Characterizing the reactivity of metallic iron in Fe0/As-rock/H2O systems 
by long-term column experiments 
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C. Noubactep 
Angewandte Geologie, Universität Göttingen, Goldschmidtstraße 3, D - 37077 Göttingen, Germany. 
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Tel.: +49 551 39 3191, Fax.: +49 551 399379, e-mail: cnoubac@gwdg.de
Abstract 
The intrinsic reactivity of four metallic iron materials (Fe0) was investigated in batch and 
column experiments. The Fe0 reactivity was characterized by the extent of aqueous fixation of 
in-situ leached arsenic (As). Air-homogenized batch experiments were conducted for 1 month 
with 10.0 g/ℓ of an As-bearing rock (ore material) and 0.0 or 5.0 g/ℓ of Fe0. Column 
experiments were performed for 2 and 3 months. Each dynamic experiment was made up of 2 
glass columns in series. The first column contained 2.5 or 5.0 g of the ore material and the 
second column 0.0 or 5.0 g of a Fe0 material. Results showed no significant reactivity 
difference in batch studies for all 4 materials, ZVI2 was by far the most reactive material in 
column experiments. This observation was attributed to the relative kinetics of production of 
aqueous As and Fe species under the experimental conditions and their impact on the 
formation of a protective film on Fe0. Accordingly, no protective film could be built at the 
surface of the least reactive materials. The results corroborated the urgent need for unified 
experimental procedures to characterize Fe0 materials.  
Keywords: Column study, Intrinsic reactivity, Ore mineral, Water treatment, Zerovalent iron. 
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Elemental metals are efficient reactive agents for the remediation of several classes of 
environmental contaminants including arsenic, azo dyes, bacteria, halogenated organic 
compounds, heavy metals, nitrates, nitroaromatics, radionuclides, and viruses (O´Hannesin 
and Gillham, 1998; Bojic et al., 2004; Bartzas et al. 2006; Bojic et al., 2007; Henderson and 
Demond, 2007; Komnitsas et al.2007; Bojic et al., 2009; Antia, 2010; Bartzas and Komnitsas, 
2010; Bundschuh et al., 2010; Luna-Velasco et al., 2010; Noubactep, 2010a; Phillips et al., 
2010; Sarathy et al., 2010, Comba et al., 2011; Giles et al., 2011; ITRC, 2011; Lin et al., 
2011; Noubactep, 2011a; Salter-Blanc et al., 2011). Metallic iron (Fe0) is currently the most 
used material for field applications (Gillham, 2010; Comba et al., 2011; Gheju, 2011; 
Henderson and Demond, 2011; ITRC, 2011; Salter-Blanc et al., 2011). 
Despite the wide variety of environmental contaminants and their possible specific 
interactions with Fe0, tested materials were characterized mainly by their surface area, size 
and interface chemistry (e.g. surface state). However, it has been traceably demonstrated that 
none of these structural and physical characteristics is really determinant for the chemical 
reactivity of Fe0 (Reardon, 1995; Landis et al., 2001: Noubactep et al., 2005; Reardon, 2005; 
Noubactep et al., 2009). For instance, Landis et al. (2001) reported that Fe0 materials of 
comparable particle size (comparable surface area) exhibited reactivity differences greater 
than a three-fold for cDCE and VC degradation rates in column studies. This example 
substantiates that a broad understanding of the chemical reactivity is urgently needed. 
Several sources of Fe0 materials have been reported in literature to be efficient for aqueous 
contaminant removal (Landis et al., 2001; Miehr et al., 2004; Leupin and Hug, 2005; 
Noubactep et al., 2005; Gheju and Iovi, 2006; Satapanajaru et al., 2006; Yang et al., 2006; 
Ngai et al., 2007; Gheju et al., 2008; Gheju and Balcu, 2010; Gheju and Balcu, 2011, Wanner 
et al. 2011). These include commercial Fe0 for contaminant removal (e.g. Connelly iron, 
Peerless iron, iron from G. Maier GmbH), commercial iron for other purposes (e.g. 
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construction steel, iron nails, steel wool), scrap iron, production by-products, Fe0 prepared in-
situ by reduction of iron salts. Although, many tested materials have been reported highly 
reactive and recommendable for field application, efficacy of a Fe
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0 in terms of high removal 
capacity for a specific contaminant in short term experiments is not a guarantee for high 
removal capacity in field applications. Moreover, researchers working with nano-scale Fe0 
usually compare their results to that of conventional micro-scale Fe0 (Noubactep et al., 2012 
and ref. cited therein). The question is what is the reference material to which innovative 
materials should be compared?  
The present study is a continuation of a series of works aiming at introducing reliable tools for 
the evaluation of the intrinsic reactivity of various Fe0 materials. A method based on the 
characterization of Fe dissolution in a 2 mM EDTA solution was first proposed (Noubactep et 
al., 2004; 2005; 2009). The method was proven less efficient for powdered materials and for 
materials with high proportion of fines (Noubactep, 2010b). On the other hand, Fe0 
dissolution is not necessarily coupled to contaminant removal. These limitations have led to 
the development of a second experimental tool in which Fe0 is characterized by the extent of 
the discoloration of methylene blue (MB) in the presence of manganese dioxide (MnO2) (MB-
test) (Noubactep, 2009). The MB-test was shown more efficient and more affordable than the 
EDTA-test but was limited by the lack of reference MnO2 materials. Both tests could enable 
an advanced material screening. However, from the tested 18 materials seven were still 
exhibiting very similar reactivity. Therefore, new approaches are needed. 
A further possibility to characterize the reactivity of Fe0 materials is to stress them in systems 
where building of a protective film at their surface is likely to occur. Such a system was 
identified recently while characterizing the solubilization of toxic species from natural rocks 
(Noubactep et al., 2008a; 2008b). It was shown that elevated amounts of As could be leached 
from an ore material for a long time (up to 99 days). Accordingly introducing the same 
amount of various Fe0 materials in system capable of producing As concentration as large as 
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1000 mg/ℓ could be a powerful tool to investigate the impact of As on the formation of the 
oxide film (mixed oxides) on the process of contaminant removal by Fe
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0. As a rule, the more 
reactive a material, the more rapid the passivation process (protective film formation). In 
other words, the system with the most reactive material will exhibit the least contaminant 
removal efficiency. 
The objective of this study is to present a new contribution to the effort for the development 
of reliable protocols for the comparison of the intrinsic reactivity of different Fe0 materials. 
For this purpose, four selected materials from former works are tested. One of the materials 
was essentially less reactive than the others. The 3 other materials were very closed in their 
reactivity by both tests described above (Noubactep, 2010b). The results confirmed the 
suitability of the used method and opened new routes for coupling the investigation of 
contaminant release and contaminant removal under relevant conditions. 
Materials and methods 
Solid materials 
As-bearing rock 
The used As ore material originates from the Otto-Stollen in Breitenbrunn/Erzgebirge 
(Saxony, Germany). The material was selected on the basis of its high arsenic content (80 %). 
A qualitative SEM analysis shows the presence of As, Ca, F, Fe, O, S and Si (Noubactep et 
al., 2008). The ore material is primarily a hydrothermal vein material and arsenic occurred as 
native arsenic (As0) and Loellingite (FeS2 - As–I) (Jones and Nesbitt, 2002) in paragenesis 
with hydrothermal vein carbonates (for example Fe-bearing calcite or dolomite). The mineral 
was ground and sieved. The particle size fraction 0.063 ≤ d (mm) ≤ 0.10 was used without 
any pre-treatment. 
Fe0 materials 
One scrap iron (ZVI1), and three commercially available Fe0 materials have been tested. The 
main characteristics of these materials are summarized elsewhere (Noubactep, 2010b). ZVI2 
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is a spherical material (d = 1.2 mm) from Würth (Germany) termed as ‘Hartgußgranulat’. 
ZVI3 are iron chips from G. Maier GmbH Rheinfelden (Germany) termed as 
‘Graugußgranulat’. ZVI4 is a direct reduced iron from ISPAT GmbH (Germany), termed 
‘Schwammeisen’.  Before used ZVI4 was crushed and sieved; the size fraction 1.0-2.0 mm 
was used. The specific surface area of the materials varies between 0.043 and 0.63 m
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2 g-1. 
These data were compiled from the literature (Tab. 1). The materials were compared solely on 
the basis of the extent of As removal by the same initial mass of Fe0 (e.g. 5.0 g in columns) 
under similar experimental conditions. The materials differ regarding their characteristics 
such as iron content, nature and proportion of alloying elements, and shape. 
The four used materials were selected from nine materials which were recently characterized 
by leaching with 2 mM EDTA in column study (Noubactep, 2010b). In turn, the tested nine 
materials were selected from eighteen materials after a screening in batch experiments using 
the EDTA-test (Noubactep et al., 2005; Noubactep et al., 2009). Both tests could not really 
differentiate the reactivity of ZVI1, ZVI3 and ZVI4. The reactivity of these three materials 
toward As removal from a natural rock was investigated in this study. For comparison the 
least reactive commercial Fe0 (ZVI2) was incorporated in this study. 
Sand  
The used sand was a commercial material for aviculture (“Papagaiensand” from RUT – 
Lehrte/Germany). Papagaiensand was used as received without any further pre-treatment nor 
characterization. This sand was the operational reference non-adsorbing material. 
Leaching solution 
The leaching solution was tap water of the city of Göttingen (Lower Saxonia, Germany). Tap 
water was discussed as a better proxy for natural groundwater than synthetic solutions 
(Noubactep, 2003; Noubactep et al., 2008). The rationale behind this assumption is that, in 
many cases, natural water is just treated for iron and manganese removal. The average 
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composition (in mg/ℓ) of the used tap water was: Cl–: 7.7; NO3–: 10.0; SO42-: 37.5; HCO3-: 
88.5; Na
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+: 7.0; K+: 1.2; Mg2+: 7.5; Ca2+: 36.1; and an initial pH 8.3. 
As leaching and immobilization 
Air homogenized batch experiments 
These experiments were conducted in special reaction vessels allowing the system to be 
homogenized by a humid current of air supplied by a small aquarist pump. The goal was to 
homogenize the experimental systems at atmospheric pressure (PCO2 = 0.035 %) without 
breaking down the materials. 10.0 g/ℓ of the ore material and 0.0 or 5.0 g/ℓ of Fe0 were 
allowed to react in sealed vessels containing 100 mℓ of tap water at laboratory temperature 
(22 ± 3 °C) for up to 30 days. At given dates, 1.5 mℓ of the solution was retrieved and diluted 
for As analysis and the same volume of tap water was added to the system. 
Column experiments 
The tap water was pumped upwards from PE bottles using a peristaltic pump (Ismatec, ICP 
24). Tygon tubes were used to connect inlet reservoir, pump, column and outlet. Ten glass 
columns (40 cm long, 2.6 cm inner diameter) were used in two series of experiments. The 
columns were mostly packed with sand. The effective length, the bulk density and the 
porosity of the packed columns were not characterized as they were not necessary for the 
discussion of the results. The extent of As dissolution by water and the extent of its removal 
by selected Fe0 materials were the sole targets. The experiments were performed at room 
temperature (22 ± 3 °C). A stable flow rate was maintained throughout the experiment.  
Five parallel experiments were performed in each series. The same mass of the rock (2.5 or 
5.0 g) was placed in a first column and 5.0 g of each tested Fe0 was placed in the second 
column (Fig. 1). In the reference system, the second column contained only sand (no Fe0). 
The experiments were stopped after 65 or 97 days when each column was leached by 19 or 
25.0 ℓ of tap water (Tab. 2). The water flow rate was constant at 12.0 mℓ/h. 
Analytical methods 
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Analysis for As was performed by inductively coupled plasma mass spectrometry (ICP-MS) 
at the Department of Geochemistry (Centre of Geosciences, University of Göttingen). All 
chemicals used for the experiments and analysis were of analytical grade. The pH value was 
measured by combining glass electrodes (WTW Co., Germany). The electrodes were 
calibrated with five standards following a multi-point calibration protocol and in agreement 
with the new IUPAC recommendation (Meinrath and Spitzer, 2000; Buck et al., 2002). 
Expression of experimental results 
The mass (m) of leached As (mg) at any time (t) is calculated from the concentration of the 
effluent using Eq. (1): 
m = P.V        (1) 
Where P is the As concentration (in mg/ℓ) and V the volume (ℓ). At the end of the experiment 
the total amount of leached As can be calculated by addition and the extend of As leaching by 
tap water deduced. Knowing the As percentage in the natural rock (80 %), the maximal 
leachable mass (m0) of As can be calculated. The percentage (P) As leaching at each time is 
given by Eq. (2): 
P = 100 * m/m0      (2) 
At each time the amount of As leached in the reference system (Pref) can be set to 100 and the 
relative leaching percent (Prel) for all other systems deduced by Eq. (3): 
P* = Prel = 100 * P/Pref     (3) 
Finally, the relative percent of As removal (Pfix) by each material is given by Eq. (4): 
Pfix = 100 - Prel      (4) 
Results and discussion 
The particularity of As-rock/Fe0 systems investigated here is that aqueous As and solid Fe 
hydroxides and oxides for their removal are generated in-situ. It has been traceably 
demonstrated, that AsIII and AsV are removed in Fe0/H2O systems by adsorption and co-
precipitation (Lackovic et al., 2000; Farrell et al., 2001; Noubactep, 2010a; Noubactep, 
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2011b; Noubactep, 2011c; Noubactep, 2012). As released from the used ore material was 
recently characterized (Noubactep et al., 2008a) and the process of As dissolution will not be 
discussed here. The basis for the characterization of Fe
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0 materials is that the smallest As 
concentration (relative to the reference system) is encountered in the system with the most 
reactive material under testing conditions. 
Batch experiments 
Fig. 2 illustrates As dissolution in the absence (reference) or the presence of tested Fe0 
materials (ZVI1 through ZVI4) as a function of time. Both leaching kinetics and the extent of 
As release substantially decreased with the addition of Fe0. From Fig. 2 not visual reactivity 
difference could be performed. It appears that the reactivity of all four materials is very closed 
to each other. A look on Pfix-values (Tab. 3) shows that the relative fixation efficiency varies 
from 30.4 to 37.3 %. A tentative order of increasing reactivity based on these values is: ZVI3 
< ZVI1 < ZVI2 < ZVI4. Remember that the order of reactivity after the EDTA-test 
(Noubactep et al., 2005; 2009) and the MB-test (Noubactep, 2009) were univocally: ZVI2 < 
ZVI1 ≅ ZVI3 ≅ ZVI4. Accordingly, air-homogenized batch experiments are not appropriate 
for the differentiation of the reactivity of ZVI1, ZVI3 and ZVI4. It is well-known that batch 
systems can not give an image of processes occurring in nature (Wang et al. 2009). The first 
reason in regard to the experimental conditions of this work is the possibility of super-
saturation of the As solution given the too long contact time (30 days) and the relative strong 
homogenisation with air-bubbles. To account for this further characterizations were 
performed under dynamic conditions. 
Column experiments for 65 days 
Fig. 3a shows the effect of tested ZVIs on As leaching from the natural ore as a function of 
time (≤ 65 days). 5.0 g of the ore material was placed in the first column and 5.0 g of ZVI in 
the second column (Fig. 1). From Fig. 3a a visual differentiation of ZVI2 is evident. It is also 
evident that the reference system exhibited the highest As concentration. The cumulative sum 
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of released As (Fig. 3b) confirmed this trend. The m- and Pfix-values from Tab. 3 confirmed 
these observations. The deduced increasing order of increasing reactivity is: ZVI3 ≅ ZVI4 < 
ZVI1 < ZVI2. This classification showing that ZVI2 was the most reactive materials is 
acceptable but the experimental conditions should be further modified to obtain a clear trend. 
The following modifications were operated: (i) the mass of ore material was halved, (ii) the 
first two litres of leaching solution in Fe
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0/As-rock systems were discarded, and (iii) the 
experimental duration was lengthened to 97 days. All other parameters (flow rate, 
temperature) were kept constant. The rationale behind discarding the first two litres was the 
elevated As concentration in this initial phase (Fig. 3a). 
The next important feature from Fig 3a is that 5.0 g of the used ore material is capable of 
producing about 17 mg/ℓ (reference system) As for more that two months. High As 
concentration was intentionally tested here. By varying the mass of the ore material and the 
particle size, As concentrations relevant for each specific size could be achieved. 
Column experiments for 97 days 
Fig. 4a shows the effect of tested ZVIs on As leaching from the natural ore as a function of 
time (≤ 97 days). A visual reactivity difference can be better performed than in Fig. 3a. The 
visual increasing order of reactivity is: ZVI4 < ZVI3 < ZVI1 < ZVI2. The m- and Pfix-values 
from Tab. 3 confirmed this trend with the additive information, that ZVI3 and ZVI4 are very 
closed in their reactivity as the percent As removal (Pfix-values) was 11.9 and 11.7 
respectively. Fig. 4b clearly confirmed the results from Tab. 3. 
From Fig. 4a is clear that 2.5 g of the As-rock is able to produce about 16 mg/ℓ As for more 
than three months. These results shows clearly that long-term experiments regarding As 
removal can be coupled with As leaching from natural ores. By reducing the ore mass, 
changing the particle size and using different ores, it is possible to perform long-term leaching 
experiments in the laboratory. Such experiments could help to bridge the huge gap between 
the laboratory and the field (Wang et al., 2009). On the other hand, parameters from such 
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systems could help to develop more reliable models to predict contaminant leaching in the 
environment. 
Discussion 
The use of Fe0 materials for environmental remediation is severely handicapped by the lack of 
methods for characterization of the chemical reactivity. The current procedure of testing the 
reactivity of Fe0 for individual contaminants (Landis et al., 2001; Miehr et al., 2004; Leupin 
and Hug, 2005; Gheju and Iovi, 2006, Wanner et al. 2011) is not very useful as no 
comparison between two independent works is possible, even for the same contaminant. 
Ideally, there should be a universally acceptable/accepted method to evaluate various Fe0 for 
their chemical reactivity. Accordingly, it is contemplated to propose protocols, which could 
be used to compare the efficiency of different Fe0. 
From available works, only the characterization of Fe0 by the extent of H2 production 
(Reardon, 1995; 2005) could be regarded as universally applicable method to characterise Fe0 
intrinsic reactivity. However, this protocol is not necessarily affordable and used relative high 
Fe0 masses (15.0 to 400 g). Accordingly, most simple and affordable tests should be 
developed. The EDTA-test (Noubactep et al., 2005; 2009) and the MB-test (Noubactep, 2009) 
are simple and affordable but they could not address the passivation of tested material. 
In using elevated As concentrations, the present work has corroborated warnings to perform 
contaminant removal experiments with over-saturated solutions (e.g. Kalin et al., 2005). 
However, more than the instability of used solutions introducing biases in the extent of 
contaminant removal by the tested process, this study has delineated the impact of elevated 
concentrations on the passivation process. In fact, in nature, contaminants are rarely available 
at high concentration (Henderson and Demond, 2011; Kümmerer, 2011) and contaminated 
water enters the zone containing Fe0 when an oxide scale is already formed at its surface. In 
other words, while using elevated contaminant concentrations, an artificial system is created 
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that could not be reproduced in nature. On the other hand, elevated contaminant 
concentrations necessarily impact the process of film formation on Fe
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0 (Noubactep, 2010c).  
Concluding remarks 
In an attempt to access their intrinsic chemical reactivity, the performance of 3 commercial 
Fe0 (ZVI2, ZVI3 and ZVI4) and one scrap iron (ZVI1) for the removal of As has been 
evaluated in long-term column studies. As was leached from a natural rock using the tap 
water of the city of Göttingen as leaching solution. The results confirmed findings from 
previous works that ZVI2 is the least reactive material (Noubactep et al., 2005; 2009; 
Noubactep, 2010a, Noubactep 2011d). It could be further shown that ZVI1 is less reactive 
than ZVI3 and ZVI4. 
The test methodology consisting in leaching As with tap water can be further improved or 
adapted to investigate several aspects of contaminant release and contaminant removal. For 
example, by reducing the mass of the ore material, As concentration relevant to field 
situations could be obtained and used to characterize the performance of Fe0 materials for As 
removal. On the other hand, using several leaching solutions could enable the characterization 
of the impact of relevant ions on the process of As leaching (and/or removal). Such 
experiments could be designed on the basis of site-specific situations. It is hoped that this new 
experimental tool will accelerate efforts to characterize the intrinsic reactivity of Fe0 
materials. 
Acknowledgments 
The used arsenic-bearing ore was purchased by the Department of Geology of the Technical 
University Bergakademie Freiberg. Dr. Klaus Simon (Centre of Geosciences, University of 
Göttingen) is acknowledged for the As analysis. Dr. Ralf. Köber (Institute Earth Science of 
the University of Kiel) kindly purchased the commercial ZVI samples. The scrap iron (Sorte 
69) was kindly purchased by the branch of the MAZ (Metallaufbereitung Zwickau, Co) in 
Freiberg. Mohammad Azizur Rahman (Angewandte Geologie, Universität Göttingen) is 
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acknowledged for technical support. The work was supported by the Deutsche 
Forschungsgemeinschaft (DFG-No 626/2-2). 
References 
ANTIA DDJ (2010) Sustainable zero-valent metal (ZVM) water treatment associated with 
diffusion, infiltration, abstraction and recirculation. Sustainability 2 2988–3073. 
BARTZAS G, and KOMNITSAS K (2010) Solid phase studies and geochemical modelling of 
low-cost permeable reactive barriers. J. Hazard. Mater. 183 301–308. 
BARTZAS G, KOMNITSAS K, and PASPALIARIS I (2006) Laboratory evaluation of Fe0 
barriers to treat acidic leachates. Miner. Eng. 19 505–514. 
BOJIC A, PURENOVIC M, and BOJIC D (2004) Removal of chromium(VI) from water by 
micro-alloyed aluminium based composite in flow conditions. Water SA 30 353–359. 
BOJIC A, PURENOVIC M, BOJIC D, and NDJELKOVIC T (2007) Dehalogenation of 
trihalomethanes by a micro-alloyed aluminium composite under flow conditions. Water 
SA 33 297–304. 
BOJIC A, BOJIC D, and ANDJELKOVIC T (2009) Removal of Cu2+ and Zn2+ from model 
wastewaters by spontaneous reduction–coagulation process in flow conditions. J. 
Hazard. Mater. 168 813–819. 
BUCK RP, RONDININI S, COVINGTON AK, BAUCKE FGK, BRETT CMA, CAMOES 
MF, MILTON MJT, MUSSINI T, NAUMANN R, PRATT KW, SPITZER P, and 
WILSON GS (2002) Measurement of pH. Definition, standards, and procedures 
(IUPAC Recommendations 2002). Pure Appl. Chem. 74 2169–2200. 
BUNDSCHUH J, LITTER M, CIMINELLI VST, MORGADA ME, CORNEJO L, HOYOS 
SG, HOINKIS J, ALARCÓN-HERRERA MT, ARMIENTA MA, and 
BHATTACHARYA P (2010) Emerging mitigation needs and sustainable options for 
solving the arsenic problems of rural and isolated urban areas in Latin America – A 
critical analysis. Water Res. 44 5828–5845. 
 12
305 
306 
307 
308 
309 
310 
311 
312 
313 
314 
315 
316 
317 
318 
319 
320 
321 
322 
323 
324 
325 
326 
327 
328 
329 
COMBA S, DI MOLFETTA A, and SETHI R (2011) A Comparison between field 
applications of nano-, micro-, and millimetric zero-valent iron for the remediation of 
contaminated aquifers. Water Air Soil Pollut. 215 595–607. 
FARRELL J, WANG J, O'DAY P, and CONKLIN M (2001) Electrochemical and 
spectroscopic study of arsenate removal from water using zero-valent iron media. 
Environ. Sci. Technol. 35 2026–2032.
GHEJU M (2011) Hexavalent Chromium Reduction with Zero-Valent Iron (ZVI) in Aquatic 
Systems. Water Air Soil Pollut. 222 103–148. 
GHEJU M, and IOVI A (2006) Kinetics of hexavalent chromium reduction by scrap iron. J. 
Hazard. Mater. 135 66–73. 
GHEJU M, IOVI A, and BALCU I (2008) Hexavalent chromium reduction with scrap iron in 
continuous-flow system: Part 1: Effect of feed solution pH. J. Hazard. Mater. 153 655–
662. 
GHEJU M, and BALCU I (2010) Hexavalent chromium reduction with scrap iron in 
continuous-flow system. Part 2: Effect of scrap iron shape and size. J. Hazard. Mater. 
182 484–493. 
GHEJU M, and BALCU I (2011) Removal of chromium from Cr(VI) polluted wastewaters by 
reduction with scrap iron and subsequent precipitation of resulted cations. J. Hazard. 
Mater. 196 131–138. 
GILES DE, MOHAPATRA M, ISSA TB, ANAND S, and SINGH P (2011): Iron and 
aluminium based adsorption strategies for removing arsenic from water. J. Environ. 
Manage. 92 3011–3022. 
GILLHAM RW (2010) Development of the granular iron permeable reactive barrier 
technology (good science or good fortune). In "Advances in environmental geotechnics : 
proceedings of the International Symposium on Geoenvironmental Engineering in 
 13
330 
331 
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333 
334 
335 
336 
337 
338 
339 
340 
341 
342 
343 
344 
345 
346 
347 
348 
349 
350 
351 
352 
353 
354 
355 
Hangzhou, China, September 8-10, 2009"; Y. Chen, X. Tang, L. Zhan (Eds);  Springer 
Berlin/London, pp. 5–15. 
HENDERSON AD, and DEMOND AH (2007) Long-term performance of zero-valent iron 
permeable reactive barriers: a critical review. Environ. Eng. Sci. 24 401–423. 
HENDERSON AD, and DEMOND AH (2011) Impact of solids formation and gas production 
on the permeability of ZVI PRBs. J. Environ. Eng. 137 689–696. 
ITRC (Interstate Technology & Regulatory Council) (2011) Permeable reactive barrier: 
Technology update. PRB-5. Washington, D.C.: Interstate Technology & Regulatory 
Council, PRB: Technology Update Team. www.itrcweb.org (access: 09.03.2012). 
JONES RA, and NESBITT HW (2002) XPS evidence for Fe and As oxidation states and 
electronic states in loellingite (FeAs2). Amer. Miner. 87 1692–1698. 
KALIN M, WHEELER WN, and MEINRATH G (2005) The removal of uranium from 
mining waste water using algal/microbial biomass. J. Environ. Radioact. 78 151–177. 
KOMNITSAS K, BARTZAS G, FYTAS K, and PASPALIARIS I (2007) Long-term 
efficiency and kinetic evaluation of ZVI barriers during clean-up of copper containing 
solutions. Miner. Eng. 20 1200–1209. 
KÜMMERER K (2011) Emerging contaminants versus micro-pollutants. Clean – Soil, Air, 
Water 39 889–890. 
LACKOVIC JA, NIKOLAIDIS NP, and DOBBS GM (2000) Inorganic arsenic removal by 
zero-valent iron. Environ. Eng. Sci. 17 29–39. 
LANDIS RL, GILLHAM RW, REARDON EJ, FAGAN R, FOCHT RM, and VOGAN JL 
(2001) An examination of zero-valent iron sources used in permeable reactive barriers. 
3rd International Containment Technology Conference (10-13 June 2001), Florida State 
University, Tallahassee. Orlando, FL. 5 pages. 
LEUPIN OX, and HUG SJ (2005) Oxidation and removal of arsenic (III) from aerated 
groundwater by filtration through sand and zero-valent iron. Water Res. 39 1729–1740. 
 14
356 
357 
358 
359 
360 
361 
362 
363 
364 
365 
366 
367 
368 
369 
370 
371 
372 
373 
374 
375 
376 
377 
378 
379 
380 
381 
LIN H, ZHU L, XU X, ZANG L, and KONG Y (2011) Reductive transformation and 
dechlorination of chloronitrobenzenes in UASB reactor enhanced with zero-valent iron 
addition. J. Chem. Technol. Biotechnol. 86 290–298. 
LUNA-VELASCO A, SIERRA-ALVAREZ R, CASTRO B, and FIELD JA (2010) Removal 
of nitrate and hexavalent uranium from groundwater by sequential treatment in 
bioreactors packed with elemental sulfur and zero-valent iron. Biotechnol. Bioeng. 107 
933–942. 
MEINRATH G, and SPITZER P (2000) Uncertainties in determination of pH. Mikrochem. 
Acta 135 155–168. 
MIEHR R, TRATNYEK GP, BANDSTRA ZJ, SCHERER MM, ALOWITZ JM, and 
BYLASKA JE (2004) Diversity of contaminant reduction reactions by zerovalent iron: 
Role of the reductate. Environ. Sci. Technol. 38 139–147.
NGAI TKK, SHRESTHA RR, DANGOL B, MAHARJAN M, and MURCOTT SE (2007) 
Design for sustainable development – Household drinking water filter for arsenic and 
pathogen treatment in Nepal. J. Environ. Sci. Health A 42 1879–1888. 
NOUBACTEP C (2003) Investigations for the passive in-situ Immobilization of Uranium 
(VI) from Water (in German). Dissertation, TU Bergakademie Freiberg, Wiss. Mitt. 
Institut für Geologie der TU Bergakademie Freiberg, Band 21, 140 pp, ISSN1433-1284. 
NOUBACTEP C (2009) Characterizing the reactivity of metallic iron upon methylene blue 
discoloration in Fe0/MnO2/H2O systems. J. Hazard. Mater. 168 1613–1616. 
NOUBACTEP C (2010a) The fundamental mechanism of aqueous contaminant removal by 
metallic iron. Water SA 36 663–670. 
NOUBACTEP C (2010b) Characterizing the reactivity of metallic iron in Fe0/EDTA/H2O 
systems with column experiments. Chem. Eng. J. 162 656–661. 
NOUBACTEP C (2010c) Elemental metals for environmental remediation: Learning from 
cementation process. J. Hazard. Mater. 181 1170–1174. 
 15
NOUBACTEP C (2011a) Metallic Iron for Safe Drinking Water Production. Freiberg Online 
Geology 27, 38 pp, ISSN 1434-7512. (
382 
www.geo.tu-freiberg.de/fog) 383 
384 
385 
386 
387 
388 
389 
390 
391 
392 
393 
394 
395 
396 
397 
398 
399 
400 
401 
402 
403 
404 
405 
406 
NOUBACTEP C (2011b) Aqueous contaminant removal by metallic iron: Is the paradigm 
shifting? Water SA 37 419–426. 
NOUBACTEP C (2011c) Metallic iron for water treatment: A knowledge system challenges 
mainstream science. Fresenius Environ. Bull. 20 2632–2637. 
NOUBACTEP C (2011d) Characterizing the reactivity of metallic iron in Fe0/UVI/H2O 
systems by long-term column experiments. Chem. Eng. J. 171 393–399. 
NOUBACTEP C (2012) Investigating the processes of contaminant removal in Fe0/H2O 
systems. Korean J. Chem. Eng., doi: 10.1007/s11814-011-0298-8. 
NOUBACTEP C, CHEN-BRAUCHER D, and SCHLOTHAUER T (2008) Arsenic release 
from a natural rock under near-natural oxidizing conditions. Eng. Life Sci. 8 622–630. 
NOUBACTEP C, FALL M, MEINRATH G, and MERKEL B (2004) A simple method to 
select zero valent iron material for groundwater remediation. paper presented at the 
Quebec 2004, 57TH Canadian Geotechnical Conference, 5TH Joint CGS/IAH-CNC 
Conference, Session 1A, pp. 6-13. 
NOUBACTEP C, LICHA T, SCOTT TB, FALL M, and SAUTER M (2009) Exploring the 
influence of operational parameters on the reactivity of elemental iron materials. J. 
Hazard. Mater. 172 943–951. 
NOUBACTEP C, MEINRATH G, DIETRICH P, SAUTER M, and MERKEL B (2005) 
Testing the suitability of zerovalent iron materials for reactive walls. Environ. Chem. 2 
71–76. 
NOUBACTEP C, SCHÖNER A, and SCHUBERT M (2008) Characterizing As, Cu, Fe and 
U solubilization by natural waters. In Merkel B.J., Hasche-Berger A. (Eds.) Uranium in 
the Environment. Springer, Berlin, Heidelberg; 549–558. 
 16
407 
408 
409 
410 
411 
412 
413 
414 
415 
416 
417 
418 
419 
420 
421 
422 
423 
424 
425 
426 
427 
428 
429 
430 
431 
NOUBACTEP C, CARÉ S, and CRANE RA (2012) Nanoscale metallic iron for 
environmental remediation: prospects and limitations. Water Air Soil Pollut. 223 1363–
1382. 
O´HANNESIN SF, and GILLHAM RW (1998) Long-term performance of an in situ "iron 
wall" for remediation of VOCs. Ground Water 36 164–170. 
PHILLIPS DH, VAN NOOTEN T, BASTIAENS L, RUSSELL MI, DICKSON K, PLANT S, 
AHAD JME, NEWTON T, ELLIOT T, and KALIN RM (2010) Ten year performance 
evaluation of a field-scale zero-valent iron permeable reactive barrier installed to 
remediate trichloroethene contaminated groundwater. Environ. Sci. Technol. 44 3861–
3869. 
REARDON EJ (2005) Zerovalent irons: Styles of corrosion and inorganic control on 
hydrogen pressure buildup. Environ. Sci. Tchnol. 39 7311–7317.
REARDON EJ (1995): Anaerobic corrosion of granular iron: Measurement and interpretation 
of hydrogen evolution rates. Environ. Sci. Technol. 29 2936–2945. 
SALTER-BLANC AJ, and TRATNYEK PG (2011) Effects of solution chemistry on the 
dechlorination of 1,2,3-trichloropropane by zero-valent zinc. Environ. Sci. Technol. 45 
4073–4079. 
SARATHY V, SALTER AJ, NURMI JT, JOHNSON GO, JOHNSON RL, and TRATNYEK 
PG (2010) Degradation of 1,2,3-trichloropropane (TCP): Hydrolysis, elimination, and 
reduction by iron and zinc. Environ. Sci. Technol. 44 787–793. 
SATAPANAJARU T, ANURAKPONGSATORN P, SONGSASEN A, BOPARAI H, and 
PARK J (2006) Using low-cost iron byproducts from automotive manufacturing to 
remediate DDT. Water Air Soil Pollut. 175 361–374. 
WANG T-H, LI M-H, TENG S-P (2009) Bridging the gap between batch and column 
experiments: A case study of Cs adsorption on granite. J. Hazard. Mater. 161 409–415. 
 17
432 
433 
434 
435 
436 
437 
438 
WANNER C, EGGENBERGER U, MÄDER U (2011) Reactive transport modelling of 
Cr(VI) treatment by cast iron under fast flow conditions. Appl. Geochem. 26 1513–
1523. 
YANG JE, KIM JS, OK YS, KIM S-J, YOO K-Y (2006) Capacity of Cr(VI) reduction in an 
aqueous solution using different sources of zerovalent irons. Korean J. Chem. Eng. 23 
935–939. 
 18
Table 1. Elemental composition and specific surface area (SSA) of iron materials used in this 
study. n.d. = not determined. Modified from Noubactep (2010b). 
438 
439 
440  
ZVI Element (%) SSA 
 C Si Mn Cr Mo Ni Fe (m2/g) 
ZVI4 1.96 0.12 0.09 0.003 n.d. <0.001 86.3 0.63 
ZVI2 3.39 0.41 1.10 0.34 n.d. 0.088 91.5 0.043 
ZVI1 3.52 2.12 0.93 0.66 n.d. n.d. 99.8 0.29 
ZVI3 3.13 2.17 0.36 0.077 n.d. 0.056 96.7 0.50 
441 
442 
 
 19
Table 2: Summary of the experimental conditions for As release from the As-mineral and As 
removal by Fe
442 
443 
444 
445 
446 
0 in batch and column studies for the four tested Fe0 materials. ‘VT’ 
is the total volume of tap water that has flowed into the individual columns. 
General conditions: pH0 = 8.3 and T = 23 ± 2 °C. 
 
  Batch column 1 column 2 
duration (d) 30.0 65.0 97.0 
As-mineral (g) 1.0 5.0 2.5 
ZVI (g) 0.5 5.0 5.0 
VT (ℓ) 0.1 19.0 25.0 
flow rate (mℓ/h) air bubbled 11.9 11.9 
447 
448 
 
 20
Table 3: Extent of As release (in mg) from the natural mineral as influenced by the presence 
of Fe
448 
449 
450 
451 
452 
453 
0 in batch and column studies. The system without Fe0 is used as reference to 
characterize the extent of As removal by individual ZVIs (Pfix in %). As a rule, the 
more reactive a material the bigger the m and Pfix values. General conditions: pH0 
= 8.3 and T = 23 ± 2 °C. 
 
 Batch Column 1 Column 2 
 m Pfix m Pfix m Pfix
 (mg) (%) (mg) (%) (mg) (%) 
reference 173 0.0 338 0.0 3745 0.0 
ZVI1 118 32.2 307 9.1 2985 20.3 
ZVI2 117 32.6 243 28.0 2782 25.7 
ZVI3 121 30.4 322 4.6 3301 11.9 
ZVI4 109 37.3 322 4.6 3306 11.7 
454 
455 
 
 21
Figure 1 455 
456  
457 
 22
Figure 2 458 
459 
460 
 
 
0 6 12 18 24 30
0
300
600
900
1200
1500
1800
 Reference
 ZVI1
 ZVI2
 ZVI3
 ZVI4
A
s 
/ [
m
g/
L]
elapsed time / [days]
 461 
462 
 23
Figure 3 
0 10 20 30 40 50 60 70
3000
6000
9000
12000
15000
18000
21000
(a)
mAs-rock = 5.0 g
mZVI = 5.0 g
0.63 < drock < 1.0
 reference
 ZVI1
 ZVI2
 ZVI3
 ZVI4
ar
se
ni
c 
/ [
μg
/L
]
elapsed time / [days]
 
462 
463  
464 
465  
0 10 20 30 40 50 60 70
0
50
100
150
200
250
300
350
(b)
mAs-rock = 5.0 g
mZVI = 5.0 g
0.63 < drock < 1.0
 reference
 ZVI1
 ZVI2
 ZVI3
 ZVI4
Σa
rs
en
ic
 / 
[m
g]
elapsed time / [days]
 466 
467 
468 
469 
470 
471 
 
 
 
 
 24
Figure 4 
0 20 40 60 80 100
7500
9000
10500
12000
13500
15000
16500
(a)
 Reference
 ZVI1
 ZVI2
 ZVI3
 ZVI4
ar
se
ni
c 
/ [
μg
/L
]
elapsed time / [days]
 
471 
472 
473 
 
 
474 
475 
476 
 
 
0 15 30 45 60 75 90
0
75
150
225
300
375
450
(b)
mAs-rock = 2.5 g
mZVI = 5.0 g
0.63 < drock < 1.0
 Reference
 ZVI1
 ZVI2
 ZVI3
 ZVI4
Σa
rs
en
ic
 / 
[m
g]
elapsed time / [days]
 477 
478 
 25
Figure Captions 478 
479 
480 
481 
482 
483 
484 
485 
486 
487 
488 
489 
490 
491 
492 
493 
494 
495 
 
Figure 1: 
Schematic diagram of the experimental design for the two-columns studies. Used materials 
were As mineral (2.5 or 5.0 g) and Fe0 (5.0 g). 
 
Figure 2: Arsenic release [mg/ℓ] from the base material as a function of time in air-
homogenized batch experiments. The lines are not fitting functions, they simply connect 
points to facilitate visualization. 
Figure 3: 
Extent of arsenic release from the two-column-systems for 65 days by tested Fe0 materials: (a) 
variation of the extent of As release with time, and (b) cumulative As release (mg). The lines 
are not fitting functions, they simply connect points to facilitate visualization. 
Figure 4: 
Extent of arsenic release from the two-column-systems for 97 days by tested Fe0 materials: (a) 
variation of the extent of As release with time, and (b) cumulative As release (mg). The lines 
are not fitting functions, they simply connect points to facilitate visualization. 
 
 26