrosion (Reardon et al. 2008;
Chen et al. 2011). However, evidence exists in the
literature, e.g. Stratmann and Müller (1994), that even
under external oxic conditions, Fe0 is oxidized by
H2O (or more precisely by H
+) and O2 by Fe
II
(Table 2). Despite the significant reaction rate
exhibited by nano-Fe0 due to its high surface area,
such processes are considered to occur (discounting
any quantum size effects) independent of particle size.
Table 2 summarizes the half reactions for the
aqueous oxidation of Fe0 under both anoxic and oxic
conditions. Thermodynamically, the major cathodic
reaction depends on the availability of molecular O2
(E0=0.81 V). In the absence of O2, Fe
0 is oxidized by
H+ (E0=0.00 V). It can therefore be stated that the
rate of Fe0 oxidation is dictated by the concentration
of dissolved O2, H
+ and H2O in proximity to Fe
0
surfaces. Le Chatelier’s principle also states that the
consumption of FeII (via oxidation to FeIII) will also
result in an increase in Fe0 oxidation.
The electrode potential of the redox couple FeII/Fe0
is −0.44 V, a value which is independent of the
particle size (nanometre, micrometre or millimetre).
The value −0.44 V is considered largely unchanged
due to the presence of alloying materials (e.g. low
alloy steel and bimetallic systems).
As a consequence, statements including “nano-Fe0
are more reactive than μm-Fe0 and mm-Fe0” are
misleading; as the reactivity of Fe0 (discounting
quantum size effects), is independent of the particle
size. Any enhanced reactivity reported is likely to be
due to the significantly high surface area of nano-Fe0
compared with other forms. A second statement
“bimetallic nano-Fe0 is more reactive that monome-
tallic nano-Fe0” is also a qualitative statement, as the
reactivity of the materials depends on numerous
factors associated with the materials synthesis route
and varies depending on the chemistry of the chosen
alloying metal. Ideally, comparisons should be made
versus standard reference materials using established
standard experimental protocols (Noubactep et al.
2009), which once established, will significantly
improve the design of future field applications.
2.2 Limitations of the Nano-Fe0 Technology
The efficiency of nano-Fe0 for aqueous contaminant
reduction faces some key issues for in situ applica-
Table 2 Relevant redox couples for the process of aqueous Fe0
dissolution and oxide scale formation in a passive remediation
Fe0/H2O system. These processes are thermodynamically the
same for all Fe0 particle sizes. Observed differences are due to
kinetics aspects
Electrode reactions Eq.
Fe0 ⟺ Fe2++2e− (I)
Oxic conditions
O2+2H2O+4e
− ⟺ 4OH− (IIa)
2H2O+2e
− ⟺ H2+2OH
− (IIb)
Anoxic conditions
O2+4H
++4e− ⟺ 2H2O (IIIa)
2H++2e− ⟺ H2 (IIIb)
These processes are thermodynamically the same for all Fe0
particle sizes. Observed differences are due to kinetics aspects
1366 Water Air Soil Pollut (2012) 223:1363–1382
tions in porous media. These challenges include: (1)
the strong tendency of aggregation/agglomeration, (2)
the rapid settlement on subsurface solid phases and
(3) the porosity and permeability loss of porous media
(Behrens et al. 2000; Dickinson and Scott 2010;
Cullen et al. 2011; Mossa et al. 2011). Aggregation
and settlement limit nano-Fe0 transport through porous
media. Porosity and permeability loss limit nano-Fe0
transport to target contaminants. It was demonstrated
that nano-Fe0 may travel only a few centimetres in
porous media from the injection position under typical
groundwater conditions (Tratnyek and Johnson 2006;
Johnson et al. 2008; Comba et al. 2011; Gheju 2011).
Accordingly, recent efforts have been made to (1)
increase the porosity of porous media, (2) mechanically
increase the distribution of nano-Fe0 and/or (3)
chemically modify nano-Fe0 for improved aqueous
mobility in porous networks.
2.3 Improving the Efficiency of the Nano-Fe0
Systems
2.3.1 Dispersion Agents
Methods to improve the aqueous mobility of nano-Fe0
have received the greatest research interest. It has
been determined that the key to improving particle
mobility is found in modifying their surface properties
such that the nano-Fe0 have significantly improved
colloidal stability and a commensurate reduction in
the likelihood of adherence to mineral surfaces.
Several synthetic methods are now available to
produce more mobile nano-Fe0. Efficiently tested
dispersants include anionic surface chargers (e.g.
polyacrylic acid), non-ionic surfactants, starch and
oil (Wu et al. 2009; Fang et al. 2011; Jiang et al.
2011; Mossa et al. 2011; Tong et al. 2011).
2.3.2 Bimetallic Combinations
In recent years, noble metals have been used to
increase the reactivity of monometallic nano-Fe0
(Nagpal et al. 2010; Yuan et al. 2010; Zhu et al.
2010; Kadu et al. 2011; Mossa et al. 2011). As
mentioned above, this appears counterintuitive as
nano-Fe0 is already highly reactive due to its size
(Wang and Zhang 1997; Zhang 2003; Macé et al.
2006; Ghauch et al. 2009) and is unstable during
synthesis, storage and application (Jiang et al. 2011).
This chemical instability has been documented as a
key reason for the observed lower efficiency exhibited
by nano-Fe0 systems compared with μm- and mm-Fe0
(Comba et al. 2011). Accordingly, it is questionable
whether further enhancing the reactivity of nano-Fe0,
e.g. by plating with more noble elements, may be of
any benefit. The reactivity of nano-Fe0 will be
discussed in the next section on the basis of
mathematical modelling.
3 Significance of Increased Reactivity
3.1 The Problem
The increased Fe0 reactivity from mm to nm size
should be better characterized. The relative reactivity
of four different materials is discussed on the basis
of 1 kg Fe0: 1 nm-Fe0 (d0=25 nm), 1 μm-Fe
0 (d0=
25 μm) and 2 mm-Fe0 (d0=250 and 1,000 μm).
Calculations for the number particles (N) in 1 kg of
each material and the number of layers (N′) in each
particle are made after the Eqs. 1 and 2 presented in
details elsewhere (Noubactep and Caré 2010a;
Noubactep et al. 2010).
N ¼ M
rFe  4=3 p  R30

   ð1Þ
N 0 ¼ 2  4=3 pR30

  
=a3 ð2Þ
where M is the mass of Fe0 (here 1 kg), ρFe is the
specific weight of Fe (7,800 kg/m3), R0 is the initial
radius of the Fe particle (d=2*R0) and a the lattice
parameter (a=2.866Å).
The results are summarized in Table 3. It can be
seen that the number of layers of Fe0 in individual
particles varies from 87.2 for nano-Fe0 to more than
3*106 for mm-Fe0 (d=1 mm). In the meantime, the
number of particles in 1 kg decreased from 1.96*1018
for nano-Fe0 to only 3.1*104 for mm-Fe0. The ratio of
the number of Fe0 layers in each particle to the
number of Fe0 layers in nano-Fe0 varies from 1 to
4*104. This ratio corresponds to the relative time (C )
as defined later (section 3.2). On the other hand, the
ratio of the number of particles in 1 kg of nano-Fe0 to
the number of particles in the same mass of each other
materials varies from 1 to 6.4*1013. These results are
Water Air Soil Pollut (2012) 223:1363–1382 1367
summarized in Fig. 1. Instead of the mass of Fe0, the
number of electrons released by the conversion of Fe0
to FeII is used to assess the kinetics of Fe0
consumption. This is discussed in the next section.
3.2 Relative Corrosion Kinetics of Fe0 Materials
For the discussion in this section, uniform corrosion
for spherical particles is assumed. Individual particles
corrode independently until material depletion. It is
further assumed for simplicity that individual layers
corrode with the same kinetics independent of particle
size. The latter assumption is conservative as larger
particles react slower than smaller (Macé et al. 2006;
Reardon et al. 2008; Nagpal et al. 2010). With these
assumptions, a relative time (C ) can be defined while
taking the time for the corrosion of the smallest
particle (here 87.2 layers of nano-Fe0) or t∞, nano as
unit.
t ¼ t=t1; nano ð3Þ
Accordingly, one unit of time corresponds to the
time to nano-Fe0 depletion. Remember that all
1.96*1018 particles in the 1 kg of nano-Fe0 simulta-
neously corrode with the same kinetics. The results of
the calculations are presented in Fig. 2. From Fig. 2a
and Table 3, it can be seen that after nano-Fe0
depletion, the material with 1,000-μm (or 1 mm)
diameter will still react for more than 3*104 times
longer than the time necessary for nano-Fe0 depletion
(C =4*104, see Table 3). Figure 2b shows that the
mm-Fe0 with 250-μm diameter is depleted after about
104*t∞, nano.
Based on the assumptions above, the service life of
a nano-Fe0 particle can be estimated. Table 4 summa-
rizes the results of such estimations while varying the
service life of a 1 mm Fe0 particle from 5 to 40 years.
This assumption is based on the fact that conventional
Fe0 walls are supposed to function for several decades
(here up to 4 decades). Results show (Table 4) that the
maximum life span of a nano-Fe0 is about 8.8 h (less
than 1 day). In other words, following approximately
9 h from subsurface deployment it is suggested that
all nano-Fe0 would be reactively exhausted. The
success of this is dependent on three key factors: (1)
the hydrodynamic conditions: pore size, porosity,
flow velocity and degree of mixing or turbulence,
(2) the water chemistry and the affinity of nano-Fe0
and its transformation products to the soil materials
and (3) the reactivity of Fe0.
It is certain that the dynamic process of transforma-
tion of concentric layers of Fe0 atoms to concentric
layers of iron (hydr)oxides cannot be linear (Noubactep
Table 3 Summary of the values of the number of particles contained in 1 kg of each material, the number of layer making up each
particle and estimation of the relative time (C )
Size d (μm) nparticles (−) nlayers (−) nlayers/nnano (−) nnano/nparticles (−) C (−)
Nano-Fe0 25*10−3 1.96*1018 87.2 1.0 1.0 1.0
μm-Fe0 25 1.96*109 87.2*103 103 109 103
mm-Fe0 250 1.96*106 87.2*104 104 1012 104
mm-Fe0 1,000 3.06*104 3.49*105 4*104 6.4*1013 4*104
The life span of nano-Fe0 is operationally considered as the unit of time while assuming uniform corrosion. C coincides with the ratio
of the number of layers of Fe in each particle to that of nano-Fe0 . The ratio of the number of particles in individual systems is also
given
0 200 400 600 800 1000
-6
-4
-2
0
2
4
 number of layers
 kinetics  e- release
lo
gX
particle size / [μm]
Fig. 1 Comparison of the evolution of the kinetics of electron
release and the number of layers in each Fe0 particle as a
function of the particle size. It is shown that smaller particles
release huge amounts of electrons within a very short time.
Calculations are made for 1 kg of Fe0 material
1368 Water Air Soil Pollut (2012) 223:1363–1382
and Caré 2010b). In fact, effects similar to “case
hardening” for food and wood drying will lead to
“surface hardened layers” (Tarvainen et al. 2006;
Fernando et al. 2008) leading to differential kinetics/
extents of Fe0 passivation for different particle size
ranges. In other words, the extent of restricted
corrosion rates through resulting surface hardened
layers will be different for nm-, μm- and mm-Fe0.
Bearing this in mind, the very short relative life span of
a nano-Fe0 estimated above will be used for the
discussion in this work. It is certain that “case-
hardening”-like effects will prolong this hypothetical
life span to some days or weeks.
3.3 Extent of Iron Corrosion from Fe0 Materials
A discussion as to the extent of Fe0 consumption is
limited in the present section to C =1 or t∞, nano. It is
considered for simplification that the sole iron
corrosion product is Fe3O4. The corresponding coef-
ficient of volumetric expansion is ηFe3O4=2.08
(Eq. 4) (Caré et al. 2008). Using ρ=M/V, the volume
of Fe corresponding to 1 kg Fe0 is calculated as
127.0 mL (V0). This is the initial volume of Fe
0 (V0).
Following corrosion, this volume is partly or totally
consumed. The volume (ΔV) corresponding to the
volume of pores occupied by the volumetric expan-
sion of corrosion products can be estimated.
Assuming that the coefficient of volumetric expan-
sion (η) (“rust expansion coefficient” or “specific
volume”) (Anstice et al. 1993; Caré et al. 2008; Zhao
et al. 2011) of the reaction products is:
h ¼ Voxide=VFe ð4Þ
where Voxide is the volume of the reaction product and
VFe the volume of the parent Fe
0.
The volumeΔV characterizing the extent of porosity
loss due to volumetric expansion is given by Eq. 5:
$V ¼ h 1ð Þ»Vconsumed Fe ð5Þ
$V ¼ V1  V0 ¼ n» hFe3O4  1ð Þ»V0 ð5aÞ
Where Vconsumed Fe is the volume of consumed Fe
0
at time t∞, V0 is the volume occupied by the initial Fe
0
particles and ν (ν≤1) is the fraction of the initial
0 10000 20000 30000
0
20
40
60
80
100
(a)
Fe0 particle size:
 25 nm
 25 μm 
 250 μm 
 1000 μm 
Fe
0  
co
n
su
m
pt
io
n 
/ [%
]
Fe
0  
co
n
su
m
pt
io
n 
/ [%
]
relative time / [-]
0 2000 4000 6000 8000 10000
0
20
40
60
80
100
(b)
Fe0 particle size:
 25 nm
 25 μm 
 250 μm 
relative time / [-]
Fig. 2 Kinetics of the process of Fe0 exhaustion at nano-,
micro- and millimetre scale as for: a d≤1,000 μm and b d≤
200 μm
Table 4 Estimation of the value of the life span (t∞) of a nano-Fe
0 particle with 25 nm diameter for barrier life spans (t) from 10 to
40 years
t (years) 5 10 15 20 25 30 35 40
tμm (days) 45.7 91.3 137.0 182.6 228.3 273.9 319.6 365.3
tμm (years) 0.2 0.3 0.5 0.7 0.9 1.0 1.2 1.4
t∞ (hours) 1.1 2.2 3.3 4.4 5.5 6.6 7.7 8.8
The considered conventional reactive wall contains granular Fe0 with a diameter of 1 mm. For comparison, the relative life span (in
days and years) of the micrometric particles is given
Water Air Soil Pollut (2012) 223:1363–1382 1369
amount of Fe0 (1 kg) which has reacted at t∞, nano. V∞
is the total volume occupied by residual Fe0 and in
situ formed corrosion products. t∞ corresponds to
nano-Fe0 depletion (25 nm in this section). In the
discussion on the reactivity at nanoscale, t∞ corre-
sponds to the depletion of the material with 10-nm
diameter (Section 4).
V1 ¼ h»Vconsumed Fe þ V0  Vconsumed Feð Þ ð6Þ
V1 ¼ V0» 1þ n» hFe3O4  1ð Þ½  ð6aÞ
The volumetric expansion (ΔV, Eq. 5) can be
characterized as percent of the initial volume (V0)
using Eq. 7:
$V %ð Þ ¼ 100»n» hFe3O4  1ð Þ ð7Þ
Table 5 summarizes the results. It is shown that at
C =1 (nano-Fe0 depletion), a volume augmentation of
108% has occurred in the nano-Fe0 system, with
volume augmentations in all other systems lower than
0.5%. This clearly shows that the porosity of the
subsurface will be significantly influenced by nano-
Fe0 at t∞. Remember that 100% reactive exhaustion of
nano-Fe0 is predicted to occur by approximately 9 h
time. During this same period the porosity loss due to
expansive iron corrosion is likely to be negligible for
all other Fe0 particle size fractions. Calculations for
Akageneite β-FeOOH (ηFeOOH=3.48) as sole corro-
sion products shows that V∞,nano=448.7 mL, ΔV=
320.5 mL or 250.4%. The examples of Fe3O4
(anoxic) and FeOOH (oxic) demonstrate the crucial
importance of the nature of formed corrosion products
for the discussion of the extent of porosity loss.
Another important aspect of Fe0 consumption is
given by the number of moles of Fe0 that have been
oxidized (Table 5). Assuming contaminant reduc-
tion, Table 5 shows that after C =1, 35.71 moles of
electrons have been released in the nano-Fe0 system
but less than 0.11 moles in all other systems. In
other words, up to 35.71 moles of electrons are
available for contaminant reduction per kg nano-Fe0
within a few hours of reaction (<9 h). But what
proportion of the electrons produced would reach
the contaminant within this period? That is the
major question to be answered for the further
development of the nano-Fe0 technology for in situ
applications.
4 Reactivity of Nano-Fe0 Materials
The presentation until now has discussed the reactiv-
ity of nano-Fe0 in comparison to larger scale Fe0.
Section 4 will focus only on the nanoscale size
fraction (d≤100 nm). Equations 1–7 will be used and
the particle size will vary from 10 to 100 nm. As
stated above t∞ is for a nano-Fe
0 of 10 nm in diameter
and the reaction proceeding until 100% reactive
exhaustion has been achieved.
4.1 Fe0 Reactivity at Nanoscale
Table 6 summarizes the results of calculations for the
number of Fe0 particles and number of layers of Fe0
in each nano-Fe0. It is shown that 1 kg of the material
with d=10 nm contains 1,000 times more particles
than a material of d=100 nm.
Table 5 Estimation of the extent of porosity loss (ΔV) due to the volumetric expansion of iron corrosion for Fe0 particles of different
sizes
Size mconsumed (kg) ν (%) V∞ (mL) ΔV (mL) ΔV (%) nFe(II) (moles) nFe3O4 (moles) nelectrons (moles) nnano/nelectrons (−)
nm-Fe0 100 100.00 264.16 137.16 108.00 17.857 5.9524 35.714 1
μm-Fe0 3*10−3 0.30 127.41 0.41 0.32 0.053 0.0178 0.107 335
mm-Fe0 3*10−4 0.03 127.04 0.04 0.03 0.005 0.0018 0.011 3,342
mm-Fe0 7.5*10−5 0.01 127.01 0.01 0.01 0.001 0.0004 0.003 13,369
The operational unit of time is arbitrarily the time to nano-Fe0 depletion (t∞)
V0 volume occupied by the initial Fe
0 particles, V∞ volume occupied by residual Fe
0 and in situ formed corrosion products. ΔV
volume of pore occupied by the volumetric expansion of corrosion products, mconsumed mass of Fe
0 consumed, ν percent of Fe0
depletion, nFe(II) number of moles of corroded Fe
0 , nFe3O4 number of moles of generated iron corrosion products, nelectrons=2*nFe(II)
number of electrons released by corroded iron, nnano/nelectrons the ratio of the number of electrons produced in by nano-Fe
0 to nelectrons
1370 Water Air Soil Pollut (2012) 223:1363–1382
Table 6 also shows that the maximum value of the
relative time (C ) is 10 (or 101). This is more practical
for graphical representations than situations where
nano-Fe0 are compared with larger particles (t≤104).
The physical significance of C is more important, it
means that if a nano-Fe0 with a diameter of 10 nm
depletes after 2 days, the material with a diameter
100 nm will deplete after 20 days. For field
applications, the selection of the particle size to be
used should be dictated by site specific character-
istics. Which diameter could quantitatively reach the
contaminants before depletion? And what fraction of
the material will have already oxidized on the path?
What is the impact of this oxidation on the transport
of nano-Fe0 in the porous aquifer? These are some
key questions to be answered in order to give this
possibly very efficient technology a scientific basis.
Figure 3 summarizes the evolution of the volumet-
ric expansion in all five nano-Fe0 systems. It can be
seen from Fig. 3a that the smallest material (d=10 nm)
experiences the 108% volumetric expansion within a
short time (C =1) while the larger materials (d=
100 nm) needs ten more time for the same expansion.
Accordingly, beside the question whether the material
will reach the contaminant under site specific con-
ditions, the question has to be answered how the
volumetric expansion will impact the aquifer porosity
(and permeability).
Figure 3b compares the variation of the volumetric
expansion for two different iron corrosion products,
Fe3O4 and FeOOH, which are considered the most
likely products in anoxic and oxic aquifers respec-
tively. It also shows that when designing a nano-Fe0
injection strategy, however, the availability of oxidiz-
Table 6 Summary of the values of the number of particles contained in 1 kg of each nano-Fe0, the number of layer making up each
particle and estimation of the relative time (C )
d (nm) nparticles (−) nlayers (−) nlayers/n10 (−) n10/nparticles (−) C (−)
10 3.06*1019 34.9 1.0 1.0 1.0
25 1.96*1018 87.2 2.5 16 2.5
50 2.45*1011 174.5 5.0 125 5.0
75 7.25*1016 261.7 7.5 422 7.5
100 3.06*1016 348.9 10.0 1,000 10.0
The life span of the material with the smallest particle site (d=10 nm) is operationally considered as the unit of time while assuming
uniform corrosion. The ratio of the number of particles in 1 kg of d=10 nm to that of other d values is also given
0 1 2 3 4 5 6 7 8 9 10
0
20
40
60
80
100
120
(a)
 10 nm
 25 nm
 50 nm
 75 nm
 100 nm
v
o
l. 
ex
pa
ns
io
n 
/ [%
]
relative time / [-]
0 20 40 60 80 100
160
200
240
280
320
360
400
440
69 %
30 %
108 %
250 %
(b)
 Fe3O4
 FeOOH
ΔV
 / 
[m
L]
particle size / [nm]
Fig. 3 Kinetics of the process of porosity loss at nanoscale as
characterized by: a the percent volumetric expansion for the
five considered particle sizes, and b the absolute value of ΔV
(mL) at C =1 for two different iron corrosion products (Fe3O4
and FeOOH)
Water Air Soil Pollut (2012) 223:1363–1382 1371
ing species (e.g. MnO2 and O2) must also to be taken
into account. Figure 3a shows that under both
conditions the trend of porosity loss is similar but
the extent is proportional to the coefficient of
volumetric expansion (η). In particular, at C =1, the
system with the material d=10 nm experiences 250%
volumetric expansion under oxic conditions and only
110% under anoxic conditions. As a result the
kinetics of more rapid Fe0 corrosion in an oxygen-
rich environment must also be considered for an
effective treatment strategy.
4.2 Fe0 Reactivity of Nano-bimetallics
The reactivity of monometallic nano-Fe0 can be
improved by combining it with a noble metal.
Assuming α (α>1) the coefficient of reactivity
enhancement, the relation between the relative time
of a bimetallic system (CFe/M) and that of a non-plated
metal (CFe) is given by Eq. 6:
tFe ¼ a»tFe=M ð8Þ
To characterize the impact of plating on nano-Fe0,
the material with the largest size (d=100 nm) will be
plated by three hypothetical metals (M1
0, M2
0 and
M3
0) to yield a reactivity factor of 2.5 (for Fe0/M1
0), 5
(for Fe0/M2
0) and 10 (for Fe0/M3
0). The considered α
values of (α≤10) are realistic and even conservative.
In fact, reported reactivity enhancement is essentially
larger (Chen et al. 2011; Zhuang et al. 2011). For
example, Zhuang et al. (2011) reported that palladized
nano-Fe0 promoted the dehalogenation kinetics for
polybrominated diphenyl ethers by orders of magni-
tude equal to 2, 3 and 4 (α≥100). The results of the
calculations for the four systems (d=10 nm) are
summarized in Fig. 4. The system with d=10 nm is
represented for comparison. It can be seen that system
Fe0/M3
0 (d=100 nm) is as reactive as Fe0 (d=10 nm).
Given that the reactivity of nano-Fe0 (d=100 nm)
could already significantly been too high in some
situations, the results from Fig. 4 strongly question
the suitability of plating at nanoscale. Accordingly,
while the application of bimetallic Fe0 is definitively
useful at μm and mm scale, it usefulness at nanoscale
is likely inappropriate. It can also be noted that by
increasing the reactivity of the material the rate at
which volumetric pore clogging also increases. As a
consequence it should be acknowledged that there
exists a conceptual play-off between increased reac-
tion rate and increased porosity loss, the impact of
which will vary depending on the physiochemical
conditions of each contaminated site.
4.3 Characterizing the Process of Reactivity Loss
To better characterize the process of porosity loss due
to the volumetric expansion of nano-Fe0, the evolu-
tion of the porosity of a sand column filled with nano-
Fe0 will be discussed as volumetric expansion
proceeds. A laboratory column with a height h (h=
75.0 cm) and diameter D (D=5.0 cm) is composed of
spherical sand particles (d=5.0 mm). The compact-
ness of the column is ideally C=0.64 (Noubactep and
Caré 2010a, Noubactep et al. 2010). The pore volume
is given by Eq. 9:
Vpore ¼ V» 1 Cð Þ ð9Þ
where V is the apparent volume of the sand column (V=
h*π*D2/4).
It is supposed that the nano-Fe0 particles fill the
inter-granular porosity of the sand column Vpore
without modifying the compactness C and the apparent
volume V of the sand column. The residual porosity of
the sand column (V′pore) is given by Eq. 10:
V 0pore ¼ V» 1 Cð Þ  V0 ð10Þ
where V0 is the volume of the Fe particles.
The evolution of the residual porosity (V′pore) as
nano-Fe0 particles undergo volumetric expansive cor-
0 2 4 6 8 10
0
20
40
60
80
100
 Fe0(d = 100 nm)
 Fe0/M01
 Fe0/M02
 Fe0/M03
 Fe0(d = 10 nm)Fe
0  
co
n
su
m
ed
 / 
[%
]
relative time / [-]
Fig. 4 Calculated extent of Fe0 exhaustion as a function of the
relative time (C ) for three ideal bimetallic systems based on the
material with 100 nm diameter. The material with 10 nm
diameter is represented for comparison
1372 Water Air Soil Pollut (2012) 223:1363–1382
rosion is considered by introducing the specific volume
(η) of the reaction products according to Eq. 11:
V
0
pore ¼ V» 1 Cð Þ  V0  Vconsumed Feð Þ  h»Vconsumed Fe
ð11Þ
V
0
pore ¼ V» 1 Cð Þ  V0  h 1ð Þ»Vconsumed Fe
ð11aÞ
where Vconsumed Fe is the volume of nano-Fe
0 particles
which is consumed at a given time.
Equations 9 through 11 are very useful to design
reactive zone. However, they are limited to describe
the initial (Vpore) and a final conditions (V′pore)
regardless on the nature of iron corrosion products
and the kinetics of the process.
Using a sand column comparable to one of those
used by Moraci and Calabrò (2010) and 1 kg of nano-
Fe0, the process of pore filling (porosity loss) can be
better characterized. For simplicity, nano-Fe0 consid-
ered as transported by a biodegradable dispersant
which does not significantly contribute to porosity
loss. As shown in Section 3.3, 1 kg of nano-Fe0
occupies a volume of 127 mL. The initial pore
volume of the sand column calculated after Eq. 9 is
530.36 mL (100% porosity), i.e. a capacity for
approximately 4.17 kg of nano-Fe0. Filling the initial
pore volume of the sand column (530.36 mL) with
1 kg of nano-Fe0 (127.00 mL) yields a 23.9%
porosity loss (Table 7). This, however, does not take
into account the expansive nature of iron during
oxidative corrosion.
Using the eight possible iron corrosion products
documented by Caré et al. (2008) and their respective
coefficient of volumetric expansion (2.08≤η≤6.40),
the extent of porosity loss is calculated and summa-
rized in Table 7. The results show that the residual
volume of pores (V′pore) decreases with increasing η
values and is zero for Fe(OH)3 and Fe(OH)3 3H2O
(100% porosity loss). Ferrihydrite (Fe(OH)3 3H2O) is
the largest known iron corrosion products. In other
words, depending on environmental conditions as
little as 1 kg of nano-Fe0 could clog the tested
column. Although this discussion considers the nature
of the corrosion products, there are other important
factors which must be considered. The negative
values (−3.04 and −282.4 mL) corresponds to the
mass of Fe0 which will not oxidized because of lack
of space for expansion (Noubactep and Caré 2010a;
Noubactep et al. 2010).
The extent of porosity loss (ΔV in %) given in
Table 7 assumes uniform distribution of nano-Fe0 in
the whole column. This is, however, not a very good
field representation. For example, if 1 kg of nano-Fe0
(V0=127 mL) is uniformly distributed only in the first
third of the column (Vpore
1/3=176.8 mL), with Fe3O4
as the primary corrosion product (ΔV=137.16 mL) a
78% porosity loss can be expected. For all other oxide
phases it is calculated that complete porosity loss
(100%) will precede nano-Fe0 reactive exhaustion.
However, in the practice a system with 78% porosity
Table 7 Summary of the extent of porosity loss (ΔVpore in %) as 1 kg of nano-Fe
0 (V0=127 mL) is corroded to various iron oxides
Fe species η (−) V∞ (mL) ΔV (mL) V′pore (mL) ΔVpore (%)
Fe0 1 127 0 403.4 23.9
Fe3O4 2.08 264.2 137.2 266.2 49.8
Fe2O3 2.12 269.2 142.3 261.1 50.8
α-FeOOH 2.91 369.6 242.6 160.8 69.7
γ-FeOOH 3.03 384.8 258.0 145.6 72.6
β-FeOOH 3.48 442.0 315.0 88.4 83.3
Fe(OH)2 3.75 476.3 349.3 54.1 89.8
Fe(OH)3 4.2 533.4 406.4 −3.0 100.0
Fe(OH)3 3H2O 6.4 812.8 685.8 −282.4 100.0
V∞ is the volume of iron oxide at Fe
0 exhaustion. The initial volume of pore (Vpore) is 530.4 mL and V′pore is the residual pore volume
at Fe0 exhaustion. The absolute value of negative values for V′pore corresponds to the mass of nano-Fe
0 which cannot oxidize because
of lack of space for volumetric expansion
Water Air Soil Pollut (2012) 223:1363–1382 1373
loss is considered as clogged. One possibility to avoid
the clogging of the entrance zone of a porous system
is to intermittently inject calculated amounts of nano-
Fe0. The volume to be injected at each event and the
time scale between two injections are necessarily
determined by site specific characteristics (e.g. poros-
ity of the aquifer and water flow rate).
5 Discussion
A primary reason behind the interest into the use of
nano-Fe0 particles over μm-Fe0 and mm-Fe0 par-
ticles for water treatment is ascribed to a significant
increase the materials efficiency (Comba et al. 2011;
Gheju 2011). For example, as reported by Vodyanitskii
(2010), Kanel et al. (2006) reported near-total
remediation of a 1 mg L−1 AsV solution within only
10 min by a nano-Fe0 with a specific surface of
24 m2/g, whereas the same goal was achieved by
mm-Fe0 (1–2 m2/g) after only 4 days or 5,760 min
(ratio of time, 570; average ratio of surface, 16).
However, this experimental evidence is highly
qualitative as neither the number of atoms directly
accessible at the surface nor the intrinsic reactivity of
individual materials are considered in both cases
(Noubactep and Caré 2010b). For a better compar-
ative result, the following three key conditions must
be considered: (1) the intrinsic Fe0 reactivity should
be characterized, (2) the amount of used materials
should be based on the reaction stoichiometry and
(3) the experimental conditions should be relevant
for field applications. In particular, the driving force
for the transport of contaminants and Fe species
should be relevant for field situations: (1) mixing
operation (type and intensity) in batch studies and
(2) flow rate and column dimensions in column
studies (Noubactep et al. 2009).
5.1 Transport of Nano-Fe0 to the Contaminants
The efficiency of nano-Fe0 for the in situ treatment of
a contaminated aquifer body is intrinsically linked to
the extent of physical contact between Fe0 and any
aqueous contaminant species present. In some cir-
cumstances, contaminants could diffuse to the sus-
pended Fe0 particles and be degraded in the aqueous
phase. However, typically the suspended Fe0 particles
must migrate to the contaminants. As Fe0 particles are
transported from the injection zone to the target
contaminant plume by natural groundwater, diffusion
experiments under relevant groundwater velocity,
using site specific aquifer materials are essential in
order to effectively assess the suitability of nano-Fe0
for in situ applications (Antia 2011).
Contaminants are typically partitioned between
sediment and water phases in a “pseudo-equilibrium”
state. Therefore, it is likely that Fe0 particles whilst
acting to reduce any soluble contaminants are also
likely to promote the dissolution of a range of
adsorbed chemical species (Le Chatelier’s principle).
However, as water is also a redox-amenable species
the specific mechanism for nano-Fe0 reactivity in a
range of conditions is difficult to resolve (Tratnyek
and Johnson 2006). In other words, nano-Fe0 is
readily oxidized by H2O during subsurface migration
to the target contaminant plume and also competes
with any other redox-amenable (including contami-
nants) present in the groundwater. Additionally,
expansive iron corrosion will yield voluminous iron
(hydr)oxides (Table 7) with limited aqueous mobility
due to (1) an increased size and weight and (2) a
possible increased affinity to aquifer material.
This discussion has intentionally neglected the
segregation between parent compounds, the reaction
products and their relative affinity to Fe0 and Fe
(hydro)oxides. The fact that the core Fe0 is always
covered by oxide layers is also neglected for
simplification. The process of preferential flow which
is crucial in predicting mass transfer in the subsurface
is also not considered (Flury and Flühler 1994; Fryar
and Schwartz 1998; Simunek et al. 2003; Clothier et
al. 2008; Allaire et al. 2009). However, it is clearly
shown, that due the acute redox sensitivity of nano-
Fe0 and the subsequent significant formation of
highly voluminous oxidative corrosion products it is
likely that for environmentally relevant distances
(metre), a significant proportion of the originally
injected nano-Fe0 will remain “clogged” in pore
spaces.
Beside the transport of nano-Fe0 to contami-
nants, the possibility of quantitative contaminant
desorption and their subsequent transformation by
suspended Fe0 could be considered. However, it is
not likely that concentration-gradient-driven mass
transfer could be quantitative at considered distances
(metre). It should be recalled that the slow kinetics
of contaminant desorption form aquifer materials is
1374 Water Air Soil Pollut (2012) 223:1363–1382
the major cause of the ineffectiveness of the pump-
and-treat technology for groundwater remediation
(McMurty and Elton 1985; Mackay and Cherry
1989; Starr and Cherry 1994).
This section has shown that it is likely that the
success of nano-Fe0 for in situ remediation is
seriously limited by the intrinsic formation of volumi-
nous iron corrosion products (Tratnyek and Johnson
2006; Comba et al. 2011; Gheju 2011; Truex et al.
2011a, b). Bearing this in mind, the next section
suggests an alternative nano-Fe0 subsurface deploy-
ment mechanism that more effectively takes into
account the aforementioned nano-Fe0 hydraulic mobil-
ity issues than conventional injection processes: the
formation of a nano-Fe0 “redox front” injection array
system for progressive contaminant reduction. The
geochemical process of redox front migration is a well-
documented one (Fryar and Schwartz 1998; Min et al.
2005; Sidborn and Neretnieks 2007).
5.2 Nano-Fe0 as Source of FeII for a Redox Front?
5.2.1 The Concept
The progressive consumption of mm-Fe0 (Fig. 1;
Table 3) is the guarantee for the long-term efficiency
of reactive barriers (Comba et al. 2011). In fact,
continuously generated small amount of high reactive
iron minerals (Gu et al. 1999; Su and Puls 2001;
Furukawa et al. 2002; Kohn et al. 2005; Noubactep
2007, 2008, 2010c, 2011b, c, d) are sufficient for the
removal of contaminants which are present in trace
amounts (Palmer and Wittbrodt 1991). As discussed
above, for nano-Fe0 however, (1) Fe0 reactive
exhaustion typically occurs in a relatively short time
scale (<9 h) and (2) it is likely that nano-Fe0
subsurface mobility is significantly retarded or even
prevented due to the volumetric expansive nature of
iron corrosion (Zhao et al. 2011). As a consequence
an alternative method of subsurface deployment is
suggested in the current work: the deployment of a
linear nano-Fe0 injection array orientated perpendic-
ular to the flow direction of the contaminant plume.
The injected nano-Fe0 can effectively form a redox
front (roll-front) which migrates through the contam-
inated zone and transforms the contaminants during
its migration as illustrated in Fig. 5. The FeII/FeIII roll-
front travels across the contaminated zone with all
possible mechanisms (e.g. diffusion, dispersion, con-
vection and preferential flow) and the contaminants
are transformed and immobilized during the cycle
FeII⇔FeIII. In other words, it is a plume of FeII/FeIII
formed from injected nano-Fe0 which migrates
through the contaminated zone and “sweeps” the
contaminants. As a consequence this method considers
all nano-Fe0 mobility issues.
5.2.2 Nano-Fe0 as FeII Generator
Nano-Fe0 in the aqueous phase is certainly a FeII/
FeIII producer. FeII species are the main reducing
agents for contaminants under both anoxic and oxic
conditions (Stratmann and Müller 1994; Nesic 2007;
Kiser and Manning 2010; Noubactep 2010c, 2011c,
d; Zhuang et al. 2011). Microbial activity could
regenerate FeII (bio-corrosion) for more contaminant
reduction (Vodyanitskii 2010). In this case, more
contaminant is reduced than can be predicted from
the reaction stoichiometry. In order words, the
operating mode of nano-Fe0 for contaminant reduc-
tion can be summarized as follows: (1) Fe0 is
oxidized to produce FeII, (2) FeII reduces the
contaminant and is oxidized to FeIII and (3) a
proportion of FeII is regenerated by the biological
reduction of FeIII. Accordingly, before Fe0 depletion,
there are three sources of FeII: (1) the Fe0-mediated
abiotic oxidation by H2O, (2) the Fe
0-mediated
abiotic oxidation by FeIII and (3) the biological
reduction of FeIII. After Fe0 depletion, the only
remaining source of FeII is the biological reduction
of FeIII. Provided that the appropriate micro-
organism species are present in the subsurface, this
process, however, could conceptually proceed for a
significantly long-time period (Cullen et al. 2011).
Evidence suggests that such micro-organism colo-
nies can be sustained by a consistent supply of FeII,
FeIII and molecular hydrogen (H/H2). Another further
process that is worth noting is the generation of
atomic or molecular hydrogen (H/H2) by Fe
0-mediat-
ed hydrolysis reactions, which is likely to aid and the
aforementioned biotic processes (Cullen et al. 2011).
The abiotic conversion of FeIII to FeII has been
successfully utilised in the hydrometallurgy industry,
for example Lottering et al. (2008) reported on the
sustainable use of MnO2 for the abiotic regeneration
of FeIII for UIV oxidation.
The fate of contaminant reduction products is
discussed in the next section.
Water Air Soil Pollut (2012) 223:1363–1382 1375
5.3 Mechanism of Contaminant Removal by Injected
Nano-Fe0
The successful application of nano-Fe0 injection
technology for in situ remediation is highly dependent
on a comprehensive understanding of the fundamental
processes governing the processes of contaminant
removal. The hitherto discussion has focused on
reductive transformations by nano-Fe0. However,
contaminant reductive transformation is not a guaran-
tee for contaminant removal (Noubactep 2010c,
2011c). Additionally, certain reaction products are
more toxic than their parent compounds (Jiao et al.
2009). Accordingly, efforts have to be focused on the
specific mechanism of aqueous contaminant removal.
Relevant removal processes include: (1) adsorption,
(2) chemical precipitation, (3) co-precipitation, (4)
size exclusion or straining and (5) volatilization
(Crawford et al. 1993a, b; Noubactep 2007, 2008,
2010c; Vodyanitskii 2010; Eusterhues et al. 2011;
Fig. 5 Schematic diagram
of the flow process of the
U-shaped redox front
through a contaminated
zone. Despite the relative
importance of preferential
flow paths, contaminants
are “swept” by the
roll-front. The red circle
is the contaminated zone.
This zone is transformed
to blue the circle upon
treatment by the
FeII/FeIII front
1376 Water Air Soil Pollut (2012) 223:1363–1382
Johnson et al. 2011; Noubactep 2011d, e). Chemical
precipitation is a characteristic of inorganic compounds
when the solubility limit is exceeded (Crawford et al.
1993a, b; Kalin et al. 2005). Volatilization is subse-
quent to chemical transformation yielding gaseous
species like AsH3, CH4, CO2, H2 and N2. In Fe
0
reactive barrier systems, contaminants are efficiently
removed by the combination of adsorption, co-
precipitation and size exclusion within the engineered
barrier (Noubactep 2010d; Noubactep and Schöner
2010; Noubactep 2011a). As a result the current
discussion concentrates on such processes.
With Fe0 (<1 m2/g) first transformed to volumi-
nous hydroxides species (>100 m2/g) and subsequent-
ly transformed to oxides (<40 m2/g), contaminant size
exclusion (straining) is driven by the dynamic cycle
of expansion/compression accompanying the cor-
rosion process (Pokhrel and Viraraghavan 2008;
Vodyanitskii 2010; Noubactep 2011a). During these
cycles contaminants are enmeshed and sequestrated
in a “matrix” of iron corrosion products.
For conventional nano-Fe0 injection arrays, size
exclusion may play an important role (1) in proximity
for Fe0 particles and (2) by reducing the pore space
during expansive corrosion of the materials. However,
if roll-fronts are formed as discussed above, the extent
of permeability loss in aquifer will be limited. The
roll-front could act as a colloidal reactive barrier for
the removal of parent contaminants and reaction
products. Species are removed or immobilized by
colloids and not because they are reduced. More
research is needed to test this hypothesis.
5.3.1 FeII/FeIII Redox Front as a Colloidal Reactive
Barrier
Aqueous contaminants have been reported to be
quantitatively removed both during abiotic and biotic
(1) oxidation of FeII and (2) reduction of FeIII
(Pokhrel and Viraraghavan 2008; 2009; Pokhrel et
al. 2009; Vodyanitskii 2010). On the other hand,
injection of FeIII salts for adsorptive contaminant
removal has been reported (Morrison and Sprangler
1993; Morrison et al. 1996). Accordingly, the migra-
tion of the FeII/FeIII redox front may be coupled to
quantitative contaminant removal by adsorption and
co-precipitation.
The primary reason for contaminant removal
during these redox reactions is the colloidal nature
of in situ generated Fe species (Fe(OH)2 and Fe
(OH)3) (Hanna and Boily 2010), which necessarily
experience volumetric contraction to form oxides (of
FeII or FeIII). Contaminants are first adsorbed by
highly reactive colloids and are co-precipitated during
transformation to amorphous and crystalline oxides
(Ghauch et al. 2010; Noubactep 2011c, d).
6 Concluding Remarks
Constructed geochemical barriers of metallic iron
(Fe0) have been used for groundwater remediation
since 1996 (O’Hannesin and Gillham 1998; Scherer et
al. 2000; Henderson and Demond 2007; Gillham
2010, Comba et al. 2011; Gheju 2011). In recent
years, however, nano-Fe0 has received proclaim as a
new tool for water treatment due to (1) improvements
in reactivity and associated aqueous contaminant
removal performance compared with conventional
materials and (2) the option of subsurface deployment
via injection for targeted in situ treatment of contam-
inant plumes (Comba et al. 2011; Gheju 2011).
Considering reactivity first, the current work has
highlighted the need for prudent use of terminology.
Discounting any quantum size effects, which are only
prevalent for Fe0 less than approximately 10 nm in
diameter, the reactivity of nano-Fe0 as a function of
surface area is no more reactive than larger forms.
Nano-Fe0 only exhibits such high reactivity due to it
significantly high surface area as a function of mass/
volume. Despite this, a recent trend in research has
been the development of bimetallic nano-Fe0 wherein
the combination of a noble metal acts to further
increase the reactivity of nano-Fe0. It is argued in the
current work that as reactive exhaustion is already
achieved by monometallic nano-Fe0 in the order of
minutes this seems counterintuitive for the majority of
environmental applications.
Considering the nano-Fe0 subsurface injection
procedure, in the current work it has been highlighted
that the hydraulic mobility of the particles is likely to
be significantly retarded by voluminous expansion
due to particle corrosion. An alternative nano-Fe0
injection procedure has been suggested herein. The
injected nano-Fe0 effectively forms an in situ migrat-
ing front which possibly reductively transforms
contaminant and removes reduced species by adsorp-
tion and co-precipitation.
Water Air Soil Pollut (2012) 223:1363–1382 1377
It is also outlined in the current work that a number
of studies with experiments “proclaimed” as analo-
gous to environmental systems are largely overlooked
a range of operational drivers including changes in
nano-Fe0 (1) reactivity and (2) voluminous as a
function of time. It is hoped that the huge literature
on redox front migration (Reynolds and Goldhaber
1978; Posey-Dowty et al. 1987; Romero et al. 1992;
Read et al. 1993) and the cycle of iron in the
hydrosphere (Vodyanitskii 2010) will now be used
for the further development of nano-Fe0 injection
technology.
Acknowledgements Mohammad A. Rahman (Angewandte
Geologie-Universität Göttingen) is acknowledged for technical
support.
Open Access This article is distributed under the terms of the
Creative Commons Attribution Noncommercial License which
permits any noncommercial use, distribution, and reproduction
in any medium, provided the original author(s) and source are
credited.
References
Agarwal, A., & Joshi, H. (2010). Environmental sciences
application of nanotechnology in the remediation of
contaminated groundwater: a short review. Recent
Research in Science and Technology, 2, 51–57.
Allaire, S. E., Roulier, S., & Cessna, A. J. (2009). Quantifying
preferential flow in soils: a review of different techniques.
Journal of Hydrology, 378, 179–204.
Anstice, C., Alonso, C., & Molina, F. J. (1993). Cover cracking
as a function of bar corrosion: part I—experimental test.
Materials and Structures, 26, 453–464.
Antia, D. D. J. (2011). Modification of aquifer pore-water by
static diffusion using nano-zero-valent metals. Water12, 3,
79–112.
Barnes, R. J., van der Gast, C. J., Riba, O., Lehtovirta, L. E.,
Prosser, J. I., Dobson, P. J., et al. (2010). The impact of
zero-valent iron nanoparticles on a river water bacterial
community. Journal of Hazardous Materials, 184, 73–80.
Bartzas, G., & Komnitsas, K. (2010). Solid phase studies and
geochemical modelling of low-cost permeable reactive
barriers. Journal of Hazardous Materials, 183, 301–308.
Behrens, S. H., Christl, D. I., Emmerzael, R., Schurtenberger,
P., & Borkovec, M. (2000). Charging and aggregation
properties of carboxyl latex particles: experiments versus
DLVO theory. Langmuir, 21, 2566–2575.
Blowes, D. W., Ptacek, C. J., & Jambor, J. L. (1997). In-situ
remediation of Cr(VI)-contaminated groundwater using
permeable reactive walls: laboratory studies. Environmental
Science & Technology, 31, 3348–3357.
Boparai, H. K., Joseph, M., & O’Carroll, D. M. (2011).
Kinetics and thermodynamics of cadmium ion removal
by adsorption onto nano zerovalent iron particles. Journal
of Hazardous Materials, 186, 458–465.
Caré, S., Nguyen, Q. T., L’Hostis, V., & Berthaud, Y. (2008).
Mechanical properties of the rust layer induced by
impressed current method in reinforced mortar. Cement
and Concrete Research, 38, 1079–1091.
Celebi, O., Uzum, C., Shahwan, T., & Erten, H. N. (2007). A
radiotracer study of the adsorption behavior of aqueous
Ba2+ ions on nanoparticles of zero-valent iron. Journal of
Hazardous Materials, 148, 761–767.
Chen, K.-F., Li, S., & Zhang, W.-X. (2011). Renewable
hydrogen generation by bimetallic zerovalent iron
nanoparticles. Chemical Engineering Journal, 170,
562–567.
Clothier, B. E., Green, S. R., & Deurer, M. (2008). Preferential
flow and transport in soil: progress and prognosis.
European Journal of Soil Science, 59, 2–13.
Comba, S., Di Molfetta, A., & Sethi, R. (2011). A comparison
between field applications of nano-, micro-, and milli-
metric zero-valent iron for the remediation of contaminat-
ed aquifers. Water, Air, and Soil Pollution, 215, 595–607.
Courcelles, B., Modaressi-Farahmand-Razavi, A., Gouvenot,
D., & Esnault-Filet, A. (2011). Influence of precipitates on
hydraulic performance of permeable reactive barrier filters.
International Journal of Geomechanics, 11, 142–151.
Crane, R. A., Dickinson, M., Popescu, I. C., & Scott, T. B.
(2011). Magnetite and zero-valent iron nanoparticles for
the remediation of uranium contaminated environmental
water. Water Research, 45, 2931–2942.
Crawford, R. J., Harding, I. H., & Mainwaring, D. E. (1993a).
Adsorption and coprecipitation of single heavy metal ions
onto the hydrated oxides of iron and chromium. Langmuir,
9, 3050–3056.
Crawford, R. J., Harding, I. H., & Mainwaring, D. E. (1993b).
Adsorption and coprecipitation of multiple heavy metal
ions onto the hydrated oxides of iron and chromium.
Langmuir, 9, 3057–3062.
Cullen, L. G., Tilston, E. L., Mitchell, G. R., Collins, C. D., &
Shaw, L. J. (2011). Assessing the impact of nano- and
micro-scale zerovalent iron particles on soil microbial
activities: Particle reactivity interferes with assay condi-
tions and interpretation of genuine microbial effects.
Chemosphere, 82, 1675–1682.
Diao, M., & Yao, M. (2009). Use of zero-valent iron nano-
particles in inactivating microbes. Water Research, 43,
5243–5251.
Dickinson, M., & Scott, T. B. (2010). The application of zero-
valent iron nanoparticles for the remediation of a uranium-
contaminated waste effluent. Journal of Hazardous Mate-
rials, 178, 171–179.
Eusterhues, K., Rennert, T., Knicker, H., Kgel-Knabner, I.,
Totsche, K. U., & Schwertmann, U. (2011). Fractionation
of organic matter due to reaction with ferrihydrite:
coprecipitation versus adsorption. Environmental Science
& Technology, 45, 527–533.
Fang, Z., Qiu, X., Chen, J., & Qiu, X. (2011). Degradation of
the polybrominated diphenyl ethers by nanoscale zero-
valent metallic particles prepared from steel pickling waste
liquor. Desalination, 267, 34–41.
Fernando, W. J. N., Ahmad, A. L., Abd. Shukor, S. R., & Lok,
Y. H. (2008). A model for constant temperature drying
1378 Water Air Soil Pollut (2012) 223:1363–1382
rates of case hardened slices of papaya and garlic. Journal
of Food Engineering, 88, 229–238.
Flury, M., & Flühler, H. (1994). Brilliant Blue FCF as a dye
tracer for solute transport studies. A toxicological review.
Journal of Environmental Quality, 23, 1108–1112.
Fryar, A. E., & Schwartz, F. W. (1998). Hydraulic-conductivity
reduction, reaction-front propagation, and preferential flow
within a model reactive barrier. Journal of Contaminant
Hydrology, 32, 333–351.
Furukawa, Y., Kim, J.-W., Watkins, J., & Wilkin, R. T.
(2002). Formation of ferrihydrite and associated iron
corrosion products in permeable reactive barriers of
zerovalent iron. Environmental Science & Technology, 36,
5469–5475.
Ghauch, A., Tuqan, A., & Abou, A. H. (2009). Antibiotic
removal from water: elimination of amoxicillin and
ampicillin by microscale and nanoscale iron particles.
Environmental Pollution, 157, 1626–1635.
Ghauch, A., Abou, A. H., & Bdeir, S. (2010). Aqueous removal
of diclofenac by plated elemental iron: bimetallic systems.
Journal of Hazardous Materials, 182, 64–74.
Gheju, M. (2011). Hexavalent chromium reduction with zero-
valent iron (ZVI) in aquatic systems. Water, Air, and Soil
Pollution. doi:10.1007/s11270-011-0812-y.
Gillham, R. W. (2003). Discussion of Papers/Discussion of
nano-scale iron for dehalogenation (by Evan K. Nyer and
David B. Vance (2001), Ground Water Monitoring &
Remediation, 21(2), pp. 41–54). Ground Water Monitor-
ing & Remediation, 23, 6–8.
Gillham, R.W. (2010). Development of the granular iron
permeable reactive barrier technology (good science or
good fortune). In Y. Chen, X. Tang, L. Zhan (Eds.),
“Advances in environmental geotechnics: proceedings of
the International Symposium on Geoenvironmental Engi-
neering in Hangzhou, China, September 8–10, 2009” (pp
5–15). London: Springer.
Gillham, R. W., & O’Hannesin, S. F. (1991). Metal-catalysed
abiotic degradation of halogenated organic compounds.
Ground Water, 29, 752–761.
Gillham, R. W., & O’Hannesin, S. F. (1994). Enhanced
degradation of halogenated aliphatics by zero-valent iron.
Ground Water, 32, 958–967.
Grieger, K. D., Fjordboge, A., Hartmann, N. B., Eriksson, E.,
Bjerg, P. L., & Baun, A. (2010). Environmental benefits
and risks of zero-valent iron nanoparticles (nZVI) for in
situ remediation: Risk mitigation or trade-off? Journal of
Contaminant Hydrology, 118, 165–183.
Gu, B., Phelps, T. J., Liang, L., Dickey, M. J., Roh, Y., Kinsall,
B. L., et al. (1999). Biogeochemical dynamics in zerova-
lent iron columns: implications for permeable reactive
barriers. Environmental Science & Technology, 33, 2170–
2177.
Hanna, K., & Boily, J.-F. (2010). Sorption of two naphthoic
acids to goethite surface under flow through conditions.
Environmental Science & Technology, 44, 8863–8869.
Henderson, A. D., & Demond, A. H. (2007). Long-term
performance of zero-valent iron permeable reactive barriers:
a critical review. Environmental Engineering Science, 24,
401–423.
Jeen, S.-W., Gillham, R. B., & Przepiora, A. (2011). Predictions
of long-term performance of granular iron permeable
reactive barriers: field-scale evaluation. Journal of Con-
taminant Hydrology, 123, 50–64.
Jiang, Z., Lv, L., Zhang, W., Du, Q., Pan, B., Yang, L., et al.
(2011). Nitrate reduction using nanosized zero-valent iron
supported by polystyrene resins: role of surface functional
groups. Water Research, 45, 2191–2198.
Jiao, Y., Qiu, C., Huang, L., Wu, K., Ma, H., Chen, S., et al.
(2009). Reductive dechlorination of carbon tetrachloride
by zero-valent iron and related iron corrosion. Applied
Catalysis B: Environmental, 91, 434–440.
Johnson, R. L., Thoms, R. B., Johnson, R. O. B., Nurmi, J., &
Tratnyek, P. G. (2008). Mineral precipitation upgradient
from a zero-valent iron permeable reactive barrier. Ground
Water Monitoring & Remediation, 28, 56–64.
Johnson, W. P., Ma, H., & Pazmino, E. (2011). Straining
credibility: a general comment regarding common argu-
ments used to infer straining as the mechanism of colloid
retention in porous media. Environmental Science &
Technology, 45, 3831–3832.
Kadu, B. S., Sathe, Y. D., Ingle, A. B., Chikate, R. C., Patil, K.
R., & Rode, C. V. (2011). Efficiency and recycling
capability of montmorillonite supported Fe–Ni bimetallic
nanocomposites towards hexavalent chromium remedia-
tion. Applied Catalysis B: Environmental, 104, 407–414.
Kalin, M., Wheeler, W. N., & Meinrath, G. (2005). The
removal of uranium from mining waste water using algal/
microbial biomass. Journal of Environmental Radioactiv-
ity, 78, 151–177.
Kanel, S. R., Greneche, J.-M., & Choi, H. (2006). Arsenic(V)
Removal from groundwater using nano scale zero-valent
iron as a colloidal reactive barrier material. Environmental
Science & Technology, 40, 2045–2050.
Karn, B., Kuiken, T., & Otto, M. (2009). Nanotechnology and
in situ remediation: a review of the benefits and potential
risks. Environmental Health Perspectives, 117, 1832–
1831.
Kim, J. Y., Park, H.-J., Lee, C., Nelson, K. L., Sedlak, D. L., &
Yoon, J. (2010). Inactivation of Escherichia coli by
nanoparticulate zerovalent iron and ferrous ion. Applied
and Environmental Microbiology, 76, 7668–7670.
Kiser, J. R., & Manning, B. A. (2010). Reduction and
immobilization of chromium(VI) by iron(II)-treated fauja-
site. Journal of Hazardous Materials, 174, 167–174.
Kohn, T., Kenneth, J. T., Livi, A., Roberts, A. L., & Vikesland,
P. J. (2005). Longevity of granular iron in groundwater
treatment processes: corrosion product development. En-
vironmental Science & Technology, 39, 2867–2879.
Korte, N. E., Zutman, J. L., Schlosser, R. M., Liang, L., Gu, B.,
& Fernando, Q. (2000). Field application of palladized
iron for the dechlorination of trichloroethene. Waste
Management, 20, 687–694.
Li, L., & Benson, C. H. (2010). Evaluation of five strategies to
limit the impact of fouling in permeable reactive barriers.
Journal of Hazardous Materials, 181, 170–180.
Li, X. Q., & Zhang, W. X. (2007). Sequestration of metal
cations with zerovalent iron nanoparticles—a study with
high resolution X-ray photoelectron spectroscopy (HR-
XPS). Journal of Physical Chemistry C, 111, 6939–
6946.
Li, X. Q., Elliott, D. W., & Zhang, W. X. (2006). Zero-valent
iron nanoparticles for abatement of environmental pollu-
Water Air Soil Pollut (2012) 223:1363–1382 1379
tants: materials and engineering aspects. Critical Reviews
in Solid State and Materials Sciences, 31, 111–122.
Litter, M. I., Morgada, M. E., & Bundschuh, J. (2010). Possible
treatments for arsenic removal in Latin American waters
for human consumption. Environmental Pollution, 158,
1105–1118.
Lottering, M. J., Lorenzen, L., Phala, N. S., Smit, J. T., &
Schalkwyk, G. A. C. (2008). Mineralogy and uranium
leaching response of low grade South African ores.
Minerals Engineering, 21, 16–22.
Macé, C. (2006). Controlling groundwater VOCs: do nanoscale
ZVI particles have any advantages over microscale ZVI or
BNP? Pollution Engineering, 38, 24–27.
Macé, C., Desrocher, S., Gheorghiu, F., Kane, A., Pupeza, M.,
Cernik, M., et al. (2006). Nanotechnology and groundwa-
ter remediation: a step forward in technology understand-
ing. Remediation, 16, 23–33.
Mackay, M. D., & Cherry, J. A. (1989). Groundwater
contamination: pump-and-treat remediation. Environmental
Science & Technology, 23, 630–636.
Masciangioli, T., & Zhang, W. X. (2003). Environmental
technologies at the nanoscale. Environmental Science &
Technology, 37, 102A–108A.
Matheson, L. J., & Tratnyek, P. G. (1994). Reductive
dehalogenation of chlorinated methanes by iron metal.
Environmental Science & Technology, 28, 2045–2053.
McMurty, D. C., & Elton, R. O. (1985). New approach to in-situ
treatment of contaminated groundwaters. Environmental
Progress, 4, 168–170.
Min, M., Xu, H., Chen, J., & Fayek, M. (2005). Evidence of
uranium biomineralization in sandstone-hosted roll-front
uranium deposits, northwestern China. Ore Geology
Reviews, 26, 198–206.
Moraci, N., & Calabrò, P. S. (2010). Heavy metals removal and
hydraulic performance in zero-valent iron/pumice permeable
reactive barriers. Journal of Environmental Management,
91, 2336–2341.
Morrison, J. S., & Sprangler, R. R. (1993). Chemical barriers
for controlling groundwater contamination. Environmental
Progress, 12, 175–181.
Morrison, J. S., Sprangler, R. R., &Morris, S. A. (1996). Subsurface
injection of dissolved ferric chloride to form a chemical
barrier: laboratory investigations. Ground Water, 34, 75–83.
Mossa, H. S., Ataie-Ashtiani, B., & Kholghi, M. (2011). Nitrate
reduction by nano-Fe/Cu particles in packed column.
Desalination. doi:10.1016/j.desal.2011.03.051.
Muftikian, R., Fernando, Q., & Korte, N. (1995). A method for the
rapid dechlorination of low molecular weight chlorinated
hydrocarbons in water. Water Research, 29, 2434–2439.
Müller, N., & Nowack, B. (2010). Nano zero valent iron—THE
solution for water and soil remediation?. Report of
workshop held in Zurich (Switzerland), November 24th
2009. http://www.observatorynano.eu/project/filesystem/
files/nZVI_final_vsObservatory.pdf. (Access 2011/04/24).
Nagpal, V., Bokare, A. D., Chikate, R. C., Rode, C. V., &
Paknikar, K. M. (2010). Reductive dechlorination of γ-
hexachlorocyclohexane using Fe–Pd bimetallic nanopar-
ticles. Journal of Hazardous Materials, 175, 680–687.
Nagpal, V., Bokare, A. D., Chikate, R. C., Rode, C. V., &
Paknikar, K. M. (2011). Reply to comment on “Reductive
dechlorination of γ-hexachlorocyclohexane using Fe–Pd
bimetallic nanoparticles”, by C. Noubactep. Journal of
Hazardous Materials. doi:10.1016/j.jhazmat.2011.04.015.
Nesic, S. (2007). Key issues related to modelling of internal
corrosion of oil and gas pipelines—a review. Corrosion
Science, 49, 4308–4338.
Noubactep, C. (2007). Processes of contaminant removal in
“Fe0–H2O” systems revisited: the importance of co-
precipitation. Open Environmental Journal, 1, 9–13.
Noubactep, C. (2008). A critical review on the mechanism of
contaminant removal in Fe0–H2O systems. Environmental
Technology, 29, 909–920.
Noubactep, C. (2010a). The suitability of metallic iron for
environmental remediation. Environmental Progress &
Sustainable Energy, 29, 286–291.
Noubactep, C. (2010b). Elemental metals for environmental
remediation: learning from cementation process. Journal
of Hazardous Materials, 181, 1170–1174.
Noubactep, C. (2010c). The fundamental mechanism of
aqueous contaminant removal by metallic iron. Water SA,
36, 663–670.
Noubactep, C. (2010d). Metallic iron for safe drinking water
worldwide. Chemical Engineering Journal, 165, 740–749.
Noubactep, C. (2011a). Metallic iron for safe drinking water
production, Freiberg Online Geology 27, 38 pp, ISSN
1434–7512. Available at: www.geo.tu-freiberg.de/fog.
Noubactep, C. (2011b). Comment on “Reductive dechlorination
of g-hexachloro-cyclohexane using Fe–Pd bimetallic
nanoparticles” by Nagpal et al [J. Hazard. Mater. 175
(2010) 680–687]. Journal of Hazardous Materials.
doi:10.1016/j.jhazmat.2011.03.081.
Noubactep, C. (2011c). Aqueous contaminant removal bymetallic
iron: is the paradigm shifting? Water SA, 37, 419–426.
Noubactep, C. (2011d). Metallic iron for water treatment: a
knowledge system challenges mainstream science. Frese-
nius Environmental Bulletin 20, 2632–2637.
Noubactep, C. (2011e). On the mechanism microbe inactivation
by metallic iron. Journal of Hazardous Materials.
doi:10.1016/j.jhazmat.2011.08.063.
Noubactep, C., & Caré, S. (2010a). Dimensioning metallic iron
beds for efficient contaminant removal. Chemical Engi-
neering Journal, 163, 454–460.
Noubactep, C., & Caré, S. (2010b). On nanoscale metallic iron
for groundwater remediation. Journal of Hazardous
Materials, 182, 923–927.
Noubactep, C., & Schöner, A. (2010). Metallic iron: dawn of a
new era of drinking water treatment research? Fresenius
Environmental Bulletin, 19, 1661–1668.
Noubactep, C., Licha, T., Scott, T. B., Fall, M., & Sauter, M.
(2009). Exploring the influence of operational parameters
on the reactivity of elemental iron materials. Journal of
Hazardous Materials, 172, 943–951.
Noubactep, C., Caré, S., Togue-Kamga, F., Schöner, A., &
Woafo, P. (2010). Extending service life of household
water filters by mixing metallic iron with sand. Clean-Soil,
Air, Water, 38, 951–959.
O’Hannesin, S. F., & Gillham, R. W. (1998). Long-term
performance of an in situ “iron wall” for remediation of
VOCs. Ground Water, 36, 164–170.
Palmer, C. D., & Wittbrodt, P. R. (1991). Processes affecting
the remediation of chromium-contaminated sites. Environ-
mental Health Perspectives, 92, 25–40.
1380 Water Air Soil Pollut (2012) 223:1363–1382
Peralta-Videa, J. R., Zhao, L., Lopez-Moreno, M. L., de la Rosa,
G., Hong, J., & Gardea-Torresdey, J. L. (2011). Nano-
materials and the environment: a review for the biennium
2008–2010. Journal of Hazardous Materials, 186, 1–15.
Pokhrel, D., & Viraraghavan, T. (2008). Arsenic removal in an
iron oxide-coated fungal biomass column: analysis of
breakthrough curves. Bioresource Technology, 99, 2067–
2071.
Pokhrel, D., & Viraraghavan, T. (2009). Biological filtration for
removal of arsenic from drinking water. Journal of
Environmental Management, 90, 1956–1961.
Pokhrel, D., Bhandari, B. S., & Viraraghavan, T. (2009).
Arsenic contamination of groundwater in the Terai region
of Nepal: an overview of health concerns and treatment
options. Environment International, 35, 157–161.
Ponder, S. M., Darab, J. G., & Mallouk, T. E. (2000).
Remediation of Cr(VI) and Pb(II) aqueous solutions using
supported, nanoscale zero-valent iron. Environmental
Science & Technology, 34, 2564–2569.
Posey-Dowty, J., Axtmann, E., Crerar, D., Borcsik, M., Ronk,
A., & Woods, W. (1987). Dissolution rate of uraninite and
uranium roll-front ores. Economic Geology, 82, 184–194.
Pradeep, T., & Anshup. (2009). Noble metal nanoparticles for
water purification: a critical review. Thin Solid Films, 517,
6441–6478.
Read, D., Lawless, T. A., Sims, R. J., & Butter, K. R. (1993).
Uranium migration through intact sandstone cores. Jour-
nal of Contaminant Hydrology, 13, 277–289.
Reardon, E. J., Fagan, R., Vogan, J. L., & Przepiora, A. (2008).
Anaerobic corrosion reaction kinetics of nanosized iron.
Environmental Science & Technology, 42, 2420–2425.
Reynolds, R. L., & Goldhaber, M. B. (1978). Origin of a
south Texas roll-type uranium deposit: I. Alteration of
iron-titanium oxide minerals. Economic Geology, 73,
1677–1689.
Romero, L., Neretnieks, I., & Moreno, L. (1992). Movement of
the redox front at the Osamu Utsumi uranium mine, Poços
de Caldas, Brazil. Journal of Geochemical Exploration,
45, 471–501.
Sakulchaicharoen, N., O’Carroll, D. M., & Herrera, J. E.
(2010). Enhanced stability and dechlorination activity of
pre-synthesis stabilized nanoscale FePd particles. Journal
of Contaminant Hydrology, 118, 117–127.
Scherer, M. M., Richter, S., Valentine, R. L., & Alvarez, P. J. J.
(2000). Chemistry and microbiology of permeable reactive
barriers for in situ groundwater clean up. Reviews in
Environmental Science & Technology, 30, 363–411.
Schrick, B., Blough, J. L., Jones, A. D., & Mallouk, T. E.
(2002). Hydrodechlorination of trichloroethylene to hydro-
carbons using bimetallic nickel–iron nanoparticles. Chem-
istry of Materials, 14, 5140–5147.
Shi, Z., Nurmi, J. T., & Tratnyek, P. G. (2011). Effects of nano
zero-valent iron on oxidation-reduction potential. Environ-
mental Science & Technology, 45, 1586–1592.
Sidborn, M., & Neretnieks, I. (2007). Long term redox
evolution in granitic rocks: modelling the redox front
propagation in the rock matrix. Applied Geochemistry, 22,
2381–2396.
Simunek, J., Jarvis, N. J., van Genuchten, M. T., & Gardenas,
A. (2003). Review and comparison of models for
describing non-equilibrium and preferential flow and
transport in the vadose zone. Journal of Hydrology, 272,
14–35.
Starr, R. C., & Cherry, J. A. (1994). In situ remediation of
contaminated ground water: the funnel-and-Gate System.
Ground Water, 32, 465–476.
Stratmann, M., & Müller, J. (1994). The mechanism of the
oxygen reduction on rust-covered metal substrates. Cor-
rosion Science, 36, 327–359.
Su, C., & Puls, R. W. (2001). Arsenate and arsenite removal by
zerovalent iron: kinetics, redox transformation, and impli-
cations for in situ groundwater remediation. Environmen-
tal Science & Technology, 35, 4562–4568.
Tarvainen, V., Ranta-Maunus, A., Hanhijärvi, A., & Forsén, H.
(2006). The effect of drying and storage conditions on
case hardening of scots pine and norway spruce timber.
Maderas Ciencia y tecnología, 8, 3–14.
Tervonen, T., Linkov, I., Figueira, J. R., Steevens, J., Chappell,
M., & Merad, M. (2009). Risk-based classification system
of nanomaterials. Journal of Nanoparticle Research, 11,
757–766.
Tong, M., Yuan, S., Long, H., Zheng, M., Wang, L., & Chen, J.
(2011). Reduction of nitrobenzene in groundwater by iron
nanoparticles immobilized in PEG/nylon membrane. Jour-
nal of Contaminant Hydrology, 122, 16–25.
Tratnyek, P. G., & Johnson, R. L. (2006). Nanotechnologies for
environmental cleanup. Nano Today, 1, 44–48.
Truex, M. J., Macbeth, T. W., Vermeul, V. R., Fritz, B. G.,
Mendoza, D. P., Mackley, R. D., et al. (2011). Demon-
stration of combined zero-valent iron and electrical
resistance heating for in situ trichloroethene remediation.
Environmental Science & Technology, 45, 5346–5351.
Truex, M. J., Vermeul, V. R., Mendoza, D. P., Fritz, B. G.,
Mackley, R. D., Oostrom, M., et al. (2011). Injection of
zero-valent iron into an unconfined aquifer using shear-
thinning fluids. Ground Water Monitoring & Remediation.,
31, 50–58.
Vodyanitskii, Y. N. (2010). The role of iron in the fixation of
heavy metals and metalloids in soils: a review of
publications. Eurasian Soil Science, 43, 519–532.
Wang, C.-B., & Zhang, W.-X. (1997). Synthesizing nanoscale
iron particles for rapid and complete dechlorination of
TCE and PCBs. Environmental Science & Technology, 31,
2154–2156.
Wu, Y., Zhang, J., Tong, Y., & Xu, X. (2009). Chromium (VI)
reduction in aqueous solutions by Fe3O4-stabilized Fe
0
nanoparticles. Journal of Hazardous Materials, 172,
1640–1645.
Xiao, S., Ma, H., Shen, M., Wang, S., Huang, Q., & Shi, X.
(2011). Excellent copper(II) removal using zero-valent
iron nanoparticle—immobilized hybrid electrospun poly-
mer nanofibrous mats. Colloids and Surfaces A: Physico-
chemical and Engineering Aspects, 381, 48–54.
Yuan, S., Zheng, Z., Meng, X.-Z., Chen, J., & Wang, L. (2010).
Surfactant mediated HCB dechlorination in contaminated
soils and sediments by micro and nanoscale Cu/Fe
Particles. Geoderma, 159, 165–173.
Zhang, W.-X. (2003). Nanoscale iron particles for environmental
remediation: an overview. Journal of Nanoparticle
Research, 5, 323–332.
Zhang, W.-X., Wang, C.-B., & Lien, H.-L. (1998).
Treatment of chlorinated organic contaminants with
Water Air Soil Pollut (2012) 223:1363–1382 1381
nanoscale bimetallic particles. Catalalysis Today, 40,
387–395.
Zhao, Y., Ren, H., Dai, H., & Jin, W. (2011). Composition
and expansion coefficient of rust based on X-ray
diffraction and thermal analysis. Corrosion Science,
53, 1646–1658.
Zhu, N., Luan, H., Yuan, S., Chen, J., Wu, X., & Wang, L.
(2010). Effective dechlorination of HCB by nanoscale Cu/
Fe particles. Journal of Hazardous Materials, 176, 1101–
1105.
Zhuang, Y., Ahn, S., Seyfferth, A. L., Masue-Slowey, Y.,
Fendorf, S., & Luthy, R. G. (2011). Dehalogenation of
polybrominated diphenyl ethers and polychlorinated bi-
phenyl by bimetallic, impregnated, and nanoscale zerova-
lent iron. Environmental Science & Technology, 45, 4896–
4903.
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