The Pennsylvania State University
The Graduate School
Department of Civil and Environmental Engineering
ARSENIC REMOVAL FROM GROUNDWATER WITH IRON
TAILORED GRANULAR ACTIVATED CARBON PRECEDED BY
PRE-CORRODED STEEL
A Dissertation in
Environmental Engineering
By
Jiying Zou
© 2009 Jiying Zou
Submitted in Partial Fulfillment Of the Requirements for
The Degree of
Doctor of Philosophy
May 2009
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The dissertation of Jiying Zou was reviewed and approved* by the following Fred S. Cannon Professor of Environmental Engineering Dissertation Advisor Chair of Committee Brian A. Dempsey Kappe Professor of Environmental Engineering Paul Painter Professor of polymer Science Department of Material Science and Engineering John M. Regan Associate Professor of Environmental Engineering Peggy A. Johnson Professor of Civil Engineering Head of the Department of Civil and Environmental Engineering *Signatures are on file in the Graduate School
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ABSTRACT
ARSENIC REMOVAL FROM GROUNDWATER BY IRON
TAILORED GAC PLUS PRECORRODED IRON Ph.D. Candidate: Jiying Zou
Thesis Advisor: Fred S. Cannon, Professor The Pennsylvania State University (University Park, PA)
Department of Civil and Environmental Engineering
Arsenic of over 50 ppb level in drinking water could cause a lifetime risk of dying from
cancer for the consumer. Although, conventional granular activated carbon (GAC) has a very
limited capacity for removing arsenic, it was found that tailoring GAC by preloading iron could
enhance its bed life, when the iron tailored GAC was coupled with precorroded iron, the GAC’s
bed life could greatly enhanced.
For carbon tailoring, incipient wetness method and organic-iron preloading method were
employed. 1-3% Fe loading was achieved with organic-iron preloading method and 3-6% Fe
loading was achieved via incipient wetness method. Compared with virgin GAC, the citric
acid-iron preloaded GAC could extend the GAC’s bedlife by over 20 times to 7000 bed volumes
of 50 ppb arsenic containing water processed before 10 ppb breakthrough. The incipient wetness
method could further extend the GAC’s bedlife by 2 times.
Precorroded iron material, coupled with Organic carboxyl-Fe preloaded granular
activated carbons (GAC), have been appraised as an innovative technique for removing arsenic
from groundwater. The effective precorroded iron materials have included Galvanized Steel
Fittings and Perforated Steel Sheets. Rapid Small Scale Column Tests (RSSCT’s) and mini
column tests had been conducted to evaluate the arsenic removal capacity of the procorroded iron
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coupled with tailored carbon. The arsenic was found to be removed by both the iron column and
the GAC column, with GAC column as the major absorber. The pH, idling and precorrosion
protocol affect the iron release and arsenic removal. The combination of a precorroded iron
column followed by a iron – tailored GAC column removed arsenic to below 10 ppb for as much
as 248,000 bed volumes (BVs) at pH 6. These tests employed Rutland, MA groundwater with
native As of 47 ~ 55 ppb. Idling the system for one time extended the bed life of by 2 time, but
caused a short period arsenic breakthrough after column restart.
Arsenic removal in the GAC column was proportional to the iron amount accumulated in
the GAC column. The iron amount accumulated in the GAC column was generally controlled by
the operating pH, but was also affected by the precorrosion conditions of the iron and the idling of
the system. The arsenic removal in the iron column was generally higher with lower pH.
Moreover, as the column just started up, the removal was also controlled by the iron pre-corrosion
condition. A longer precorrosion period has promoted arsenic removal in the iron column. The
arsenic removal was generally lower with aged PSSs as the column just started, this was attributed
to the release of iron (hydr)oxides particles from the iron column; but with longer aging period of
more than 10 days, arsenic removal by aged PSSs could be greatly increased.
The precorrosion protocol influenced the formation of surface corrosion layer of the iron,
which in turn, affected how the iron was released and accumulated in the GAC column, especially
when the column just restarted. The morphology and structure of surface corrosion products on
precorroded steel sheets were studied via scanning electron microscope (SEM), X-ray diffraction
(XRD) and X-ray photoelectron spectroscopy (XPS) method. The results showed that the
morphology of surface corrosion products was highly related to iron release and arsenic removal.
v
Fresh precorroded steel sheets have a uniform surface, while aged precorroded steel sheets
exhibited a heterogeneous surface with some areas covered with thick, porous scales.
Lepidocrocite (γ-FeOOH), humboditine (FeC2O4(H2O)2) and clinoferrosilite (Fe1.5Mg0.5Si2O6) are
the mainly component on the fresh precorroded steel sheet, while goethite (α-FeOOH),
lepidocrocite and magnetite (Fe3O4)are the primary component of the aged precorroded steel sheet
surface. After they were employed in the column for arsenic removal, the primary phase on
precorroded steel sheet changed to goethite and magnetite, calcite was also detected. Arsenic
extracted from precorroded steel in iron columns contain only As(III) when the column was
operated at pH < 7 and had been idled. XAFS study of the GAC in pH 7.5 column indicated the
presence of reduced iron phases such as FeO and green rust, some As(V) has also been reduced to
As(III). Idling the columns for 7 days is promoted a reduction reaction in both the iron and the
GAC columns.
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TABLE OF CONTENTS
LIST OF TABLES ...........................................................................................................................ix LIST OF FIGURES ..........................................................................................................................x Acknowledgements........................................................................................................................ xii CHAPTER 1 .....................................................................................................................................1 CHAPTER 2 .....................................................................................................................................5
2.1 ARSENIC............................................................................................................................5 2.1.1 Background ..............................................................................................................5 2.1.2 Toxicology................................................................................................................6 2.1.3 Regulatory................................................................................................................7 2.1.4 Chemistry of Arsenic................................................................................................8
2.1.4.1 Immobilization of arsenic..............................................................................8 2.1.4.2 Arsenic speciation .......................................................................................10
2.1.5 Treatment Technologies .........................................................................................11 2.1.5.1 Modified conventional treatment methods – precipitation/coprecipitation methods ...................................................................................................................11
2.1.5.1.1 Precipitation by Alum.......................................................................12 2.1.5.1.2 Precipitation by Iron.........................................................................12 2.1.5.1.3 Lime softening .................................................................................13
2.1.5.2 Adsorption and Ion exchange reactions ...............................................14 2.1.5.2.1 Adsorption by activated carbon........................................................14 2.1.5.2.2 Adsorption by Activated Alumina ....................................................16 2.1.5.2.3 Adsorption by iron hydroxide/iron oxides........................................17 2.1.5.2.4 Adsorption by zero valent iron (ZVI)...............................................21 2.1.5.2.5 Adsorption by other low cost adsorbents..........................................24 2.1.5.2.6 Adsorption by Iron Based Sorbents..................................................24
2.2 ACTIVATED CARBON ...................................................................................................26 2.2.1 The Physical Characteristics and Surface Chemistry of Activated Carbon............26 2.2.2 Fe loading onto Activated Carbon for Arsenic Removal........................................28
2.2.2.1 Impregnation ...............................................................................................28 2.2.2.2 Precipitation .............................................................................................29 2.2.2.3 With Chelating Agent...............................................................................29
2.3 IRON CORROSION.........................................................................................................30 2.3.1 Corrosion process...................................................................................................30
2.3.1.1 Anaerobic iron corrosion.............................................................................30 2.3.1.2 Iron corrosion with the presence of oxygen or other oxidizer.....................32 2.3.1.3 Reduction of surface corrosion product on Fe0 ........................................33
2.3.2 Corrosion product characterization ........................................................................34 2.3.2.1 Corrosion scales on iron pipes in water distribution systems......................34
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2.3.2.2 Corrosion layers on the surface of iron used in contaminant removal ........36 2.3.3 Surface corrosion products and contaminant removal ...........................................39
2.3.3.1 Iron corrosion and contaminant reduction in PRBs ....................................40 2.3.3.2 Iron corrosion and contaminant adsorption in PBRs...................................41
2.4 THE MECHANISMS OF ARSENIC REMOVAL BY IRON BASED SORBENTS.......42 2.4.1 Adsorption of Arsenic by iron oxide/hydroxide—As removal mechanisms..........42 2.4.2 Arsenic removal by ZVI.........................................................................................43
2.4.2.1 Iron corrosion and arsenic removal on ZVI – the process...........................43 2.4.2.2 Rate controlling arsenic removal by ZVI ....................................................44
2.4.3 Redox reaction in ZVI system................................................................................45 2.4.4 Arsenic release .......................................................................................................46 2.5 REFERENCES.......................................................................................................48
CHAPTER 3 ...................................................................................................................................56
3.1 INTRODUCTION ............................................................................................................56 3.2 MATERIALS AND METHODS....................................................................................59
3.2.1 Materials.................................................................................................................59 3.2.2 Organic carboxylic-Fe preloaded carbon. ..............................................................60 3.2.3 Preparation of Fe-GAC through incipient wetness impregnation (IWI). ...............60 3.2.4 Adsorption Isotherm...............................................................................................61 3.2.5 Column tests...........................................................................................................61 3.2.6 Chemical Analysis..................................................................................................61
3.3 RESULTS AND DISSCUSIONS...................................................................................62 3.3.1 Organic acid-Fe loading onto GAC........................................................................62 3.3.2 Iron loading via incipient wetness method..........................................................63 3.3.3 Batch test of the citric acid-iron preloaded activated carbon ..............................64 3.3.4 Isotherm results ...................................................................................................65 3.3.5 Rapid Small Scale Column Tests ........................................................................67
3.4 CONCLUSIONS............................................................................................................68 3.5 REFERENCES...............................................................................................................68
CHAPTER 4 ...................................................................................................................................75
ABSTRACT............................................................................................................................75 4.1 INTRODUCTION .........................................................................................................76
4.1.1 Background ............................................................................................................76 4.1.2 Arsenic Removal Technology ................................................................................76 4.1.3 pH Effect on Arsenic Removal by ZVI and Iron (hydr)oxides ..............................77 4.1.4 Iron Corrosion and Iron Release ............................................................................77
4.2 MATERIALS AND METHODS....................................................................................79 4.2.1 Materials..............................................................................................................79 4.2.2 Citrate-Fe preloaded carbon................................................................................80 4.2.3 Iron Pre-corrosion ...............................................................................................80 4.2.4 Column tests........................................................................................................80 4.2. 5 Chemical Analysis..............................................................................................82
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4.3 RESULTS AND DISCUSSION.....................................................................................84 4.3.1 Arsenic removal with and without precorroded iron..............................................84 4.3.2 pH effect on Arsenic removal.................................................................................87 4.3.3 Idle Effect on Arsenic Removal .............................................................................92 4.3.4 Precorrosion iron amount effect .............................................................................96 4.3.5 Iron release and arsenic removal in iron column – A summary ..........................96
4.4 CONCLUSIONS............................................................................................................98 4.5 REFERENCES.............................................................................................................100
CHAPTER 5 .................................................................................................................................110
ABSTRACT..........................................................................................................................110 5.1 INTRODUCTION ....................................................................................................... 111
5.1.1 Surface corrosion layer and its effect on contaminant removal ........................ 111 5.1.2 Arsenic – iron redox reaction and As removal by ZVI .....................................112 5.1.3 As release ..........................................................................................................114
5.2 MATERIALS and METHODS ....................................................................................115 5.2.1 Precorroded steel sheets. ...................................................................................115 5.2.2 Scanning Electron Microscopy and energy-dispersive X-ray spectroscopy (SEM-EDS) tests...........................................................................................................116 5.2.3 X-ray Diffraction (XRD) Measurements. ............................................................117 5.2.4 X-ray Diffraction (XRD) Measurements of the powders collected from PSS surface. ..........................................................................................................................117 5.2.5 X-ray Photoelectron Spectroscopy (XPS) analysis. ..........................................117 5.2.6 Digestion of precorroded steel sheets for arsenic speciation.............................118
5.3 RESULTS and DISCUSSION....................................................................................118 5.3.1 SEM result............................................................................................................118 5.3.2 XPS results........................................................................................................120 5.3.3 XRD result ........................................................................................................123 5.3.4 Arsenic extraction from precorroded steel sheets in iron column.....................125 5.3.5 XAFS result.......................................................................................................125
5.4 Conclusions..................................................................................................................126 5.5 REFERENCES.............................................................................................................127
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LIST OF TABLES
Table 1.1 pKa Values of Arsenate and Arsenite .............................................................10 Table 3.1 Water quality characteristics of Cool Sandy Beach Groundwater (Rutland,
MA).................................................................................................................................62 Table 3.2 Fe loading result a..............................................................................................63 Table 3.3 Iron loading result b. .........................................................................................64 Table 3.4 Iron loading via incipient wetness method. ........................................................64 Table 3.5. Arsenic adsorption capacity with respect to water pH and carbon properties
.........................................................................................................................................66 Table 4.1: Configuration of Rapid Small Scale Column Tests (RSSCTs) and mini
columns...........................................................................................................................82 Table 4.2. Water quality characteristics of Cool Sandy Beach Groundwater (Rutland,
MA).................................................................................................................................83 Table 4.3. Column operating parameters and 10 ppb breakthrougha..............................85 Table 4.4. Arsenic distribution in GS #1 (iron - tailored GAC coupled with corrosion of
galvanized steel fittings) after 250,000 BV ..................................................................87 Table 4.5. Correlation of 10 ppb As breakthrough..........................................................92 Table 4.6. Fe release amount and arsenic removal in iron column................................97 Table 5.1 The pretreatment precorroded steel sheets and columnoperating conditions
.......................................................................................................................................116 Table 5.2 Quatitative analysis of precorroded steel sheets – atomic percentage of each
element .........................................................................................................................122
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LIST OF FIGURES
Figure 1.1 Molecular configurations of arsenite and arsenate..........................................55 Figure 1.2 (A) arsenate and (B) arsenite speciation as a function of pH..........................55 Figure 3.1. Adsorption Isotherm of Citrate-Fe preloaded GAC and Virgin GAC.
(A)Freudlich Isotherm (B) Langmuir Isotherm .........................................................70 Figure 3.2 Kinetics tests of CA-Fe (1.2) and CA-Fe-Mg (2.18). .....................................71 Figure 3.3 Pore volume analysis of virgin Ultracarb and various iron loaded
Ultracarb........................................................................................................................72 Figure 3.4 Kinetics tests of Fe loaded carbon made via incipient wetness method. .....73 Figure 3.6 RSSCT’s of amorphous iron oxide preloaded GAC. ....................................74 Figure 4.1 RSSCT of iron tailored GAC with (solid triangle, GS #1) and without
(hollow square, #1) corroded iron, both columns operated at pH 6±0.3. Rutland groundwater as influent (As 47~55 ppb, Fe < 3 ppb). Dashed line indicated where the column (solid triangle) was stopped and ceased for 6 days. ..............................103
Figure 4.3 pH effect on (A) Total Fe release. (B) Filterable Fe release. (C) Fe accumulated in GAC column. ....................................................................................105
Figure 4.4 Arsenic removal with no idle (open diamond, PS #3), one idle (solid reactangle, PS #1) and 3 idle (solid triangle, PS #2). (A) As effluent from GAC column. (B) As removal in Fe column. (C) Filterable arsenic from Fe column. Solid line indicate where PS #2 was stopped for 7 days, dashed line indicate where PS #1 was stopped for 7 days. All columns were operated at pH 6±0.3 ............................106
Figure 4.6 The effect of precorroded iron amount on arsenic removal. (A) Arsenic breakthrough curve. (B) Arsenic removal by Fe column. Both columns were operated at pH 7.5. Dashed line indicated where Run #6 was idled for 7 days, solid line indicated where PS #5 was idled for 7 days. ......................................................108
Figure 4.7 The effect of precorroded iron amount on (A) Total Fe release, (B) Filtrable Fe release, (C) Fe accumulation in GAC column. Both columns operated at pH 7.5........................................................................................................................................109
Figure 5.1 SEM of precorroded steel sheets (A) Fresh precorroded steel sheets – clean surface (B) Aged precorroded steel sheets – rough and rusty (C) Steel surface in PS # 6 (pH 7.5, idle once) – amorphous and uniform. (D) Steel surface in PS # 7(pH 7.5, idle once) – amorphous and uniform. (E) Steel surface in PS # 4 (pH 6-6.5, idle once) – rough with lots of precipitates. (F) Steel surface in PS #2 (pH 6, idle 3 times) – rough with lots of precipitates. (G) Steel surface in PS #2 (pH 6, idle 3 times) – porous (H) Steel surface in PS #4 (pH 6-6.5, idle once) – porous ............................131
Figure 5.2 Various crystals on surfaces of PSSs # 2 and PSSs #4. (A) to (E) Iron oxides, (F) Calcium oxides and iron oxides............................................................................132
Figure 5.3 Elements identification on precorroded steel sheets by XPS survey. Top spectrum– iron particles ultrasounded from PSS #2 (pH 6, idle three times); 2nd spectrum–iron particles detached from PSS #1 (pH 6, idle once); 3rd spectrum – fresh precorroded steel sheets; 4th spectrum – PSS #3 (pH 6, no idle); 5th spectrum – PSS #4 (pH 6 ~ 6.5, idle once). .................................................................................133
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Figure 5.4 XPS survey of Fe 2p peak. Note that Fe are FeOOH or iron oxide (Fe2O3 & Fe3O4). Top spectrum– iron particles ultrasounded from PSS #2 (pH 6, idle three times); 2nd spectrum–iron particles detached from PSS #1 (pH 6, idle once);3rd spectrum – fresh precorroded steel sheets; 4th spectrum – PSS #3 (pH 6, no idle); 5th spectrum – PSS #4 (pH 6 ~ 6.5, idle once)............................................................133
Figure 5.5 XPS survey of O 1s peak. Note that O is mainly hydroxide or iron oxide. 1st spectrum– iron particles ultrasounded from PSS #2 (pH 6, idle three times); 2nd spectrum–iron particles detached from PSS #1 (pH 6, idle once);3rd spectrum – fresh precorroded steel sheets; 4th spectrum – PSS #3 (pH 6, no idle); 5th spectrum – PSS #4 (pH 6 ~ 6.5, idle once). .................................................................................134
Figure 5.8 X-ray diffraction patterns of the powdered rust collected from steel chamber after runs PS #3 (no idling-top pattern); PS #2 ( thrice-idled-bottom pattern). Peak designations: G = Goethite-α-FeOOH; M = Magnitite Fe3O4; W = Wustite FeO; H = Humboltine (hydrous ferrous oxalate). ......................................135
Figure 5.10 Arsenic edges of GAC collected from Run PS#1 (pH 6 & idle once) after column stopped. Note that As(V)/As(III) ratio dropped from 3 (Rutland water) to 0.43 indicating As(V) reduction occurred..................................................................137
Figure 5.11 Arsenic edges of GAC collected from Run PS#2 (pH 6 & idle three times) after column stopped. Note that As(V)/As(III) is less than 1 indicating As(V) reduction occurred. .....................................................................................................137
Figure 5.12 Arsenic edges of GAC collected from Run PS#4 (pH 6 ~6.5 & idle once) after column stopped. Note that As(V)/As(III) ratio dropped from 3 (Rutland water) to 2 indicating As(V) reduction occurred. .....................................................138
Figure 5.13 Arsenic edges of GAC collected from Run PS#5 (pH 7.5 & idle once) after column stopped. Note that As(V)/As(III) ratio dropped from 3 (Rutland water) to 1.3 indicating As(V) reduction occurred....................................................................138
Figure 5.14 Fe edge from XAFS result of GAC from Run PS #5 and PS #6, both column were operated at pH 7.5 and idle once. Note that iron are best fit with FeO and green rust, indicating Fe(III) reduction occurred. ............................................139
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Acknowledgements
There are many people who have helped to complete this work, and I owe thanks to all of them.
First and foremost, I would like to express my appreciation to Dr. Fred S. Cannon, for giving me
this opportunity to do this work, and for his guide and support throughout this study. It has been a
pleasure working for him.
I would like to thank Dr. Brian A. Dempsey, Dr. Paul Painter, and Dr. John M. Regan for their
willingness to serve on my doctoral committee.
I would also like to thank Dr. Robert Parette, Dr. Weifang Chen, Fenglong Sun, Dr. Adam Redding,
Dr. Wang Yujue and Huang He for their help with some Laboratory procedure and all the graduate
students and staffs in the Kappe Laboratory for their kindly help.
This study was supported by the American Water Works Association Research Foundation. We
thank Siemens Water Technologies and Cool Sandy Beach Community Water System, Inc.
Rutland, MA for their support and service.
1
CHAPTER 1
INTRODUCTION
Throughout the world, arsenic is creating potentially serious environmental problems for
humans and other living organisms. Most reported arsenic problems are found in groundwater
water supply systems and are caused by natural processes such as mineral weathering and
dissolution resulting from a change in the geo-chemical environment to a reductive condition
(Astrup et al. 2000; Namasivayam and Senthilkumar 1998).
Millions of people in Western Bengal and Bangladesh have been drinking groundwater from
wells that contain 100-2,000 µg/L As, and many of these people have succumbed to diseases that
are caused by the arsenic contaminated ground water (Mandal et al. 1996). In the United States
over 35,000 people may be drinking water contaminated with more than 50 μg/L of arsenic and
over 2.5 million people could be supplied with water having arsenic levels over 25 μg/L (Smith et
al. 1992). Consumption of arsenic at the 50 µg/L level is estimated to cause mortality due to
lung, kidney, or bladder cancer in 1 out of every 1,000 to 10,000 people. Because of this concern,
the WHO in 1993 and USEPA in 2001, lowered the arsenic standard from 50 ppb to 10 ppb; and
the USEPA likewise has dictated that, all United States public water systems must comply with the
new 10 ppb standard as of January 1st 2006. In initial projections, USEPA and AwwaRF had
estimated the costs to meet this MCL to be $102 to 550 million per year (Frey 2000; USEPA
2001).
Modified conventional iron coagulation and filtration can be cost effective arsenic
2
removal for the larger municipalities; but such treatment may not be practical for small and very
small water utilities, which commonly employ simple well head treatment systems. Thus there has
been an urgent need to devise simple arsenic removal systems that are suitable for small utilities.
This research aimed to devise a means to remove arsenic from groundwater in a cost
effective manner for small and very small water systems. Activated carbon has been widely used
in the water treatment industry. The inherently simple features about activated carbon are that
GAC column is easy to operate and very applicable to small and very small water systems.
Studies revealed that iron (III) had high affinity toward inorganic arsenic species and very
selective in the sorption process. Granular ferric hydroxide (GFH) can be effective to remove both
As (V) and As (III) from aqueous solutions, it is physically weak, and will crumble and crush and
lost its capacity during the arsenic removal process. Recent researches are focused on creating
cheap and stable iron bearing adsorbents, such as iron oxide coated sand (Gupta, 2005), Iron oxide
impregnated activated carbon(Vaughan, 2005; Reed, 2000), and GAC based iron containing
adsorbent (Gu, 2005). GAC has large surface area, high pore volume, and rigid structure, which
renders it an ideal backbone for hosting a considerable quantity of iron, the authors aimed to
preload GAC with an effective way so as to improve the GAC’s arsenic removal capacity without
blocking GAC pored with too much unavailable iron.
Zero valent iron (ZVI) has been successfully used as a filter medium to remove arsenic
from water. ZVI’s bed life is not so long compared with GFH, but ZVI is relatively cost effective.
Researchers observed that the ZVI filter easily clogged with the iron oxidation; and to prevent this
clogging, iron filings need to be mixed with sand homogenously, but the homogenously mixed
ZVI/sand filter released effluent iron as high as 70 mg/L (Nikolaidis et al, 2003). Adding a
3
separate sand filter could control the iron effluent to less than 0.3 mg/L (bang, 2005); but this
added to system complexity.
The iron loaded GAC plus precorroded iron could be an effective way of arsenic removing
from groundwater. The precorroded iron could serve as an arsenic remover with its iron
(hydr)oxides corrosion products, it could also provide fresh iron for arsenic removal by adsorption
or coprecipitation in GAC column.
Research Objectives
The objectives of this research were:
1. To extend the bed life of activated carbon for arsenic removal.
2. To test the hypothesis that by preloading organic acid-Fe onto activated carbon surface, even
with just 1.2% Fe loading, the resultant carbon would be more effective in arsenic adsorption.
3. To test the hypothesis that when coupled with precorroded iron source, the iron preloaded
carbons are more effective in removing arsenic.
4. To test the arsenic removal as a function of pH, idle times, aging etc; and to study the influence
of iron release on arsenic removal during the column operation period.
5. To characterize the precorroded steel sheets so as to obtain a better understanding of how the
corrosion surface affect arsenic removal.
6. To test the hypothesis that arsenic removal in the precorroded iron plus iron loaded GAC
system is highly related to iron accumulated in GAC.
7. To study the arsenic and iron speciation in precorroded iron and GAC, so as to
explore the mechanism of arsenic and iron interaction.
4
REFERENCES Astrup, T., Stipp, S. L. S., and Christensen, T. H. (2000). "Immobilization of chromate from coal fly
ash leachate using an attenuating barrier containing zero-valent iron." Environmental Science & Technology, 34(19), 4163-4168.
Frey, M. (2000). " Cost implications of a lower arsenic MCL. Final report." Awwa Research Foundation.
Mandal, B. K., Chowdhury, T. R., Samanta, G., Basu, G. K., Chowdhury, P. P., Chanda, C. R., Lodh, D., Karan, N. K., Dhar, R. K., Tamili, D. K., Das, D., Saha, K. C., and Chakraborti, D. (1996). "Arsenic in groundwater in seven districts of West Bengal, India - The biggest arsenic calamity in the world." Current Science, 70(11), 976-986.
Namasivayam, C., and Senthilkumar, S. (1998). "Removal of Arsenic(V) from Aqueous Solution Using Industrial Solid Waste: Adsorption Rates and Equilibrium Studies." Ind. Eng. Chem. Res., 37, 4816-4822.
USEPA. (2001). "National primary drinking water regulations. Arsenic and clarifications to compliance and new source contaminants monitoring. Final Rule. Fed. Reg." 66(14).
5
CHAPTER 2
LITERATURE REVIEW
2.1 ARSENIC
2.1.1 Background
Throughout the world, arsenic is creating potentially serious environmental problems for
humans and other living organisms. Most reported arsenic problems are found in groundwater
water supply systems and are caused by natural processes such as mineral weathering and
dissolution resulting from a change in the geo-chemical environment to a reductive condition
(Astrup et al. 2000; Namasivayam and Senthilkumar 1998). Arsenic contamination is also
caused by human activities such as mining wastes, petroleum refining, sewage sludge, agricultural
chemicals, ceramic manufacturing industries and coal fly ash (Grossl et al. 1997; Manning
and Goldberg 1997; Viraraghavan et al. 1999).
Millions of people in Western Bengal and Bangladesh have been drinking groundwater
from wells that contain 100-2,000 µg/L As, and many of these people have succumbed to diseases
that are caused by the arsenic contaminated ground water (Mandal et al. 1996). In the United
States over 35,000 people may be drinking water contaminated with more than 50 μg/L of arsenic
and over 2.5 million people could be supplied with water having arsenic levels over 25 μg/L
(Smith et al. 1992). Consumption of arsenic at the 50 µg/L level is estimated to cause mortality
due to lung, kidney, or bladder cancer in 1 out of every 1,000 to 10,000 people. The World
Health Organization (WHO) announced that water containing more than 50 µg/L of arsenic is
unsuitable due to acute and chronic toxicity. Owing to epidemiological evidence linking arsenic
6
and cancer, the safe limit of arsenic in drinking water was reduced from 50 µg/L to 10 µg/L in
1993 by WHO (Johnston and Heijnen 2001; Tokunaga et al. 1999). The Clinton administration
promulgated a new maximum concentration level (MCL) of 10 µg/L As, and the EPA announced
on October 31, 2001 that public water supplies nationwide should reduce arsenic concentration
levels to below 10 µg/L by 2006. Complying with these stringent limits on arsenic could impose
a heavy financial burden on small public water system (Woods 2001). The overall objective of
this research has been to discern a less expensive means of removing arsenic from groundwater,
particularly for small municipalities.
2.1.2 Toxicology
Arsenic in drinking water may cause chronic arsenic intoxication (arsenicosis), which
may lead to harm of respiratory, digestive, renal circulatory, neural systems and internal organs
(ATSDR, 2000; IPCS, WHO, 2001). There are reported clinical effects and symptoms including
Raynaud’s syndrome, hypertension, cerebral infarction (Chen et al. 1995), damage of the
peripheral nerve bodies (Bansal et al. 1991), diabetes mellitus (Chen et al. 1994), and circulatory
disorders. In large regions of Bangladesh and West Benghal, India, the drinking water contains
arsenic concentrations as high as 1 mg/L; and as many as 50-65 million people are being poisoned
by this (Driehaus et al. 1998). In this area, 170,000 people have exhibited symptoms of chronic
arsenicosis(Paty et al. 1995).
The toxicity of arsenic is related to its chemical form and oxidation state. Inorganic
arsenic compounds normally are more toxic than organic compounds. The most significant
consequence of chronic arsenic intoxication is the induction of cancers in various organs.
Therefore, arsenic has been recognized as Class I human carcinogen and is a public concern due to
7
its widespread usage in both industry and agriculture. An area in Taiwan has had drinking water
sources in which arsenic concentrations ranged from 170 to 800 ppb. On the basis of the cancer
that was observed there, Smith et al. (1992) surmised that a 50 ppb arsenic level would translate to
a lifetime risk that 13 people per 1000 could die from cancer to the liver, lung, kidney, or bladder.
Arsenic also causes skin cancer at low concentrations; and it poisons the heart and gastrointestinal
tract at high concentrations.
Inorganic arsenic in low and micro molar doses can cause great genotoxicity. Sodium arsenite
is reported to induce chromosome aberrations, sister chromatic exchanges, and DNA-protein crosslinks
(Dong and Luo 1993).
2.1.3 Regulatory
Arsenic exceeds 10 ppb in at least 4000 community and non-community wells that
appear in more than 45 U.S. states (Frey and Edwards 1997). Half of all the states in America
have more than ten community wells that exceed this new limit; and they are (from roughly west
to east): Alaska, California, Oregon, Washington, Nevada, Idaho, Montana, Utah, Arizona, New
Mexico, Colorado, Texas, Oklahoma, Nebraska, South Dakota, North Dakota, Minnesota,
Wisconsin, Michigan, Indiana, West Virginia, New Jersey, Massachusetts, Vermont, Maine, and
Florida. (http://co.water.usgs.gov/trace/pubs/geo_v46n11/fig1.html; Welch et al. 2000). Many of
these wells service small and very small community water systems; and for the majority of these,
an arsenic removal facility will represent the first treatment system that the small providers have
had to install, above mere chlorination.
In early 2001, the USEPA published a revised arsenic standard of 10 ppb in drinking
water. This is considerably lower than the previous 50 ppb standard, which was established in
8
1942. All public water systems must comply with this 10 ppb standard within 5 years after this
rule was published (i.e. by 2006). The USEPA estimates that 3,000 community water systems
will need to take measures to lower their arsenic levels. The USEPA projects that throughout the
nation, it will cost these communities a cumulative $195-$675 million to comply; and this will
translate to $58-327 / household / year. Other individuals have projected yet higher compliance
costs. The cost burden for removing arsenic will be greatest on very small community systems,
which have traditionally employed no treatment beyond simple chlorination. Thus, there is great
need to devise new and innovative technologies that are inexpensive to use, easy to operate, and
durable through long-term use.
2.1.4 Chemistry of Arsenic
2.1.4.1 Immobilization of arsenic
Arsenic is of concern in water treatment because of its health effects. In general,
inorganic arsenic compounds are more toxic than organic arsenic compounds, and arsenite
(As(III)) is more toxic than arsenate (As(V)). The molecular structure of both arsenate and arsenite
are shown in Figure 1.1. The double-bonded oxygen in arsenate has a large effect on the ionization
due to the loss of hydrogen ions. The tendency of ionization is expressed by pKa (the
dissociation constant). For arsenic species, acid-base equilibria and pKa values are summerized
in Table 1.1. Figure 1.2 shows a schematic of the pH relationship between arsenic species and
illustrates the significant difference in the pH values of ionization steps that occur between
arsenate and arsenite. The pE-pH relationship is important for understanding the mobility of
arsenic species in groundwater and the effectiveness of arsenic treatment systems (Sun and Doner
9
1998). Inorganic arsenic species mainly exist in the +3 or +5 oxidation state. These oxidation
states are controlled by micro-organisms, redox potential, and pH, as well as reactions with other
chemical compounds in the soil and sediments such as iron sulfides, iron/manganese/aluminum
oxides and hydroxides, dissolved organic matter, etc. (Loeppert et al. 1995).
Components of soils and sediments are involved with ionic species in two types of adsorptive
reactions. The first type of adsorption reaction is based on ion exchange between charged
adsorptive sites and charged soluble ions. The second type is London Van der Waals bonding
and is the result of complex interactions between the electron clouds of molecules, molecular
polarity, and the attractive forces of an atomic nucleus for electrons beyond its own electron cloud.
The change of groundwater to a reductive condition could cause the arsenate attached in the soil or
sediment to be released into the liquid phase due to the chemical reduction of arsenate to arsenite
(especially predominant species H3AsO3 at below pH 9.22), which is more mobile due to its weak
adsorption on most mineral surfaces (Manning and Goldberg 1997; Scott 1991). The redox
alterations incurred when drawing reduced groundwater out of the ground can increase the arsenic
levels in the extracted water.
10
Table 1.1 pKa Values of Arsenate and Arsenite
Species Acid-base equilibria pKa
+− +⎯→← HAsOHAsOH 4243 2.20
+− +⎯→← HHAsOAsOH 24
-42
6.97 Arsenate (Arsenate)
+− +⎯→← HAsOHAsO 34
-24
11.53
+− +⎯→← HAsOHAsOH 3233 9.29
Arsenite (As(III)) +− +⎯→← HHAsOAsOH 2
3-32
12.10
2.1.4.2 Arsenic speciation
In natural environment, arsenic is rarely encountered as a free element. It can occur as
the semi-metallic element (Aso), arsenate (As5+), arsenite (As3+), arsine (As3-),
monomethylarsonate (MMAA), and dimethylarsinate (DMAA). The amount of each of these
species depends on the redox conditions and the nature of anthropogenic input and biological
activity. However, the organic (methylated) arsenic usually occurs at natural concentrations of
less than 1 μg/L and is not of major significance in drinking water treatment (Edwards 1994).
The most prevalent species of arsenic in drinking water are arsenate (+V valence) and arsenite
(+III valence). The occurrence, distribution, mobility and speciation of arsenic rely on a lot of
factors including the pH, reduction-oxidation reactions, distribution of other ionic species, aquatic
chemistry and microbial activity (Chen et al. 2005). Oxidation-reduction potential (Eh) and pH
are the most important parameters controlling arsenic speciation. The relationship between Eh, pH
and arsenic speciation are illustrated in Figure 1.3. The arsenate prevails in oxidized or anoxic
waters, while the arsenite prevails in reduced waters that also contain hydrogen sulfide. Clifford
11
and Ghurye (2000) compiled data indicating that arsenate represented more than 80% of the
arsenic species in the wells that were tested in California, New Mexico, Arizona, Taiwan, and
Chile; while arsenite predominated in Bangladesh, West Bengal, and Alaska wells. The acid/base
species of arsenate (V) are H3AsO4, H2AsO4-. HAsO4
2-, and AsO43-with corresponding pKa’s of
2.35, 6.75, and 11.6. This means that when the water pH is between 2.35 and 6.75, the H2AsO4-
species will prevail; and when the water pH is between 6.75 and 11.6, the HAsO42-
species will
prevail. Since both of these species that predominate in the near-neutral pH region are charged,
charge-based processes will remove arsenate. Moreover, when the pH is above neutral, the
arsenate exchange bonding will be greater (with the double negative charge) than below neutral
(with the single negative charge).
Similarly, the acid/base species of arsenite (III) are H3AsO3, H2AsO3-, HAsO3
2-, and AsO33-,
with pKa’s of 9.23, 12.11, and 13.41. This means that below pH 9.23, the non-charged H3AsO3
species will predominate, and charge-based processes will not remove arsenite. However, Ghurye
and Clifford (2000) observed that arsenite will oxidize to arsenate when it is exposed to chlorine
for one minute; while dosing with just three times the stoichiometrically required level of chlorine.
This means that in typical groundwaters, a chlorine dose of <0.1 mg/L would convert all arsenite
to arsenate. Most groundwater-based municipalities already have adopted chlorination; and thus
oxidation of As(III) to As(V) will not be an additional issue.
2.1.5 Treatment Technologies
2.1.5.1 Modified conventional treatment methods – precipitation/coprecipitation methods
Modified conventional treatment methods include coagulation with Alum or iron,
12
Fe-Mn oxidation and lime softening. Dissolved arsenic can form low solubility metal arsenates
(e.g. calcium arsenate) upon the addition of appropriate chemicals. This solid, commonly present
as a floc, can then be removed by sedimentation and filtration. A large fraction of the larger
utilities affected by the new MCL standards have already installed these methods. For them, Chen
et al. (Chen et al. 1999) proposed the lowest-cost As removal method is to remove arsenic and
reduce hardness with these conventional methods. But for very small water systems, it is not cost
effective to install a new coagulation or softening process for removing arsenic alone.
2.1.5.1.1 Precipitation by Alum
Aluminum compounds like Al2(SO4)3 could be used to remove arsenic from water. For
example, arsenic has been removed by coprecipitation with Al2(SO4)3. This process is most
effective for the removal of arsenate, but not arsenite. In order to remove As(III) by this process,
the arsenite must first be oxidized to As(V). Moreover, the removal of arsenate was only effective
below pH 6. At pH 8, the removal efficiency is only half of that at pH 5. At pH 5, with an initial
arsenic level of 400 ug/L and alum dose of 30 mg/L (as alum-MW 600), the percent arsenic
removal could reach 90% when chloride was added as an oxidizer (Kartinen and Martin 1995). So
a number of chemicals must be involved in this treatment process: chlorine would need to be
added to oxidize arsenite, an acid should be used to lower the water pH, and post-treatment of the
clarified water would need caustic to increase the water pH to an acceptable level.
2.1.5.1.2 Precipitation by Iron
The coprecipitation of arsenic with iron compounds is more effective than that with
alum compounds. With this process, adjustment of the pH does not appear to be as important as
13
alum. The best arsenic removal rates are obtained at pH of less than 8.5. As with alum
precipitation, chlorine still needs to be added to improve the arsenic removal efficiency from 50%
to 95% when As(III) is present (Kartinen and Martin 1995). Jekel et al. (1993) observed for raw
water that in the pH range of 6.1 to 7.2, a molar ratio of iron to arsenic of 20 is high enough for the
removal of arsenic. An arsenic removal process based on the fluid-bed technology was proposed
by Stamer and Nielsen (2000). The fluid bed contains an inert carrier medium, typically fine
quartz sand grains. Ferrous ions and stoichiometric amount of hydrogen peroxide was added
continuously with the arsenic-containing water. As a result, iron oxyhydroxide precipitated on the
surface of the sand grains as a very dense coating which provided an active and fresh source for
arsenic to bind with. The arsenic that accumulated on the granules was about 50 g/kg. Several
operations occurred within a single reactor, including adding chemicals, mixing, precipitation, and
solid/liquid separation.
2.1.5.1.3 Lime softening
Lime softening, to reduce carbonate hardness, is a successful technology for achieving greater
than 90% As(V) removal. Arsenic in the pentavalent arsenate form is more readily removed than
the trivalent arsenite form. The optimum pH for As(V) removal by softening is approximately
10.5 and the optimum pH of As(III) removal is approximately 11.0 (USEPA 2000a, 2000e).
Addition of iron improves As(V) removal and the presence of sulphate and carbonate in the raw
water does not interfere with As(V) removal at pH 11. As(V) removal, however, is reduced in the
presence of carbonate at pH 10 to 10.5 and the presence of orthophosphate at pH less than 12.0.
However, where levels in water are high, these methods may not remove sufficient arsenic to be
adequately protective and various alternative technologies have been developed /adapted to reduce
14
arsenic to trace levels.
Many small water treatment systems are not set up to employ multi-unit operation systems
with alum or iron coagulant that involve precipitation, flocculation, settling and filtration. Thus,
for the work herein, we have sought a less complicated process.
2.1.5.2 Adsorption and Ion exchange reactions
Some solids, including iron and aluminium hydroxide flocs, have a strong affinity for
dissolved arsenic. Arsenic is strongly attracted to sorption sites on the surfaces of these solids, and
is effectively removed from solution. Ion exchange involves the reversible displacement of an ion
adsorbed onto a solid surface by a dissolved ion.
Several adsorption technologies have successfully been applied for the removal of
arsenic; and the research herein has built on this prior research. In general, adsorption processes
require only simple operations, low costs, and little maintenance. They are adaptable to various
kinds of well sites. Low dosing of chemicals is required (if any) and also the amount of residuals
is low when adsorbents with high adsorption capacities are used.
2.1.5.2.1 Adsorption by activated carbon
A lot of researches have been done regarding arsenic removal by ommercial activated
carbon and synthesized activated carbon (Chuang et al. 2005; Daus et al. 2004; Gu and Deng 2005;
Huang and Fu 1984; Lorenzen et al. 1995; Nagarnaik et al. 2003; Olesen et al. 2004; Rajakovic
1992; Vaughan and Reed 2005). Van der Waals forces are claimed to mechanisms of arsenic
adsorption by the commercial activated carbon (Eguez and Cho 1987) and only 1~2 wt% arsenic
removal can be achieved through this way; while after impreganated with copper, the arsenic is
removed by forming insoluble metal arsenate with the impregnated copper and by adsorption of
15
carbon through Van der Waals forces (Rajakovic, 1992). Campos, et al. (2002) studies the arsenic
removal by combination of GAC and carbon steel wool, the iron-As electrochemical reactions is
claimed to be the mechanisms of arsenic removal in this system.
Metal treated activated carbon for arsenic removal including Fe containing GAC (Gu et al.
2005), zirconium loaded GAC (Daus et al. 2004), Fe oxide – impregnated activated carbon
(Vaughan and Reed 2005) and iron salt solution pretreated GAC (Huang and Fu 1984). All these
metal treated activated carbon are able to improve arsenic removal, a 10 fold increase was claimed
by Huang, et al.(1984). The optimum pH for arsenate removal is 5-7, while that for arsenite
removal is 9-11.
Granular activated carbon (GAC) based iron containing adsorbents was developed by Gu
et al. (2005). This media was made in 2 steps, first Fe (II) was adsorbed onto GAC, and then the
Fe (II) was oxidized to Fe (III) by O2, H2O2 or NaClO. When lignite based carbon was employed,
the iron loading could reach 7.9%, but the best arsenic removal was achieved with an Fe loading
of 6%. The BET surface area, pore volume and porosity of the carbon are decreased after Fe
loading, indicating some micropores and mesopore blockage. X-ray diffraction of the virgin GAC
and Fe loaded GAC are the same, which indicated that the iron loaded was in amorphous state.
The adsorbent could remove arsenic to 7500 bed volumes before reaching 10 ppb breakthrough
when influent contained 50-60 ppb of As (V) or As (III) (Gu, et al. 2005).
Vaughan et al.(2005) developed iron oxide impregnated activated carbon. By precipitating
iron salts onto activated carbon (presumably primarily external loading), these authors could
achieve an iron loading of 7%. At pH 7, with 1 mg/L arsenic, and 0.2 g/L Fe oxide impregnated
activated carbon, this adsorbent could get an As (III) adsorption of 4.7 mg/g or an As (V)
16
adsorption of 4.5 mg/g (Vaughan and Reed 2005).
2.1.5.2.2 Adsorption by Activated Alumina
Activated alumina, prepared by thermal dehydration of aluminum hydroxide, has a high
surface area and a distribution of both macro- and micropores. Arsenic(V) sorption occurs best
mostly between pH 6.0 and 8.0 where AA surfaces are positively charged. As(III) adsorption is
strongly pH dependent and it exhibits a high affinity towards AA at pH 7.6 (Singh and Pant 2004).
Singh and Pant (2004; 2006) removed arsenites from water with AA and iron
oxide-impregnated AA (Kuriakose et al. 2004). The effect of adsorbent dose, pH, and contact time
were investigated. As(III) removal was strongly pH dependent. Both Freundlich and Langmuir
adsorption isotherms were fit by the experimental data. Adsorption kinetics were governed by a
pseudo first order rate equation in both cases. The adsorption capacity of iron oxide impregnated
AA (12 mg/g) was much higher than AA (7.6 mg/g) (Kuriakose et al. 2004).
Conventional AA has ill-defined pore structures, low adsorption capacities and exhibits
slow kinetics (Kim, 2004). An ideal adsorbent should have uniformly accessible pores, a three
dimensional pore system, a high surface area, fast adsorption kinetics and good physical and/or
chemical stability. Kim et al, developed a mesoprous alumina (MA) with a large surface area
(307m2/g) and uniform pore size (3.5 nm) for arsenic removal (Kim et al. 2004). A sponge-like
interlinked pore system was developed through a post-hydrolysis. The resulting MA was insoluble
and stable at pH 3–7 and its adsorption kinetics were rapid. The maximum As(V) uptake by MA
was seven times higher (121 mg of As(V)/g and 47 mg of As(III)/g) than that of conventional AA.
This adsorbent’s surface area did not greatly influence the adsorption capacity. The key factor is a
uniform pore size.
17
Activated alumina is very efficient and can be regenerated in situ to extend the useful life.
However, sorption efficiency is highest only at low pH and arsenites must be pre-oxidized to
arsenates before adsorption.
2.1.5.2.3 Adsorption by iron hydroxide/iron oxides
Iron oxides, oxyhydroxides and hydroxides, including amorphous hydrous ferric oxide
(FeO-OH), goethite (α-FeOOH) and hematite (α-Fe2O3), are promising adsorbents for removing
both As(III) and As(V) from water (Roberts et al. 2004; Saha et al. 2005; Wilkie and Hering 1996).
Amorphous Fe(O)OH has the highest adsorption capability since it has the highest surface area.
Surface area is not the only criterion for high removal capacities of metal ions and other
mechanisms (ion exchange, precipitation) play an important role. Most iron oxides are fine
powders that are difficult to separate from solution. Therefore, the EPA has proposed iron
oxide-coated sand filtration as an emerging technology for arsenic removal at small water
facilities (Thirunavukkarasu et al. 2003). Another shortcoming of amorphous FeOOH is its
tendency to form low surface area crystalline iron oxides during preparation, greatly reducing its
As removal capacity. Different types of ferrihydrites, ion hydroxide and iron oxides were prepared
and tested. Some recent studies are now discussed.
Raven et al. (Raven et al. 1998)1998) compared the adsorption behavior of arsenite and
arsenate on ferrihydrite [(Fe3+O3·0.5(H2O)]. Arsenate adsorption was faster at low As
concentrations and low pH. Adsorption maxima at pH 4.6 (pH 9.2 in parentheses) of 0.60 (0.58)
and 0.25 (0.16) molAs/molFe were achieved for arsenite and arsenate, respectively. Overall
arsenite and arsenate have strong affinities for ferrihydrite, and arsenite retained in much larger
amounts than arsenate at high pH (approximately >7.5) or at high As concentrations in solution.
18
The high arsenite retention was due to the fact that ferrihydrite was transformed to a ferric arsenite
phase and not simply adsorbed at the surface.
Hydrous ferric oxide (HFO) has been extensively studied as a promising adsorptive
material for removing both arsenate and arsenite from aqueous phase due to its high iso-electric
points (8.1) (Dixit and Hering 2003) and selectivity for arsenic species. Ranjan et al. (2003)
synthesized hydrous ferric oxide, for arsenic sorption. As(V) sorption strongly depended on the
system’s concentration and pH, while As(III) sorption was pH insensitive. As(III) required less
contact time to attain equilibrium.
Ford (2002) studied the influence of laboratory controlled aging on the stability of arsenate
coprecipitated with hydrous ferric oxide (HFO). The relationship between the transformation of
HFO and the stabilization of arsenic was addressed. The rate of arsenate stabilization
approximately coincided with the rate of HFO transformation at pH 6 and 40 ◦C. Extraction data
and X-ray diffraction results confirmed that hematite and goethite were the primary crystalline
products. HFO transformation was highly related to the arsenic loading amount. The
transformation was significantly retarded at, or above, arsenate loadings of 29.4 mg As/kg HFO.
However, HFO transformation proceeded rapidly to hematite (XRD studies) for arsenate solid
loadings of 4.2 and 8.4 g As/kg HFO.
The HFO is mostly used as powder and its low hydraulic conductivity (Zeng 2003) made it
unsuitable for use in column applications. However, in the purification technology, the
precipitates cannot be advantageously used since large amounts of hazardous waste materials are
produced daily and the disposal cost can be tremendous. To overcome this disadvantage,
granulation techniques have been developed.
19
Granular ferric hydroxide (GFH) consists of a poorly crystallized β-FeOOH (250-300 m2/g
surface area; 75%-80% porosity). It was synthesized from ferric chloride solution by
neutralization and precipitation with sodium hydroxide, followed by centrifugation and
granulation under high-pressure (Thirunavukkarasu et al. 2003, Gu et al. 2005). There are no
drying procedures in the preparation, so GFH pores are filled with water, which renders it high
amount of adsorption sites and thus high adsorption capacity. The adsorption of As(V) present in
concentrations ranging from 100 to 750 μg/L over the pH range of 4–9 on ferric hydroxide
(GFH) was investigated (Lenoble et al. 2002). The adsorption decreased as the pH of the solution
increased, and optimal adsorption was at pH 4. The competitive effect of phosphate on the uptake
of arsenate at pH 4 by GFH was also investigated. GFH had a greater affinity for arsenate
adsorption compared to phosphate.
At slightly above neutral pH values, GFH was discovered to have nearly equal
adsorption capacity for both arsenate and arsenite (Driehaus et al. 1998). Driehaus et al. explored
this process’s arsenic removal capacity with influent water which has pH 8.2 and arsenic content
of 16 ug/L. With GFH grain sizes 0.2-0.4 mm and EBCT 32 second, this process is able to remove
As for 34,000 bed volumes until the 10 ppb As breakthrough occurs (Driehaus et al. 1998).
Westerhoff et al. had conducted several lab columns with GFH. The results revealed that at pH
7.6, GFH could remove As from 50 ppb to 10 ppb MCL for 25,000 BV when the EBCT is 2.5
minutes (Westerhoff et al. 2005). But suppliers have commented that these iron oxide granules can
crumble and disintegrate when they experience prolonged use, Selvin reported that GFH with a
media particle size at 0.8-2.0 mm will need a backwashing every 5000 Bed volumes (Selvin et al.
2002). Also, after backwashing, a significant amount of head loss pressure will build up in the
20
system (Gu,et al., 2005).
Arsenic adsorption on magnetite (Fe3O4) nanoparticles was conducted by Mayo et
al.(Mayo et al. 2007). The effect of Fe3O4 particle size on the adsorption and desorption behavior
of both As(III) and As(V)was reported. As the particle size was decreased from 300 to 12 nm the
adsorption capacities for both As(III) and As(V) increased nearly 200 times.
Some researches are conducted regarding redox reaction and arsenic removal, as
discussed below:
Roberts et al. (2004) studied the arsenic removal by oxidizing Fe(II) to iron(III) (hydr)oxides by
aeration. Arsenic removal was achieved by coprecipitation with the thus formed iron(III) species.
Application of Fe(II) instead of Fe(III) was advantageous, because the dissolved oxygen used for
oxidation of Fe(II) causes partial oxidation of As(III). Furthermore iron(III) (hydr)oxides formed
in this way have higher sorption capacities. Multiple additions of Fe(II) followed by aeration
further increase As(III) removal. Lee and Hering (2003) investigated the stoichiometry, kinetics,
and mechanism of arsenite [As(III)] oxidation and coagulation by ferrate [Fe(VI)]. As(III) was
oxidized to As(V) (arsenate) by Fe(VI), in a 3:2 (As(III):Fe(VI)) stoichiometry. As (III) oxidation
with Fe (VI) was first-order in both reactants. An oxygen transfer mechanism was proposed for
the oxidation of As(III) by Fe(VI). Fe(VI) was very effective in arsenic removal from water at a
low Fe(VI) dose level (2.0 mg/L). In addition, the combined use of a small amount of Fe(VI)
(below0.5 mg/L) and Fe(III) as a major coagulant was effective for removing arsenic.
Electrochemical peroxidation (EPC) at steel electrodes with H2O2 is an emerging As(III)
remediation technology (Arienzo et al. 2002). ECP effectively removed arsenic from the
aqueous solutions, with > 98% of the As(III) adsorbed on solid hydrous ferric oxides. The reaction
21
is rapid and removal was complete within 3 min, independent of the initial aqueous pH (3.5–9.5).
The optimal operating conditions were pH < 6.5, [H2O2] = 10 mg/L and a process time ≤3 min.
Binding of arsenite to ferric hydroxide using several density functional theory methods
was investigated by Zhang et al. (2005 a). Calculated and experimentally measured As–O and
As–Fe bond distances confirmed that arsenic formed bidentate and monodentate corner-sharing
complexes with Fe(III) crystalline. Edge-sharing As(III) complexes were less energetically
favored and had As–O and As–Fe distances that deviated from experimentally measured values
more than corner-sharing complexes.
2.1.5.2.4 Adsorption by zero valent iron (ZVI)
The use of Fe0 to remove arsenic has been actively investigated by many groups (Bang
et al. 2005; Farrell et al. 2001; Kim et al. 2003; Krishna et al. 2001; Lackovic et al. 2000; Lien and
Wilkin 2005; Manning et al. 2002; Melitas et al. 2002; Nikolaidis et al. 2003; Su and Puls 2001b;
Su and Puls 2003; Su and Puls 2004).
In this method, arsenic is adsorbed onto corrosion products of zero-valent iron (ZVI) as the
ZVI converts so such species as iron (oxyhydr) oxide. Studies have shown that the performance of
ZVI is limited by its initial removal capacity and any additional capacity that may come about
after iron metal corrodes in water (Lackovic, Nikolaidis et al. 2000; Su and Puls 2001). Possible
arsenic removal processes in zero-valent iron system include surface adsorption onto corrosion
products, e.g. iron (oxyhydr)oxides (Dixit and Hering 2003), precipitation such as formation of
symplesite (Fe3(AsO4)2· 8H2O) (Nikolaidis, et al. 2003) , co-precipiration ( e.g. arsenic
co-precipitation with carbonate green rust) (Lien and Wilkin 2005) or redox reaction such as As
22
(III) oxidized to As(V) by corrosion products or impurities such as MnO2 (Manning, et al. 2002;
Melitas, et al. 2002). Both As (III) and As (V) could be removed effectively from aqueous solution
using zero valent iron according to studies by Nikolaidis et al. (2003). The arsenic removal
capacity was determined to be approximately 7.5 mg/g Fe or 0.1 mmol As/g Fe. Melitas et al.
(2002) found that rates of arsenate removal by ZVI were highly dependent on the continuous
generation of iron oxide adsorption sites by comparing arsenate removal by freely corroding and
cathodically protected iron. The presence of 100 μg/L As(V) decreased the iron corrosion rate by
up to a factor of 5 compared to a blank electrolyte solution. However, increased As(V)
concentrations (100–20,000 μg/L) caused no further decrease in the iron corrosion rate. Arsenate
removal kinetics ranged between zeroth- and first-order versus the aqueous As(V) concentration.
The surface area exposed plays a major role in both the adsorption kinetics and
capacities. Kanel et al. (2005, 2006) synthesized nanoscale (1–120 nm diameter) zerovalent iron
(NZVI) for rapid, first order As(III) and As(V) removal (kobs = 0.07–1.3 min−1) This rate was
about 1000 times faster then that of micron-sized iron. The maximum As(III) adsorption
Freundlich capacity was 3.5 mg of As(III)/g for NZVI. Light scattering electrophoretic mobility
measurements confirmed a NZVI-As(III) inner-sphere surface complexation mechanism.
Bang et al. (2005) utilized zero-valent iron filings for arsenic remediation. Arsenic removal
was dramatically affected by oxygen content and pH. Arsenate removal by Fe(0) filings was faster
than arsenite under oxic conditions. Greater than 99.8% of the As(V) was removed whereas 82.6%
of the As(III) was removed at pH 6 after mixing for 9 h. When dissolved oxygen was removed by
nitrogen purging, less than 10% of the As(III) and As(V) was removed. High dissolved oxygen
content and low solution pH increased the iron corrosion rate. Thus, arsenic removal by Fe0 was
23
attributed to adsorption onto iron hydroxides generated from Fe0. The As(III) removal rate was
higher than that for As(V) when iron filings (80–120 mesh) were mixed with nitrogen-perged
arsenic solutions in the pH range of 4–7 (Bang et al. 2005). XPS spectra demonstrated As(III)
surface reduction to As(0). As(V) was reduced to As(III) with Fe0 under anoxic conditions, but no
As(0) was detected in solution after 5 days.
Zero-valent iron mechanisms for arsenate removal from drinking water were also
investigated by Farrell et al. (2001). Batch experiments using iron wires suspended in anaerobic
arsenate solutions were performed to determine arsenate removal rates as a function of the
arsenate solution concentration. Batch reactor removal kinetics were described by a dual-rate
model. Arsenate removal was pseudo-first-order at low concentrations and approached zero-order
in the limit of high arsenate concentrations. Arsenate decreased iron corrosion rates as compared
to those in a blank 3mM CaSO4 electrolyte solutions.
Karschunke and Jekel (2002) investigated arsenic removal with fine iron wool. The iron
wool was corroded in oxygenated water, and the iron hydroxide formed by the corrosion adsorbed
As (V). They also tested if galvanic corrosion could improve arsenic removal by enhancing iron
release. The results revealed that when combined with a sand filtration process, the galvanic
corrosion system could remove Arsenic from 500 ppb down to 50 ppb for 32,000 Bed Volumes.
According to Karschunke and Jekel, there are two major concerns that prevent this process from
implementing, they are the cathodic formation of hydrogen and the risk of copper release when the
contact breaks down. Nikolaidis, et al. (2003) conducted pilot tests with a ZVI filter which
contains ZVI plus sand with a weight ratio of 1:1. At pH 6, this filter removed arsenic from 294
ug/L down to 20 ug/L for 18,000 BVs ~ 21,600 BVs. Bang et al. (2005) removed Arsenic from
24
100 ug/L down to < 10 ppb for 34, 000 BVs by using an iron filter followed by a sand filter. ZVI’s
bed life is not so long compared with GFH, but ZVI is relatively inexpensive, only $0.2/lb
compared to $4.6/lb for GFH (Westerhoff et al. 2005). Researchers also pointed out that the ZVI
filter easily clogged with the iron oxidation; and to prevent this clogging, iron filings need to be
mixed with sand homogenously, but the homogenously mixed ZVI/sand filter was found to exhibit
heavy iron leachant – as high as 70 mg/L was reported by Nikolaidis et al (2003). With a separate
sand filter, Bang et al. (2005) was able to control the iron effluent to less than 0.3 mg/L; but this
added to system complexity.
2.1.5.2.5 Adsorption by other low cost adsorbents
Clays, silica, sand, etc. are in fact low-cost arsenic adsorbents (and substrates). They are
available worldwide. These can also be regenerated in situ. Unfortunately, they have lower
adsorption efficiency than most of the other adsorbents. Additionally, other water contaminants
can deactivate the clays, further lowering the sorption efficiency. Solid wastes like chars, red mud,
furnace slag, fly ash, are a vexing societal problem mandating attention to recycling. Recycled
product quality is not always high or recycle may not be feasible. However, conversion of solid
wastes into effective low-cost adsorbents for wastewater treatments could decrease costs for
removing arsenic. Only initial laboratory evaluations of the adsorptive capacities of adsorbents
developed from such wastes were available by far. Future studies are needed for the field
application with careful cost evaluations.
2.1.5.2.6 Adsorption by Iron Based Sorbents
A number of novel removal technologies are under development. Many new materials
are being tested for their ability to remove arsenic, including low-tech iron coated sand and
25
greensand, novel iron-based sorbents, and specially engineered synthetic resins. Studies revealed
that iron (III) has a high affinity toward inorganic arsenic species; and it is very selective to
arsenic in the sorption process. Recent research has focused on creating cheap and stable iron
bearing adsorbents. Some of these media are discussed below.
(1) Iron oxide coated sand. For this media, the sand only serves as an (inert) support, and
it is the iron oxide coating that removes arsenic. Gupta et al. (2005) observed that the adsorption
of As (III) onto iron oxide coated sand could reach 0.03 mg/g at pH 7.5.
(2) Iron oxide impregnated activated carbon. By precipitating iron salts onto activated
carbon (presumably primarily external loading), these authors could achieve an iron loading of
7%. At pH 7, with 1 mg/L arsenic, and 0.2 g/L Fe oxide impregnated activated carbon, this
adsorbent could get an As (III) adsorption of 4.7 mg/g or an As (V) adsorption of 4.5 mg/g
(Vaughan and Reed 2005).
(3) Fe (III)-loaded cellulose sponge. The sponge is claimed to contain free available
ethyleneamine and iminodiacetate groups, which interact with Fe (III) by chelation and ion
exchange. The Fe (III) loading capacity was 0.25 mmol Fe/g sponge, which corresponded to a
1.4% Fe content. The media had an As (V) adsorption capacity of 1.83 mmol As /g sponge; and an
As (III) adsorption capacity of 0.24 mmol As/g sponge (Munoz, et al. 2002).
(4) Granular activated carbon (GAC) based iron containing adsorbents. This media was
made in 2 steps, first Fe (II) was adsorbed onto GAC, and then the Fe (II) was oxidized to Fe (III)
by O2, H2O2 or NaClO. When lignite based carbon was employed, the iron loading could reach
7.9%, and they proposed that the impregnated iron was mostly in coordinated form with various
functional groups on GAC, but not in polymeric iron hydroxide form. The adsorbent could remove
26
arsenic to 7500 bed volumes before reaching 10 ppb breakthrough when influent contained 50-60
ppb of As (V) or As (III) (Gu, et al. 2005).
(5) Amorphous iron oxide preloaded media. Amorphous oxide had been loaded on sand
(Vagliasindi, et al, 1998), and in incinerator melted slag (Zhang and Itoh 2005; Zhang et al. 2005),
iron has also been embedded in a macroporous cation exchange bead (DeMarco, et al. 2003) and
the latter two had been tested for arsenic removal. The porous cation exchange resins could obtain
an iron loading of 25% based on dry bead weight, and 50% of the iron loading is in the form of
FeOOH. This product could remove arsenic from 50 ppb down to 10 ppb for 45,000 BVs.
DeMarco, et al. (2003) also pointed that the HFO agglomerates in the beads were very stable,
turbulence and mechanical stirring did not result in any loss of HFO. The iron oxide loaded on
melted slag was believed to be chemically bonded with the Si inside the slag, which prevented the
crystallization of FeOOH, and Zhang (2005) claim this material had better arsenic removal
(arsenate of 78.5 mg/g slag) capacity than FeOOH.
Some of this research is promising, but these technologies are still under development and
not tested in the field. These may also be too expensive for application.
2.2 ACTIVATED CARBON
2.2.1 The Physical Characteristics and Surface Chemistry of Activated Carbon
This research focused on activated carbon as the porous media for iron loading. Activated
carbon is comprised of graphene planes that are packed together and then bonded together. Each
graphene plane consists of a benzene ring lattice. The pi-electrons in this lattice can exhibit
27
pi-bonding energy with other graphene layers, and also with adsorbents. Hydrophobic organic
compounds prefer to adsorb onto these non-polar regions of the activated carbon, rather than
staying in polar water. The edges of the graphene planes can host a number of oxidized sites,
including the oxygenated substituents: carboxyls, phenolics, carbonyls, and lactones. In contrast,
the interiors of the graphene planes can pose a localized low-redox potential; since N can be
substituted for C in the lattice structure, creating an electron-rich region (Leon and Radovic 1994).
Activated carbon is created by thermally treating carbon-based solids, such as bituminous
coal, lignite coal, or wood. The pyrolysis step in thermal treatment creates narrow fissures
between graphene planes; and the oxidation step facilitates the gasification of some graphene
layers so as to create slightly wider spaces between the layers. Following activation, the edge
sites can be left with incomplete electron configurations; and are therefore reactive. Oxygen can
chemisorb to such reactive sites, and form oxygenated groups (Nowack and Stone 2002).
The spaces between graphene planes are generally planar, or slit-shaped. In conventional
bituminous granular activated carbons (GAC’s), the large majority of pores have widths of 4-30 Å;
and organic molecules can just barely fit into these pores. Iron-citrate complexes (with tag-along
waters of hydration) have dimensions of 5-20 Å. Perhaps the most useful pore widths for
adsorbing molecules are 1 to 13 times their dimension, i.e. 5-250 Å (Nowack et al. 2002, Krupa
and Cannon 1996, Rangel-Mendez et al. 1996). Based on a mass/volume basis; a single
continuous flat graphene plane would exhibit a surface area (top and bottom) of 2000 m2/g; and
commercial activated carbons generally have N2BET surface areas of 900-1200 m2/g. If
rigorously accurate, this indicates that about half of all the graphene planes have two surfaces
exposed. These surface areas are 2-3 times higher than for granular iron media.
28
The large surface area, high pore volume, and rigid structure of GAC render it an ideal
backbone for hosting iron species or iron complexes.
2.2.2 Fe loading onto Activated Carbon for Arsenic Removal
2.2.2.1 Impregnation
Pore volume impregnation, and specifically incipient wetness, is one of the most prevalent
methods used in the area of catalyst production. The impregnation procedure comprises two steps.
First, saturate the pores of the porous support with aqueous metal salt solution; then dry and
calcine the impregnated support in order to convert the metal salt to metal oxide. It had been
reported that 10-12 % Cu impregnation could be achieved via this method (Montanari et al. 1997,
Marchi et al, 2003).
The impregnation of carbons brings to mind the impregnation by metals, a subject that has
been widely studied in heterogeneous catalysis. Metals and their oxides, dispersed as small
particles on high surface area carbons and other supports, are being used as catalysts for various
industrial applications. The impregnation of metals in carbonaceous materials modifies the
gasification characteristics and varies the porous structure of the carbon product. For example,
loading of iron(III) was explored by Na et al. (2002) to change the surface chemistry of activated
carbon. The theory behind this is that the carbon surface is more positively charged by the iron
loading.
The impregnation procedure affords relatively large metal crystallites concentrated at the
surface of the support particle (Vaishya and Gupta 2003). Crystallized iron oxides are far less
29
effective for arsenic removal compared to amorphous iron oxides.
2.2.2.2 Precipitation
The precipitation method, also called precipitation-deposition, comprises inducing
precipitation of a dissolved metal species which then deposits upon a finely powdered solid
support. Conventionally, the most widely studied chemical method to prepare iron oxides has been
the precipitation of iron ions from aqueous solutions of their nitrate, chloride, perchlorate, or
sulfate salts (Lee et al. 1996). The precipitation of ferric ions is usually driven by thermolysis or
by the addition of base to the aqueous solution. The characteristics of the final product, i.e. oxide
phase, particle size and surface area, depend highly on the precipitation conditions, especially the
concentration of the iron ions, the nature of the counter-ions present, and the pH of the solutions.
Goethite (α-FeOOH), ferrihydrite (Fe5HO8·H2O) or akagenite (β-FeOOH) are usually the initial
precipitates, which are converted to low surface area α-Fe2O3 under moderate heat treatment
(Schwertmann 1991).
High metal loading can be achieved via precipitation (e.g., 50% or more). But aggregation
might accompany this, such that the metal is not uniformly distributed on the support. Khaleel
(2004) had discovered that the metal crystallites size produced by this method can be as large as
30 nm.
2.2.2.3 With Chelating Agent
Surface modification of activated carbon by immobilizing organic compounds is recognized
as an effective approach for enhancement of heavy metal removal. Surface modification of GAC
by citric acid had been reported to enhance copper adsorption by 140% (Chen et al. 2003). The
30
presence of ethylenediamine tetra-acetic acid (EDTA) was reported to improve the cadmium
adsorption by 2-3 times (Preston et al, 1995). It is noted that iron, citric acid, EDTA and fatty acids
are non-toxic and commonplace in water and foods. No primary drinking water standards exist for
any of these species.
Citric acid is found to serve as a redox species, and it also facilitates considerably greater
internal sorption of iron than could be achieved with mere salts such as FeCl3 (Chen et al. 2003).
Liu and Huang (2003) observed that citrate presence also inhibited the formation of iron-based
crystals, even after 135 days of precipitation. Rather, the citrate promoted short-range ordered
materials.
2.3 IRON CORROSION
2.3.1 Corrosion process
2.3.1.1 Anaerobic iron corrosion
In water solutions, Fe0 corrosion is expected as the stability field of Fe0 lies below the water redox
line (Pourbaix 1973), indicating that Fe0 should oxidize in the presence of water (Matheson and
Tratnyek 1994). Taking under consideration the different mechanisms proposed in literature over
the past 50 years (Kabanov, et al. 1947; Bockris, et al. 1961; Burke, etal. 1986; McDougall, et al.
1995; Heusler, et al. 1958; Cornell, 1996), the most probable iron surface reactions can be
summarized in the following schema:
Fe0 +OH- → Fe(OH)ads + e- (1) Fe(OH)ads → Fe(OH)ads
+ + e- (2)
The surface complex Fe(OH)ads+ can be regarded as amphoteric, in acidic solutions, the simple
31
dissolution reaction becomes significant (Burke, et al. 1986; Cornell, 1996)
Fe0 +2 H2O = Fe2+ + H2+2OH- (3)
In this case no iron surface product is anticipated.
The iron dissolution reaction can also be expressed as reaction (4)
Fe(OH)ads+ +H+ + nH2O → Fe(H2O)n+1
2+ (4)
at pH > 9, the corrosion of Fe and pH increase leads to iron precipitate and passive layer forms.
Fe(OH)ads+ + OH- → Fe(OH)2 (5)
Oxidation of Fe0 can proceed along several reaction pathways (Schwertmann and Cornell 1991).
Odziemkowski (1998) found that anaerobic corrosion of iron leads to the initial formation of
ferrous hydroxide at the beginning of the reaction. Independent of the ground water composition,
however, the final corrosion product is magnetite. The spontaneous (no current applied) formation
of magnetite takes place by a dissolution/precipitation mechanism with the separation of anodic
and cathodic sites across the surface film. The cathodic reaction, which takes place at the porous
film/solution interface, requires the film to be electron conducting.
3Fe(OH)2(s) = Fe3O4 +H2(g) +2H2O +2e- (6) (transformation into magnetite)
Under anoxic conditions, hydroxide produced by slow iron oxidation reaction (3) will cause the
local pH value to increase. Under subneutral to neutral pH valules, which is typical for most
surface and groundwater, the pH may increase to a final local value of 9.0 or higher, contributing
to the formation of ferrous hydroxide precipitates that coat the metal surface. In a solution with
high concentration of carbonate, as the hydroxide ions are consumed, soluble carbonate ions are
formed as reaction (7) and (8). Buildup of carbonate ions eventually results in the precipitation of
carbonate solid species:
H2CO30 +2OH- = CO3
2- +2H2O(l) (7)
32
HCO3- +OH- = CO3
2- +2H2O(l) (8) Ca2+ + CO3
2- = Ca CO3(s) (9) Fe2+ + CO3
2- = Fe CO3(s) (10)
The carbonate coatings on the reactive Fe0 surface could prevent adsorption and precipitation of
the reducible radionuclides from the contaminated water.
2.3.1.2 Iron corrosion with the presence of oxygen or other oxidizer
Atenas, et al (2005), studied the iron corrosion under aerobic weak acidic conditions, the overall
reaction, which depends on pH, is as the following:
At pH 2-3
Fe0 + 2H+ + 1/2 O2 + 5 H2O = Fe(6H2O)(aq)2+ (11)
At pH 4-5
Fe0 + 3/4 O2 + 1/2 H2O = Fe(OOH)(surf) (12)
We note that equation 11 consumes H+, while equation 12 consumes neither acid or base. Atenas,
et al (2006), claimed that at pH 5, where the hydroxyl group concentration in solution is higher
and at interface of both cathodic and anodic sites, Fe (OH)2 is formed, inhibiting further iron
dissolution. But under aerobic condition, the produced ferrous hydroxide is transformed quickly to
lepidocrocite (γ-FeOOH), which is the major observed surface product of iron corrosion at these
solution conditions. The lepidocrocite is probably formed from ferrous hydroxide in the following
reaction (Mazaudier, et al. 2003):
2Fe(OH)2(s) + 1/2 O2 → 2Fe(OOH) + H2O (13)
Lepidocrocite is unstable, and will convert to goethite with time (Pourbaix, 1975). Goethite
(α-FeOOH) is the main constituent of passive films on iron formed by the
dissolution–precipitation mechanism (Kruger, et al. 1984)
Green rusts ([Fe4(OH)8Cln·H2O] or [Fe6(OH)12][CO3·nH2O])as an intermediate product during
the hydrolytic oxidation of Fe2+ solutions often transform to FeOOH (Olowe and Genin 1991).
33
Green rusts oxidize to Fe(III) species that include goethite (α–FeOOH), lepidocrocite (γ–FeOOH),
hematite (Fe2O3), or magnetite (Fe3O4), depending on the rate of oxidation and dehydration
(Myneni et al. 1997). Wang et al (2001) observed the formation of Green rust type compounds
under oxygen-deficient conditions. They also observed that upon contact with air Green rust are
oxidized into iron oxyhydroxides, lepidocrocite or magnetite.
Oxidizers like nitrate can also cause passivation of metals by initially reacting with the metal,
causing an increase in pH and a positive shift in potential. As a result, high valency oxide species
like Fe2O3 and FeOOH form and remain stable at the metal surface, creating a film which acts as a
barrier to further reaction with the metal (Ritter et al. 2003). Ritter’s research indicated that
magnetite rather than hematite was thermodynamically with contact of nitrate-free water, while
with nitrate water, hematite is the primary corrosion products.
2.3.1.3 Reduction of surface corrosion product on Fe0
It has been proposed that hematite and maghemite are removed from the Fe0 surface by an
autoreduction reaction (Odziemkowski and Gillham 1997) under anoxic conditions, whereby
water and electrons from iron metal participated in the reduction of these surface species. Under
open circuit potential conditions, this process has been proposed to occur as a reductive
dissolution reaction, releasing Fe2+ into solution (Pryors and Evans, 1950):
Fe2O3 + Fe + 6H+ → 3 Fe2+ + 3H2O (14)
with metallic iron being oxidized according to the reaction:
Fe → Fe2+ + 2e- (15)
and supplying the electrons for the reaction:
Fe2O3 + 6H+ → 2 Fe2+ + 3H2O (16)
Ritter et al. (2002) proposed that magnetite may have formed by a dissolution–precipitation
reaction in which the corrosion of iron (by electron transfer through the magnetite layer) by water
34
or TCE released Fe2+ into solution, which then precipitated back onto the surface as Fe (OH)2.
Ferrous hydroxide is thermodynamically unstable, and will convert to magnetite (Reardon, 1995).
At ambient temperatures, this transformation is catalyzed by the presence of iron metal (Shipko
and Douglas, 1956), and occurs according to the Schikorr is proportionation reaction (Schikorr,
1929):
3Fe(OH)2 →Fe3O4 + H2 + 2H2O (17)
Ritter et al. (2000) observed stable existence of green rust in a permeable iron wall
throughout the TCE removal process of at least 25 days. They attributed the formation of green
rust from partial oxidation of Fe(OH)2, which in turn may have either been a product of
autoreduction of hematite or corrosion of Fe0, or both. Green rust is similar to magnetite with
respect to the fact that it will not prevent electron transfer (Simard et al., 1997), and may further
act as a reductant, and hence may be beneficial to the technology.
2.3.2 Corrosion product characterization
Corrosion is the oxidation of metallic iron, and this process releases iron into solution, or
results in iron scales. Corrosion may uniformly attack a metal surface (uniform corrosion) or it may
be focused at specific sites. While uniform corrosion results in the development of uniform scales,
localized corrosion can produce growing amount of corrosion products (Obrecht and Pourbaix,
1967), called tubercles (Larson and King, 1954; Larson and Skold, 1957).
2.3.2.1 Corrosion scales on iron pipes in water distribution systems
Corrosion scales had a layered structure with a porous interior, which was covered by a
dense shell-like layer. A loosely held surface layer was present above the shell-like layer at the
35
water-scale interface. Wet chemical analyses from the pipe wall indicated that the oxidation state
of iron increased with distance from the pipe wall. High Fe(II) concentrations – sometimes even >
90% by weight – were found inside the scales, whereas only Fe(III) was detected at the top layers
of the scales.
Sarin, et al (2001, 2004) studied the iron pipe corrosion scales in water distribution
system, According to their report, the pipe inner surface usually has 2–5mm thick, reddish-brown,
semi-uniform corrosion layer (Sarin et al. 2004). Corrosion scales originate at the pipe wall and
grow radially inwards towards the pipe center, typical iron corrosion products formed on aged pipe
surfaces can be described using three different layers: (a) a macroporous layer (ferrous hydroxide,
Fe(OH)2 and magnetite, Fe3O4) in contact with the metal, (b) a microporous shell like film of
mixture of Fe2+ and Fe3+ species that covers the macrolayer, and (c) a top layer of red rust (mainly
goethite, α-FeOOH and hematite, α-Fe2O3) in contact with water. The exact composition and
structure of iron corrosion scales, however, varies significantly with water qualities as well as flow
properties (Butler, 1966; Tuovinen, 1980). Similar scale structure has also been reported by other
researchers (Herro and Port, 1993, Kölle and Rosch, 1980; Sontheimer et al. 1981).
Depending on the composition of the water next to the pipe surface, corrosion of iron may result in
the formation of very complexed compounds. Benjamin et al. (1996) documented that the
compounds usually found in iron corrosion scales include goethite (α-FeOOH), lepidocrocite
(γ-FeOOH), magnetite (Fe3O4), siderite (FeCO3), ferrous hydroxide (Fe(OH)2), ferric hydroxide
(Fe(OH)3), ferrihydrite (5Fe2O3 ·9H2O), green rusts (e.g. Fe4IIFe2
III(OH)12(CO3))and calcium
carbonate (CaCO3) (Benjamin et al. 1996).
Inside the scale, close to the pipe wall, iron phases in the oxidation state Fe(II) such as
36
siderite and ferrous hydroxide are expected. Ferrous hydroxide is formed under reducing conditions.
However, in the presence of carbonic species, siderite (FeCO3) is the stable ferrous solid (Singer and
Stumm, 1970). Benjamin et al. (1996) and Sontheimer et al. (1981) have also documented the
presence of siderite as a corrosion product. A high concentration of Fe(II) inside the porous interior
of the scale has also been reported (Sarin et al. 2001)
The role of shell-like layer scales has been examined in considerable depth by others
(Kölle and Sontheimer, 1977; Kölle and Rosch, 1980; Sontheimer et al., 1981). Goethite and
magnetite have been reported as the phases comprising the dark, relatively hard and brittle layer
(Benjamin et al., 1996). Sarin, et al, 2004 studied the powder samples from top layers (surface and
shell) of scales, and magnetite, goethite and lepidocrocite were identified. Considering the top
layers has come into contact with waters having high DO and chlorine concentrations. The
presence of lepidocrocite may be formed by rapid oxidation of ferrous hydroxide, as proposed by
Cornell and Schwertmann (1996). Synthesis studies have shown that goethite formation
predominated over lepidocrocite formation at a slow oxidation rate (Benjamin et al., 1996). A
high pH and a slower rate of oxidation favored formation of magnetite over lepidocrocite (Cornell
and Schwertmann, 1996).
Sarin, 2001, observed of higher Fe(II) concentrations in wet scales compared to dry scales
indicated that Fe(II) phases were converted to Fe(III) phases during drying. The dry samples had
ferrous ions present only in the form of magnetite (Fe3O4), and the conversion of a phase such as
Fe(OH)2, to magnetite by oxygen account for the observed change.
2.3.2.2 Corrosion layers on the surface of iron used in contaminant removal
A permeable reactive barrier (PRB) is defined as an in situ method for remediating
37
contaminated ground water that combines a passive chemical or biological treatment zone with
subsurface fluid flow management. Treatment media may include zero-valent iron, chelators,
sorbents, and microbes to address a wide variety of ground-water contaminants, such as
chlorinated solvents, other organics, metals, inorganics, and radionuclides. The contaminants are
concentrated and either degraded or retained in the barrier material, which may need to be
replaced periodically.
The Master Builder iron used in field-scale PRBs usually is a low-grade cast iron (carbon
content >2%) that contains a number of chemical elements in addition to iron. These “impurities”
include graphite inclusions (Burris et al. 1995; Deng et al. 1997), alloying metals (e.g., manganese,
copper, chromium), metalloids (e.g., silicon), and nonmetals (e.g., phosphorus, sulfur) (Landis, et
al. 2001).
Kohn et al. (2005) studied the un-reacted granular iron that would be used in TCE
treatment. The unreacted iron grains exhibit a thin scale (<1000 nm), mainly consisted of a single
dense phase that completely covers the grain. The compositional map shows that the oxygen
content in the scale is higher than that in the iron core, indicating the presence of oxides. After
used in the PRBs for TCE removal. The oxide layer on the reacted grains is generally very thick,
sometimes ranging to over 100 μm, as compared to less than 2μm on the raw grains. The pores
of the reacted grains generally contained substantial amounts of precipitates (Kohn et al. 2005).
The precipitates present on the exposed grains (for TCE removal) exhibits a layered morphology.
In intimate contact with the metal is a very thin layer of nanocrystalline magnetite or maghemite.
Adjacent to this nanocrystal layer, there is a gap filled with carbonaceous material. This material
is probably part of the alteration structure and represents carbon associated with the metal after
38
some of the iron has dissolved. The next layer contains a mixture of fayalite (Fe2SiO4) and
magnetite/hematite, with fayalite crystals (200nm) elongated roughly parallel to the interface.
SEM study of the surface morphology of ZVI (Gu et al. 1999; Phillips et al. 2000) showed
two authigenic precipitate morphology: acicular aggregates and cryptocrystalline clusters. The
acicular aggregates were largely composed of green rust minerals, goethite, lepidocrocite, and
calcium carbonate phases, whereas cryptocrystalline clusters contained mackinawite and poorly
crystallized iron (oxy-) hydroxides. XRD(X-ray Diffraction ) data indicate that after contact with
untreated groundwater, the iron corrosion products contains magnetite, carbonate green rust,
mackinawite, lepidocrocite, calcite and aragonite.
Furukawa and co-workers (2002) investigated the precipitates found in two peerless
iron-based PRBs used to treat chlorinated solvents and hexavalent chromium. The most common
precipitates were claimed to be ferrihydrite (Fe5HO8·4H2O) and another phase they identified is
magnetite. Odziemkowski and coworkers (1995), using in-situ Raman spectroscopy, determined
that magnetite and possibly ferrous hydroxide (Fe(OH)2) were the primary corrosion products,
independent of solution composition. Similarly, Gui and Devine (1995), using surface-enhanced
Raman spectroscopy (SERS), found that the precipitate film consisted of Fe(OH)2 and a mixture
of Fe3O4 and γ-Fe2O3, regardless of the anions present in solution. A number of more highly
oxidized precipitates have been encountered in column studies or PRBs exposed to solutions
containing oxyanions such as Cr(VI) or U(VI), as well as in laboratory and field studies with
aerobic groundwater . (Mackenzie et al. 1999). These precipitates include goethite (α-FeOOH),
akaganeite (β-FeOOH), lepidocrocite (γ-FeOOH), hematite (α-Fe2O3), and amorphous iron
(hydr)oxides (Gu et al. 1999). Commonly reported carbonate species include calcite (CaCO3),
39
aragonite (CaCO3), carbonate green rust, and siderite (FeCO3) (Vogan et al. 1999; Gu et al. 1999;
Roh et al. 2002; Philip et al. 2000; Fukukawa, et al. 2002). As noted, a wide variety of surface
precipitates have been detected in granular iron systems, but little is known about the temporal and
spatial variability of these phases.
For the ZVI that employed in columns, more highly oxidized phases (e.g., goethite,
hematite) were typically found near the column inlet, and more reduced phases (e.g., green rust)
were detected near the outlet. Kohn et al.(2005) studied the corrosion products in different places
in the Fe0 PBR by Raman spectroscopy method. For grains extracted near the inlet, magnetite,
maghemite, and surface associated carbonate are the primary phases. For iron grains extracted
from ports ~30 and ~80 mm from the inlet, aragonite (CaCO3) also present. As discussed below,
samples from the same column subjected to XRD analysis indicate that calcite was the
predominant carbonate mineral phase in this column after 475 days. Near the column outlet,
carbonate precipitate disappered and green rust compounds became the major phase (Bonin et al.
2000; Legrand et al. 2001).
2.3.3 Surface corrosion products and contaminant removal
The effect of iron corrosion on contaminant removal is debatable. The surface oxidized layer
could act as a mediator in electron transfer between metallic iron and solution and play a crucial
role in the decomposition of water pollutants. The importance of the iron surface layer was already
recognized in recent reports (Balko and Tratnyek 1998; Oh et al. 2002; Weber 1996).
Water-pollutant decomposition efficiency, reaction pathway, and decomposition products are
strongly related to the properties of the mediator surface layer (Mielczarski et al. 2005). Previous
40
laboratory and field studies have shown the development of surface corrosion and authigenic
precipitates (Mackenzie et al. 1999; Puls et al. 1999; Roh et al. 2000). Such surface precipitates
may mask the redox active sites where exchange of electrons between Fe0 and contaminants is
facilitated. The corrosion products may also reduce the barrier permeability by occupying
available pore space. On the other hand, the formation of iron oxyhydroxides with large surface
areas may be beneficial for the immobilization of certain contaminants (e.g., arsenic) through
sorption or coprecipitation (Lackovic et al. 2000; Melitas et al. 2002).
2.3.3.1 Iron corrosion and contaminant reduction in PRBs
A permeable reactive barrier consists of a zone of reactive material, such as granular iron or
other reduced metal, lime, electron donor-releasing compounds, or electron acceptor-releasing
compounds, installed in the path of a plume of contaminated groundwater. As the groundwater
flows through this zone, contaminants are degraded and transformed to innocuous components,
adsorbed, or chemically altered so that they form insoluble precipitates and leave solution. The
most commonly applied permeable reactive barrier consists of granular zero valent iron used to
remediate dissolved chlorinated solvents (USEPA, 1999).
The material that is used in iron walls is not pure iron, but rather is a commercial product,
consisting of scrap metal, mostly cast iron and low alloy steels, and the material is furthermore
covered with a passive oxide film (Ritter et al. 2002). Roh, et al, (2000) studied the master builder
Fe0 filings used for dechlorination with SEM micrograph, results also proved that the Fe was
covered with Fe (hydr)oxides.
Several potential precipitates, such as magnetite (Fe3O4) (Gregory et al. 2004; Lee et al.
2002; Sivavec and Horney 1996) and green rust (e.g., [Fe42+Fe2
3+(OH)12][CO3·2H2O]) (Erbs et
41
al. 1999), as well as Fe2+ adsorbed onto iron (hydr)oxides (Elsner et al. 2004; Klausen et al. 1995;
Pecher et al. 2002), have been shown to be able to reduce organic redox-active contaminants.
Legrand et al. (2004), studied the reduction of chromate by Fe(II)/Fe(III) Carbonate Green Rust.
Results indicate the formation of ferric oxyhydroxy carbonate and the concomitant precipitation of
CrIII monolayers at the surface of the iron compound that induce passivation effects and
progressive rate limitations. Thick green rust particles formed by the corrosion of iron in
permeable reactive barriers, makes FeII not accessible for efficient CrVI removal.
Stratmann et al. (1994) studied the mechanism of the oxygen reduction on rust-covered
metal, the results show that oxygen is predominantly reduced within the rust scale and not at the
metal/electrolyte phase boundary. In order to allow any oxygen reduction, the rust layers have to
be reduced. Oxidized rust scales, which are nearly free of Fe2+ states, inhibit the reduction of
oxygen completely.
The degradation characteristics changes with time in the PBR, these changes are attributed
to (1) reduction in Fe surface reactivity caused by passivation of Fe0 by precipitates, including Fe
(hydr)oxides and Fe sulfides, and (2) alternation of flow paths through Fe filings as a result of
precipitation and cementation.
Bacteria may cause a potential negative consequence of biofouling because the proliferation
of bacteria in an improperly designed reactive barrier could reduce the hydraulic conductivity of
the barrier, thereby hindering the flow of groundwater through it (Weathers and others 1997).
2.3.3.2 Iron corrosion and contaminant adsorption in PBRs
Furukawa, 2002 studied the fine-grained fractions of permeable reactive barrier (PRB)
42
samples for groundwater treatment. They claimed that if adsorption is mechanism for contaminant
removal, Fe0-PRBs may remain effective for a longer period of time in slightly oxidized
groundwater systems where ferrihydrite formation occurs compared to oxygen-depleted systems
where magnetite passivation occurs.
2.4 THE MECHANISMS OF ARSENIC REMOVAL BY IRON BASED
SORBENTS
2.4.1 Adsorption of Arsenic by iron oxide/hydroxide—As removal mechanisms
The mechanisms of As sorption to the iron oxide/hydroxide surfaces based on the
spectroscopic, sorption, and EM measurements are as follows: arsenate forms inner-sphere surface
complexes on Fe oxide, while arsenite forms both inner- and outer-sphere surface complexes on
amorphous Fe oxide (Goldberg and Johnston 2001). Adsorption on ferrihydrite occurs by ligand
exchange of the As species for OH2 and OH− in the coordination spheres of surface structural Fe
atoms (Jain et al. 1999). While arsenate adsorption resulted in the net release of OH− at pH 4.6 and
9.2, arsenite adsorption resulted in net OH− release at pH 9.2 and net H+ release at pH 4.6. The
amount of H+ or OH− released/adsorbed As (mol/L) varied with the As surface coverage,
indicating that different mechanisms of arsenic adsorption predominate at low versus high
coverage. The results provide evidence that during arsenite adsorption at low pH, i.e., pH 4.6, the
oxygen of the Fe–O–As bond remained partially protonated as Fe–O(H)–As (Jain et al. 1999).
For Fe3O4, α-FeOOH, γ -Fe2O3, and amorphous Fe(OH)3, values of pHzpc = 6.5–8.5 were
obtained (Stumm, 1981). Hlavay et al.(2005) studied the surface properties of iron
hydroxide-coated alumina adsorbents for arsenic removal, results revealed that: the total capacity
43
of the adsorbent was 0.12 mmol/g, and the pH of zero point of charge, pHzpc = 6.9 ± 0.3.
Depending on the pH of solutions, the adsorbent can be used for binding of both anions and
cations, if pHeq < pHzpc anions are sorbed on the surface of adsorbent (S) through c and {S–OH}
groups. Values of pHiep = 6.1 ± 0.3 for As(III) and pHiep = 8.0 ± 0.3 for As(V) ions were found.
The amount of surface charged groups (Q) was about zero within the a pH range of 6.5–8.6, due to
the practically neutral surface formed on the adsorption of As(V) ions. At acidic pH (pH 4.7), Q =
0.19 mol/kg was obtained.
2.4.2 Arsenic removal by ZVI
The use of Fe0 to remove arsenic has been actively investigated by many groups t al. (Farrell
et al. 2001; Krishna et al. 2001; Manning et al. 2002; Su and Puls 2001a; Su and Puls 2003). In
this method, arsenic is adsorbed onto corrosion products of zero-valent iron (ZVI) as the ZVI
converts so such species as iron (oxyhydr) oxide. Possible arsenic removal processes in
zero-valent iron system include surface adsorption onto corrosion products, e.g. iron
(oxyhydr)oxides (Manning et al. 2002, Dixit and Hering 2003), precipitation such as formation of
symplesite (Fe3(AsO4)2· 8H2O) (Nikolaidis et al 2003) , co-precipiration ( e.g. arsenic
co-precipitation with carbonate green rust) (Lien and Wilkin 2005) or redox reaction such as As
(III) oxidized to As(V) by corrosion products or impurities such as MnO2 (Melitas et al. 2002,
Manning et al. 2002).
2.4.2.1 Iron corrosion and arsenic removal on ZVI – the process
In anoxic environment, upon contact with water, the corrosion of ZVI may happen as an
autoreduction process, as discussed in Section 2.3.1.3. Continual corrosion of ZVI to generate iron
44
oxides is needed for the continuous removal of As by ZVI. It is expected that, once the free iron
metal is depleted or complete passivation occurs, As removal capacity will decrease and
eventually cease.
With respect to pH, an optimum range for As(III) adsorption by ZVI is expected because: (1)
acidic conditions favor ZVI corrosion; and (2) maximum adsorption of As(III) on iron oxides
occurs between pH 7 and 9.2. ZVI corrosion results in the release of Fe2+ and OH− into solution,
which in turn forms Fe(OH)2 initially and ferric oxides with time. The optimum pH range for
removal of As(III) was found to be between 7 and 8 (Yu et al. 2006).
Aging maybe beneficial for arsenic removal. It was observed that after aging ZVI for two
months, significant improvement was observed in the percentage removal of As(III) at pH 9 (Yu
et al. 2006).
Carbonate effect on arsenite removal by ZVI debatable. Carbonates are known to stimulate
iron corrosion (Evans 1982) however, they may also interfere with As(III) adsorption onto iron
oxides. A recent assertion by Kim et al. (2000) was that complexation between carbonate and As
was responsible for the observed correlation between soluble As with carbonate concentration in a
Michigan groundwater. Yu et al. (2006) found that in typical groundwater conditions, when
alkalinity is below 200 mg/L as CaCO3, competition of HCO3−/CO3
2− with As(III) for adsorption
sites on iron oxides will most likely be negligible.
2.4.2.2 Rate controlling arsenic removal by ZVI
Mass transfer efficiency was found to play an important role in the removal of arsenic by
ZVI. After an initial period of arsenic rapid adsorption to surface rusts formed during
manufacturing and exposure to air, arsenic removal rate is most likely controlled by the rate of
45
iron corrosion and the diffusion of arsenic to adsorption sites in ZVI/iron oxides (Yu, 2006). In a
batch study of As(V) adsorption to ferrihydrite, Fuller et al. (1993) reported that, following the
fast saturation of available surface sites, diffusion of As(V)to inner adsorption sites was the
rate-limiting step.
Liu’s research (2006) has proposed differences in iron aging effect on TCE removal by Fe0
in column and batch results. As reported, long term column tests, decline in dechlorination. The
decline was attributed to an increase in the mass transfer resistance of contaminants due to
insoluble Fe-oxides and Fe-(oxy)- hydroxides formed on particle surface, or to porosity loss and
decreased access to iron particles in the column. In contrast, long-term batch studies on the
corrosion behavior of micrometer-scale iron filings in unbuffered water reported a constant
(zero-order) H2 corrosion rate over a 125-160 day period suggesting that the iron corrosion rate,
hence reactivity, is not changing as the iron ages (Reardon 1995; Reardon 2005).
2.4.3 Redox reaction in ZVI system
The issue of arsenic redox reactions in iron filter media has not been resolved. Several
investigations using column reactors packed with iron filings have reported that the relative
concentrations of As(V) to As(III) in the effluent solutions were the same as those in the feed
solutions (Lackovic et al. 2000; Melitas et al. 2002). Spectroscopic analyses of iron filings from
column reactors treating As(V) have found no discernible As(III) or As(0) associated with the iron
particles, even after more than 1 year of operation (Farrell et al. 2001). These observations suggest
that there is no reduction of As(V) in iron media filters. Although column studies have not
observed changes in the arsenic oxidation state, reduction of As(V) to As(III) and As(III) to As(0)
46
have been observed in batch experiments conducted in nitrogen purged solutions containing iron
filings (Bang et al. 2005). Su and Puls reported that the ratio of As(V) to As(III) on iron filings
after 60 days elapsed was approximately 1:3. This ratio was independent of whether As(V) or
As(III) was the initial reactant, which strongly suggests that the 1:3 ratio is representative of
equilibrium between As(V) and As(III) on the iron surfaces.
Melitas’ study (2002) showed that bound arsenic species decrease the corrosion rate of
zerovalent iron and that bound or solution-phase As(V) may be reduced to As(III). Reduction of
bound As(V) occurs at higher potentials than reduction of aqueous arsenate. At lower potential
that favors the arsenate reduction, the electrochemical adsorption of arsenate was retarted because
of the negatively charged iron surface. Thus As(III) adsorption was favored over As(V) at the iron
surface. The stronger binding of As(III) results in an elevated As(III) to As(V) ratio on the iron
surface with respect to their bulk solution ratio. The elevated As(III) concentrations on the iron
surface decrease the equilibrium potential for further As(V) reduction. Melitas et al. (2002)
concluded that the pH and potential conditions necessary for significant As(V) reduction will be
difficult or impossible to achieve in an open system under freely corroding conditions. Therefore,
in the absence of biological reduction, there will be little conversion of As(V) to As(III) in zero
valent iron filter media.
2.4.4 Arsenic release
The principal mechanisms of arsenic mobilization associated with geochemical conditions
have been identified as desorption in alkaline conditions, competitive sorption, and reductive
release, especially as associated with the dissolution of iron oxides. Of these, the reductive release
47
of arsenic and/or arsenic-bearing minerals especially iron(III) (hydr)-oxides, appears to be the
primary cause of elevated arsenic levels under most conditions. (Cummings et al. 1999; Nickson
et al. 2000; Pfeifer et al. 2004).
In drinking water distribution systems, arsenic released could be related to iron based solids. It
was reported that solids released from cast iron pipes could have an arsenic content of 83 ug As/g
solid, while hydrant flushed solid contain nearly 2000 ug As/g solid (Lytle et al. 2004). Those iron
oxide solids are loosely deposited at the pipe surface and can become re-suspended by hydraulic
flow.
The dissolution and transformation of the iron (hydr)oxides will impart a pronounced effect on As
partitioning. Ferrihydrite, a short-range order material common in soils and sediments, is
transforming to lower surface area minerals such as goethite and magnetite in the presence of
aqueous Fe(II) (Benner et al. 2002; Hansel et al. 2003). Thus, iron reduction should be expected to
induce As release (desorption) from Fe(III) (hydr)oxides dissolved or are transformed to lower
surface area minerals.
As(III) binds to Fe(III) (hydr)oxides more extensively than As(V) under circumneutral conditions
(Dixit and Hering 2003), but was contrarily shown to be more mobile under flow conditions than
As(V) (Gulens, 1979; Jenne, 1979). Thus, the reduction of As(V) to As(III) will also cause arsenic
release.
Arsenic associated with poorly crystalline iron oxides can also be mobilized as a result of
dissimilatory iron reduction by microorganisms (Cummings et al. 1999; Nickson et al. 2000;
Pfeifer et al. 2004; Van Geen et al. 2004; Zobrist et al. 2000)
48
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55
Figure 1.1 Molecular configurations of arsenite and arsenate
pH2 4 6 8 10 12
Spec
iatio
n (%
)
0
20
40
60
80
100
120
AsO43-
HAsO42-
H2AsO4-
H3AsO4
A
pH2 4 6 8 10 12
Spec
iatio
n (%
)
0
20
40
60
80
100
120
AsO33-
HAsO32-
H2AsO3-
H3AsO3
B
Figure 1.2 (A) arsenate and (B) arsenite speciation as a function of pH
56
CHAPTER 3
Arsenic Removal from Groundwater by Iron tailored GAC
3.1 INTRODUCTION
In recent years, arsenic contamination of groundwater has emerged as a major concern on a global
scale (AwwaRF). At 50 ppb arsenic level, 13 people out of 1000 could die from cancer to the liver,
lung, kidney, or bladder (Smith et al. 1992). This situation is especially serious in Taiwan, India,
Chile and Bangladesh. For this concern, the WHO in 1993 and USEPA in 2001, lowered the
arsenic standard from 50 ppb to 10 ppb. In USA, all public water systems must comply with the
new standard since January 1st 2006. The USEPA and AwwaRF had estimated the costs to meet
this new MCL to be $102 to 550 millions per year. So there is urgent need for simple and cost
effective technologies for arsenic removal.
Many researches had been focused on developing new arsenic adsorbents because adsorption
systems have the characters motioned above. Studies have shown that iron oxides, such as
granular ferric hydroxide (GFH) (Driehaus et al. 1998; Driehaus et al. 1995) and hydrous ferric
oxide (HFO) (Dixit and Hering 2003) can be effective to remove both As (V) and As (III) from
aqueous solutions. GFH is reported to have a high treatment capacity of 30 000-40 000 bed
volumes to a 10 μg/L breakthrough (Driehaus et al. 1998). Dixit and Hering reported maximum
sorption of about 0.2 moles As per mole of iron in HFO (Dixit and Hering 2003). But these iron
oxide granules can crumble and disintegrate when they experience prolonged use, it was reported
that GFH with a media particle size at 0.8-2.0 mm will need a backwashing every 5000 Bed
57
volumes (Selvin et al. 2000). Also, after backwashing, there would be significant amount of
headloss pressure built up in the system (Gu and Deng 2005).
Studies revealed that iron (III) had high affinity toward inorganic arsenic species and very
selective in the sorption process. Recent researches are focused on creating cheap and stable iron
bearing adsorbents. There are: (1) iron oxide coated sand. Sand only serve as a support, it's the
iron oxide who removes arsenic. Gupta et al. (2005) found out the adsorption of As (III) onto iron
oxide coated sand could reach 28.57 ug/g at pH 7.5. (2) Iron oxide impregnated activated carbon.
The adsorbent could get an iron content of 7%. At pH 7, 1mg/L arsenic, 0.2 g/L Fe oxide
impregnated activated carbon, this adsorbent could get an As (III) adsorption of 4.67 mg/g, an As
(V) adsorption of 4.5 mg/g (Vaughan and Reed 2005). (3) Fe (III)-loaded cellulose sponge. The
sponge is claimed to contain free available ethyleneamine and iminodiacetate groups, which
interact with Fe (III) by chelation and ion exchange. The Fe (III) loading capacity was 0.25 mmol
Fe/g sponge, corresponding to 1.4% Fe content. And with this 1.4% iron loading, the media has a
high As (V) adsorption capacity as 1.83 mmol As /g, and a fairly As (III) adsorption capacity as
0.24 mmol As/g (Munoz et al. 2002). (4) Granular activated carbon (GAC) based iron containing
adsorbents. This media was made in 2 steps, first Fe (II) was adsorbed onto GAC, then the Fe (II)
was oxidized to Fe (III) by O2, H2O2 or NaClO. When lignite based carbon was employed, the iron
loading could reach 7.89%, and the author proposed that the impregnated iron was mostly in
coordinated form with various functional groups on GAC, but not in polymeric iron hydroxide
form. The adsorbent could remove arsenic to 7500 Bed volume before reached 10ppb
breakthrough (Gu et al. 2005).
Among all these medias, GAC is of the most concern. GAC has large surface area, high pore
58
volume, and rigid structure, which renders it an ideal backbone for hosting a considerable quantity
of iron and also GAC had been used for water treatment for decades, there are no investment for
new systems needed.
Surface modification of activated carbon by immobilizing organic compounds is recognized as an
effective approach for enhancement of heavy metal removal. Surface modification of GAC by
citric acid had been reported to enhance copper adsorption by 140% (Chen et al. 2003). The
presence of ethylenediamine tetra-acetic acid (EDTA) was reported to improve the cadmium
adsorption by 2~3 times (Patrick, et al. 1995). Besides, iron, citric acid, EDTA and fatty acids are
non-toxic and commonplace in water and foods. No primary drinking water standards exist for
any of these species.
Pore Volume Impregnation, also called incipient wetness, is one of the most prevalent
methods used in the area of catalyst production.
The impregnation procedure comprise saturating the pores of the porous support with
aqueous metal salt solution, drying and then calcining the impregnated support to convert the
metal salt to metal oxide. It had been reported that 10-12 % Cu impregnation could be achieved
via this method (Montanari et al. 1997; Marchi et al, 2003).
Unfortunately, this impregnation procedure affords relatively large metal crystallites
concentrated at the surface of the support particle (Delmon 1979), which means most metals are so
buried inside so that they are not active.
Amorphous oxide had been loaded on sand (Benjamin et al. 1996), incinerator melted slag (Zhang
and Itoh 2005), and also in a macroporous cation exchange bead (DeMarco, et al. 2003) and the
latter two had been tested for arsenic removal. The porous cation exchange resins could get an
59
iron loading of 25% based on dry bead weight, and 50% out of the iron loading is in the form of
FeOOH. This product could remove arsenic from 50 ppb down to 10 ppb for 45,000 BVs.
DeMarco, et al. also pointed out that the HFO agglomerates in the beads were very stable,
turbulence and mechanical stirring did not result in any loss of HFO. The iron oxide loaded on
metled slag was believed to chemically bonded with the Si inside slag, which prevented the
crystallization of FeOOH, and they claim this material had better arsenic removal (arsenate of 78.5
mg/g) capacity than FeOOH.
This paper will focus on developing iron bearing GAC for arsenic removal, the authors had
studied two iron loading methods, incipient wetness method and preloading GAC with organic
carboxylic-iron. Batch test and column test was conducted to test the arsenic removal capacity of
Fe tailored GAC.
3.2 MATERIALS AND METHODS
3.2.1 Materials.
All chemicals were reagent grade. The experiments had employed 0.01M EDTA solutions
from VWR scientific products; Palmitic acid ( CH3(CH2)14CO2H) from ALDRICH; and Citric
acid (HOC(COOH)(CH2COOH)2) from J.T.Baker company. Metal ions employed include Ferric
chloride (FeCl3·6H2O), Magnesium chloride (MgCl2·6H2O) and Manganese chloride (MnCl2 •
4H2O) from Fisher Scientific company. As (V) solution was made from Na2HAsO4·7H2O (Alfa
Aesar).
Activated carbon employed here is Ultracarb from USFilter # WESTATE. The non
tailored carbon was designated as virgin carbon.
60
3.2.2 Organic carboxylic-Fe preloaded carbon.
Several kinds of organic acids were tested, including pamaltic acid, EDTA, citric acid and
L-Glutamic acid. Certain amount of Ultracarb was mixed with the organic carboxyl-Fe solution at
certain concentration, and then agitated on a shaking table at 100-120 RPM for 2~3 days, from
literature, this time is long enough to reach equilibrium. The adsorbent was filtered out and
washed with distilled water until no color in the washing water could be discerned. The tailored
carbon was dried at 104ºC overnight and stored in desiccators before use. Detailed information is
presented in table 2.
The authors also employed evaporation method. In this process, the GAC was added to
50 ml 0.05M Citrate-Fe solution, and then mixed on a hot plate with magnetic stirring until the
solution volume lowered to 5~10 ml.
3.2.3 Preparation of Fe-GAC through incipient wetness impregnation (IWI).
Iron nitrate nonahydrate [Fe(NO3)3·9H2O], was incorporated as a precursor of iron oxide
into the pores of granular-size porous GAC. Some times, citrate was also added as a precursor. The
following coating procedure has been developed to achieve this impregnation as homogeneously
as possible: (1) dissolve the iron precursor in deionized water at given concentrations to have the
final volume of an iron-dissolved solution of 1–1.5 mL, (2) disperse well the iron precursor
solution using a 1-mL micropipette over the dried GAC (1 g), (3) dry the solids at room
temperature for one day, and (4) put in a rotary evaporator for oxidative precipitation of iron
nitrate at a temperature selected from the range of 60–90 ºC for a time of 4–12 h.
61
3.2.4 Adsorption Isotherm.
In this experiment, a prescribed amount of activated carbon ( 10 mg, 20 mg, 50 mg, 80 mg
or 100 mg) was added to 50 mL arsenic-spiked Rutland groundwater (Total arsenic concentration
is 550 ppb). The water pH had been adjusted to 6 with 0.1 M HCl. The mixtures were then put
on the horizontal shaking table and shaken at 120-150 rpm for 48 hours.
3.2.5 Column tests.
Rapid small-scale column tests (RSSCT’s) were conducted to evaluate GAC’s arsenic
adsorption capacity; it was designed to simulate the adsorption conditions that would occur in a
full-scale bed. The RSSCT’s in this paper was designed to simulate a full scale Column with
EBCT of 20 minutes. Detailed configuration of the columns had been discussed in previous work
of our Penn State team (Chen et al. 2003).
All small-scale column tests were carried out at room temperatures 20-23°C. The ground
water originated from the well of the Cool Sandy Beach Community Water System of Rutland,
MA. The total Arsenic in this groundwater was 47-55 ppb depending on weather. Characteristics
of the groundwater were presented in table 3.1.
3.2.6 Chemical Analysis.
To test the iron loading on tailored GAC, a portion of the fully loaded GAC were ashed in
a muffle furnace at 600°C for 24 hours. The ashed GAC was dissolved in 25 mL of concentrated
HCl. After a minimum contact time of 24 hours, the solution was filtered and the filtrate was then
diluted to 250 mL. Solutions were analyzed for iron by the ICP-MS method.
62
Table 3.1 Water quality characteristics of Cool Sandy Beach Groundwater (Rutland, MA)
Cations Concentration Units Anions Concentration UnitsCalcium
(Ca) 59
mg/l CaCO3
Bicarb (HCO3)
64.2 mg/l
CaCO3 Magnesium
(Mg) 11.3
mg/l CaCO3
Fluoride (F)
0.670 mg/l
CaCO3
Sodium (Na) 27.5 mg/l
CaCO3 Chloride
(Cl) 9.32
mg/l CaCO3
Potassium (K)
4.2 mg/l
CaCO3 Nitrate
(NO3) 0.041
mg/l CaCO3
Iron (Fe) < 0.003 mg/lPhosphate
(PO4) < 0.080
mg/l CaCO3
Manganese (Mn)
0.003 mg/l Sulfate
(SO4) 26.4
mg/l CaCO3
Aluminum (Al)
<0.006 mg/lSilica
(SiO2) 12.5
mg/l CaCO3
Zinc (Zn) 0.004 mg/l Arsenic* 47-55 μg/lOther
parameters Units
Other parameters
Units
pH 7.4-7.6 Total
Hardness 70.30
mg/l CaCO3
Turbidity 0.08 NTU TOC (C) 0.851 mg/l
Conductivity 165 μS Free
(CO2) 3.6
mg/l CaCO3
* 25% of arsenic was As(III) and 75% of arsenic was As(V).
Arsenic concentrations were determined with ICP-MS method and also a
hydride-vapor-generation flame atomic absorption spectrometry (HVG-AAS, Shimadzu®) method,
as per Chen et al. (2007).
3.3 RESULTS AND DISSCUSIONS
3.3.1 Organic acid-Fe loading onto GAC
Palmatic acid, citric acid, EDTA and L-Glutamic acid were tested for chelating iron onto
Ultracarb. The results were illustrated in table 3.2. Since EDTA and citrate acid has a high
63
complexation capability, they has higher iron loading result compared to other organic acid, like
long chain fatty acid-pamaltic acid, and L-Glutamic acid. Through evaporation, the iron loading
amount could be increased 1.5 to 2 times.
Table 3.2 Fe loading result a.
Organic acid Fe loading amount
(%)* Pamaltic acid 0.2-0.4
Citric acid 0.97-2.1 Citrate acid with
evaporation method 4.4-5
EDTA 0.7-3.7 EDTA with
evaporation method 5
L-Glutamic acid 2 L-Glutamic acid with
evaporation method 3
* Virgin Ultracarb has a Fe content of 0.15%; the data in this column = virgin Ultracarb Fe content + Fe loaded on Ultracarb.
The iron loading by organic acid chelating could be affected by the initial agent concentration and
pH, as shown in Table 3.3. Highest iron loading was achieved with 0.2 M initial preloading
solution concentration. Higher concentration doesn’t help to load more Fe onto GAC, on the
contrary, the Fe loading amount decreased. When pH increased from 1.7 to 4.2, the iron loading
amount also increased from 1.3% to 1.8%. Other metals like Magnesium, Manganese and copper,
seems have no significant effect on the iron loading.
3.3.2 Iron loading via incipient wetness method
Through incipient wetness method, higher iron loading were achieved, as indicated by
Table 3.4. Under higher temperature, the citric acid helps to increase the iron loading a little bit.
But this trend was not observed at low temperature.
64
Table 3.3 Iron loading result b.
reagent Concentration(mol/L) pH Fe
loading(%) 0.001 2.5 1.05 0.01 1.7 1.3 0.01 4.2 1.8 0.2 1 2.1 0.5 1.6 1 0.3 1.42
Citrate-Fe
2 1.35
Citrate-Fe-Mg 0.0005 ~ 0.01 2.77 0.97 ~
2.4 Citrate-Fe-Mn 0.001 3.34 1.36 Citrate-Fe-Cu 0.05 1.54
Table 3.4 Iron loading via incipient wetness method.
Temperature solution/carbon
mass (ml/g)
Fe loading amount (%)
Fe-Ul (10.1) 10.1
CA(2%)-Fe-Ul (9.5)
9.5 Heated at
60ºC,
CA (10%)-Fe-Ul
(9.9) 9.9
Fe-Ul (3.8) 3.8 Fe-Sd (5.4) 5.4 Heated at
80-90ºC
CA (2%)-Fe-Sd (6.1)
6.1
3.3.3 Batch test of the citric acid-iron preloaded activated carbon
To test the arsenic adsorption capacity with respect to the carbon properties and water pH, we
did a series of batch test. In these tests, 10 mg Ultracarb preloaded via different approaches was
added to 50ml Rutland groundwater and agitated for 48 hours. The pH of the water was adjusted
to 4, 5 and 7. The arsenic adsorption result with respect to carbon properties were shown in table
3.5.
Virgin Ultracarb has a pHpzc value of 10.42, with the citrate-Fe loaded on it’s surface, the
65
carbon’s pHpzc value decreased to 4~6. Second metal may have different effect on the carbon’s
pHpzc value. Cu, Mg decrease the citrate-Fe loaded Ultracarb to lower than 5. While addition of
Mn doesn’t seem to affect the pHpzc value at all.
Generally, the arsenic adsorption increases with iron loading amount. It’s also affected by
water pH and pHpzc of the carbon. For a specific surface, when pH < pHzpc, the surface tends to
be positively charged and will attract anions such as HAsO42- and H2AsO4
-; on the other side,
when pH > pHzpc, the surface tends to be negatively charged and will repel anions. The pKa
values of arsenate are 2.2, 6.97, 11.53. at pH value 4 and 5, the arsenate mainly exist as H2AsO4-,
while at pH 7, it may exist as HAsO42- and H2AsO4
-. As a result, when pH increases from 4 to 7,
the arsenic adsorption dropped by 23%~50%. Best arsenic adsorption was achieved at pH slightly
lower than the carbon’s pHpzc value. For example, CA-Fe-Mg (2.18) has a pHpzc value of 4.72, it
maximum arsenic adsorption was achieved at pH 4.
3.3.4 Isotherm results
We conducted isotherm tests with two organic acid-Fe preloaded GAC, CA-Fe (1.2) and
CA-Fe-Mg (2.18). The results are illustrated in Figure 3.1. The results fit with Langmuir Isotherm
better. The qmax value from the Langmuir Isotherm is 7.5 mg/g for CA-Fe (1.2), and 5.7 mg/g for
CA-Fe-Mg (2.18). Although CA-Fe-Mg has higher Fe loading, it has lower arsenic removal
capacity compared to CA-Fe at pH 6. This could probably be attributed to the carbon’s low pHpzc
value. We also conducted kinetics tests with these two carbons. Results (Figure 3.2) also proved
that the presence of Magnesium interfered the binding of Fe toward As.
66
Table 3.5. Arsenic adsorption capacity with respect to water pH and carbon properties
carbon Fe conc.
(%) pHpzc pH Ce As/g carbon
4 40.4 0.037
5 40.7 0.0355 Virgin 0.16 10.42
7 50.2 ~ 0
4 2 0.229 5 3.9 0.2195
CA-Fe-Mg (2.18)
2.18 4.72
7 19.3 0.1425 4 6.95 0.20475
5 6.6 0.2395 CA-Fe-Mn
(1.36) 1.36 5.8
7 27.4 0.102
4 3.1 0.2235 5 7 0.204
CA-Fe-Cu (1.54)
1.54 4.59
7 24.5 0.1165 4 3.9 0.2195
5 2.6 0.226 CA-Fe (2.1) 2.1 5.6
7 14.4 0.167
4 1.6 0.231 5 4.9 0.2145
CA-Fe (1.78)
1.78 5.84
7 20.2 0.138
As mentioned in the iron loading part, incipient wetness method could produce iron loading
GAC with a high iron loading amount as 10%. With the presence of citrate, the iron loading is
slightly lower than 10%. To figure out whether citrate could help to make nano-sized FeOOH, we
conducted pore volume analysis with the Fe-Ul (10.1), CA (2%)-Fe-Ul(9.5), CA (10%)-Fe-Ul(9.9).
Results were illustrated in Figure 3.3. From this result, citrate does help to make more porous iron
oxide loaded GAC. Both CA (2%)-Fe-Ul (9.5) and CA (10%)-Fe-Ul (9.9) has more pores than the
Fe-Ul (10.1). Also, CA (2%)-Fe-Ul (9.5) has slightly lower iron loading compared to CA
(10%)-Fe-Ul (9.9), but it has more pores.
Kinetic tests were conducted and results were illustrated in Figure 3.4. Although citrate-Fe
loaded carbon has higher pore volume, it doesn’t help adsorb more arsenic, on the contrary, the
67
arsenic adsorption decreased, especially for the carbon made from 10% citrate. When citrate bond
with iron and form amorphous iron oxide, it already takes some active site of the iron oxide, so
less active sites available for arsenic.
3.3.5 Rapid Small Scale Column Tests
Arsenic breakthrough behaviors for virgin carbon and various kinds of tailored
carbon are explored with rapid small scale column tests (RSSCT’s), and results are illustrated in
Figure 3.5 and Figure 3.6. All RSSCT’s herein were operated with pH 6, except as noted
otherwise.
Virgin Ultracarb has a 30 ppb arsenic breakthrough at merely 200 ppb. Citrate-Fe preloaded
carbons could be fairly effective for arsenic removal. This column exhibited 10 ppb breakthrough
at 5500-7000 bed volumes (BV); and they reached 25 ppb breakthrough at 8500-12,000 BV. In
accordance with the isotherm test and the kinetics test, CA-Fe-Mg (2.4) has a very sharp
breakthrough curve compared to CA-Fe (1.36).
GAC made with incipient wetness method has slightly longer bed life, with a10 ppb arsenic
breakthrough achieved at 9000 BV and 25 ppb breakthrough at 19,000 BV. The CA (2%)-Fe-Sd
(6.1) has a better performance than Fe-Sd (5.4). This may be attributed to two reasons: first, the
CA (2%)-Fe-Sd (6.1) has a higher iron loading than the Fe-Sd (5.4); secondly, at higher
temperature, the Fe state on the Fe-Sd (5.4) is more crystallized and thus less active compared to
the Fe on the CA (2%)-Fe-Sd (6.1).
68
3.4 CONCLUSIONS
Organic acid-Fe preloading carbon could achieve and iron loading amount of 1-3%. Highest
iron loading was achieved with an initial citrate-Fe concentration of 0.2 M. With incipient wetness
method, higher iron loading amount as 10% could be achieved.
Arsenic removal is generally higher with higher iron loading amount. But it also affected by
the pHzpc value of the carbon. With the addition of second metal Magnesium, higher iron loading
amount could be achieved, but the arsenic removal capacity was lower compared to citrate –Fe
preloaded carbon.
When loading Fe to carbon with incipient wetness method, the addition of 2% and 10%
(molar percent to Fe) citric acid made more porous carbon. But the arsenic removal capacity of
this carbon was lower because citrate took up some active site.
Citrate-Fe preloaded GAC could be fairly effective for arsenic removal, with a 10 ppb
arsenic breakthrough at 5500-7000 BV. The citrate acid-Fe preloaded GAC made through
incipient wetness method has slightly better arsenic removal.
3.5 REFERENCES
Benjamin, M. M., Sletten, R. S., Bailey, R. P., and Bennett, T. (1996). "Sorption and filtration of metals using iron-oxide-coated sand." Water Research, 30(11), 2609-2620.
Chen, J. Paul; Wu, Shunnian; Chong, Kai-Hau. Surface modification of a granular activated carbon by citric acid for enhancement of copper adsorption. Carbon. 2003, 41, 1979-1986.
Chen, Weifang ; Cannon, Fred S.; Rangel-Mendez, Jose R. Ammonia-tailoring of GAC to enhance perchlorate removal. II: Perchlorate adsorption. Carbon, 2005, 43, 581-590.
Clesceri, Lenore S. Standard methods for the examination of water and wastewater. American Public Health Association:Washington, DC. 1998.
Daus, Birgit; Wennrich, Rainer; Weiss, Holger. Sorption materials for arsenic removal from water:
69
A comparative study. Water Res., 2004, 38, 2948-2954. DeMarco, Matthew J.; SenGupta, Arup K.; Greenleaf, John E. Arsenic removal using a
polymeric/inorganic hybrid sorbent. Water Research, 2003, 37, 164-176. Dixit, S., and Hering, J. G. (2003). "Comparison of arsenic(V) and arsenic(III) sorption onto iron
oxide minerals: Implications for arsenic mobility." Environmental Science & Technology, 37(18), 4182-4189.
Driehaus, W., Jekel, M., and Hildebrandt, U. (1998). "Granular ferric hydroxide - a new adsorbent for the removal of arsenic from natural water." Journal of Water Services Research and Technology-Aqua, 47(1), 30-35.
Driehaus, W., Seith, R., and Jekel, M. (1995). "Oxidation of Arsenate(Iii) with Manganese Oxides in Water-Treatment." Water Research, 29(1), 297-305.
Gu, Z. M., and Deng, B. L. (2005). "Arsenic redox transformation and adsorption by GAC-based iron-containing adsorbents." Abstracts of Papers of the American Chemical Society, 230, U1577-U1577.
Gupta, V.K. ; Saini, V.K.; Jain, Neeraj. Adsorption of As (III) from aqueous solutions by iron oxide-coated sand. J. Colloid Interface Sci. 2005, 288, 55-60.
Jekel, M.; Seith, R. Comparison of conventional and new techniques for the removal of arsenic in a full scale water treatment plant. Source: Water Supply, 2000, 18, 628-631
Kolker, Allan; Huggins, F.E.; Palmer, C.A.; Shah, Naresh; Crowley, S.S.; Huffman, G.P.; Finkelman, R.B. Mode of occurrence of arsenic in four US coals. Fuel Proc. Technol. 2000, 63, 167-178.
Munoz, J.A.; Gonzalo, A. ; Valiente, M. Arsenic adsorption by Fe (III)-loaded open-celled cellulose sponge, Thermodynamic and selectivity aspects. Environ. Sci. Technol. 2002, 36, 3405-3411.
Patrick, John W. Porosity in carbons : characterization and applications. Halsted Press, 1995. Selvin, N.; Messham, G.; Simms, J.; Pearson, I.; Hall, J. The development of granular ferric
media-arsenic removal and additional uses in water treatment. Proceedings-Water Quality Technology Conferences, Salt Lake City, UT, 2000; 483-494.
Smith, A. H., Hopenhaynrich, C., Bates, M. N., Goeden, H. M., Hertzpicciotto, I., Duggan, H. M., Wood, R., Kosnett, M. J., and Smith, M. T. (1992). "Cancer Risks from Arsenic in Drinking-Water." Environmental Health Perspectives, 97, 259-267.
Thirunavukkarasu, O.S.; Viraraghavan, T.; Subramanian, K.S. Arsenic removal from drinking water using iron oxide-coated sand. Water, Air, and Soil Pollut. 2003, 142, 95-111.
Vaughan, R. L., and Reed, B. E. (2005). "Modeling As(V) removal by a iron oxide impregnated activated carbon using the surface complexation approach." Water Research, 39(6), 1005-1014.
Zhang, F. S., and Itoh, H. (2005). "Iron oxide-loaded slag for arsenic removal from aqueous system." Chemosphere, 60(3), 319-325.
70
Freudlich Isotherm
y = 0.0342x + 0.7263R2 = 0.8708
y = 0.3782x + 0.3684R2 = 0.9427
0.5
0.55
0.6
0.65
0.7
0.75
0.8
-0.1 0.1 0.3 0.5 0.7 0.9log Ce
log
qeCA-Fe (1.2)
CA-Fe-Mg (2.18)
Linear (CA-Fe-Mg(2.18))Linear (CA-Fe(1.2))
Langmium Isotherm
y = 0.4885x + 0.1335R2 = 0.9481
y = 0.014x + 0.1742R2 = 0.996
0.1
0.15
0.2
0.25
0.3
0 0.2 0.4 0.6 0.8 1 1.2
1/Ce
1/qe
CA-Fe (1.2)
CA-Fe-Mg (2.18)
Linear (CA-Fe(1.2))Linear (CA-Fe-Mg (2.18))
Figure 3.1. Adsorption Isotherm of Citrate-Fe preloaded GAC and Virgin GAC. (A)Freudlich Isotherm (B) Langmuir Isotherm
B
A
71
Kinetics
0
0.5
1
1.5
2
2.5
3
0 60 120 180 240 300
time (mins)
Ads
orpt
ion
(mg/
g)
CA-Fe-Mg (2.18)CA-Fe (1.2)
Figure 3.2 Kinetics tests of CA-Fe (1.2) and CA-Fe-Mg (2.18).
72
cumulated pore volume
0
0.1
0.2
0.3
0.4
0.5
1 10 100 1000 10000
pore width (Å)
cum
ulat
ed v
olum
e (c
m3 /g
)
CA-Fe-Ul (1.3%)virgin UlCA (2%)-Fe-Ul-IWCA (10%)-Fe-Ul-IWFe-Ul-IW
Figure 3.3 Pore volume analysis of virgin Ultracarb and various iron loaded Ultracarb.
73
Kinetics
Fe-Ul
CA (2%)-Fe-Ul
CA (10%)-Fe-Ul
0123456
0 100 200 300 400
time (min)
Ads
orpt
ion
Cap
.(mg/
g)
Figure 3.4 Kinetics tests of Fe loaded carbon made via incipient wetness method.
0
5
10
15
20
25
30
35
40
45
50
0 3000 6000 9000 12000 15000Bed volumes
Ce (ug/L)
Citrate-Fe(1.32)
Citrate-Fe-Mg(2.4)
Citrate-Fe-Mn(1.36)-pH 5
virgin
Figure 3.5 RSSCT’s of virgin and Citrate-Fe preloaded GAC.
74
0
10
20
30
40
50
60
70
80
90
100
0 10000 20000 30000 40000 50000
BVs
Ce (ug/L)
C i t rate-Fe-Sd(6.06)Fe-Sd (5.39)
Fe-Ul (3.18)
Figure 3.6 RSSCT’s of amorphous iron oxide preloaded GAC.
75
CHAPTER 4
Arsenic Removal from Groundwater with Iron Tailored GAC
Preceded by Precorroded Iron
By Jiying Zou, Fred. S. Cannon, Weifang Chen, Brian A. Dempsey
ABSTRACT
Precorroded iron material, coupled with Organic carboxyl-Fe preloaded granular activated
carbons (GAC), have been appraised as an innovative technique for removing arsenic from
groundwater. The favorable precorroded iron materials have included Galvanized Steel Fittings,
and Perforated Steel Sheets. Rapid Small Scale Column Tests (RSSCT’s) and mini column tests
had been conducted to evaluate the arsenic removal capacity of the procorroded iron coupled with
tailored carbon. The arsenic was removed through both the iron column and the GAC column,
with the GAC column serving as a major absorber and filter media. The water pH, precorrosion
condition and idling practice all affected both iron release and arsenic removal. The precorrosion
condition of iron has a short period effect on arsenic removal as the column just started, while
long term performance of the column was decided by the water pH and idling of the system. At
pH 6, the iron-GAC bed combination removed arsenic to below 10 ppb for as much as 250,000
bed volumes (BVs), with BV as measured through the GAC media. Idling the system extended the
system’s bed life by 2 times, but also caused brief periods of arsenic breakthrough after the
columns restarted.
76
4.1 INTRODUCTION
4.1.1 Background
In recent years, arsenic contamination of groundwater has emerged as a major concern on a
global scale. Lifetime exposure to arsenic in water can cause cancer to the liver, lung, kidney or
bladder on consumption of 1L/day of water at 50 µg/L arsenic level (USEPA 2001). Because of
this concern, the USEPA has dictated that all United States public water systems must comply with
a new 10 µg/L standard as of January 1st 2006. The work herein addresses a system that employs
iron-tailored GAC coupled with solubilization of zero valent iron; and this has been appraised
particularly for very small systems.
4.1.2 Arsenic Removal Technology
A number of authors have applied zero valent iron (ZVI) for removing arsenic (Bang et al.
2005; Karschunke and Jekel 2002; Leupin et al. 2005; Nikolaidis and Chheda 2001; Su and Puls
2003). For example, Nikolaidis, et al. (2003) conducted pilot tests with a ZVI filter that contained
ZVI plus sand with a weight ratio of 1:1. At pH 6, this filter removed arsenic from 294 ug/L down
to 20 ug/L for 18 000 BVs ~ 21 600 BVs. However, such an integrated ZVI/sand filter released
effluent iron as high as 70 mg/L (Nikolaidis and Chheda 2001). With a separate sand filter that
followed ZVI corrosion, Bang et al. (2005) was able to control iron effluent to less than 0.3 mg/L;
but this added to system complexity.
Much research had focused on developing iron-based arsenic adsorbents such as granular
ferric hydroxide (GFH) (Driehaus et al. 1998; Selvin et al. 2002; Westerhoff et al. 2005). The GFH
can effectively remove both As (V) and As (III) from aqueous solutions, and this media can be
77
used for a wide range of pH (pH < 9). GFH is reported to have a high treatment capacity of from
50 000-300 000 bed volumes to a 10 μg/L breakthrough level (Driehaus et al. 1998; Selvin et al.
2002; Westerhoff et al. 2005). However granular ferric oxide, as a poorly crystallized β-FeOOH,
is physically weak; and this media can crumble and disintegrate when employed for prolonged use.
Indeed, it has been recommended that GFH should be operated with pressures under 10 psi so as
to prevent the media from crumbling; and that GFH columns should receive backwashing
approximately every 3-4 weeks (Westerhoff et al. 2005).
The inherently simple features about activated carbon are that GAC columns are easy to
operate and very applicable to small water systems. Moreover, the GAC grains will remain
intact under considerable pressure; and GAC has been used for decades to treat water.
4.1.3 pH Effect on Arsenic Removal by ZVI and Iron (hydr)oxides
The pH of the water, relative to the pHzpc of iron oxide/hydroxide is a critical factor for
arsenic removal by iron systems. For a specific surface, when pH < pHzpc, the surface tends to be
positively charged and will attract anions such as HAsO42- and H2AsO4
-. Conversely, if pH >
pHzpc, the surface tends to be negatively charged and will electrostatically repel arsenic anions.
For Fe3O4, α-FeOOH, γ -Fe2O3, and amorphous Fe(OH)3, values of pHzpc = 6.5–8.5 have been
noted (Cornell and Schwertmann, 1996). The pKa values of arsenate species (As-V) are 2.2,
6.97, 11.53, while those values for arsenite species (As-III) are 9.29, 12.10 and 13.4. Thus,
arsenate removal by iron oxide/hydroxide is favored at acidic pH, while arsenite removal by those
media is favored at pH 7-8.
4.1.4 Iron Corrosion and Iron Release
Under anoxic conditions, iron corrosion can be dictated by autoreduction reactions in which
78
ZVI is oxidized (partially) to Fe(II), as per the following Equations 1-2 (Kuch 1988; Ritter et al.
2002; Ritter et al. 2003). Most Fe(II) species are soluble. Also, Fe(II) can convert to Fe(II-III)
in the form of magnetite (Equation 4); and this can be insoluble. The iron scale formed by such
dissolution/precipitation processes are claimed to be porous (Odziemkowski et al. 1998); and the
authors herein propose that this can be favorable for arsenic sorption (see below).
Fe2O3 + Fe + 6H+ = 3Fe2+ +3H2O (1) Fe + 2H+ = Fe2+ + H2 (2) Fe2+ + 2OH- = Fe(OH)2 (3) 4Fe2O3 + Fe = 3Fe3O4 (4)
Moreover, when aerobic conditions prevail at neutral pH, iron reactions can be governed by the
oxidation of ZVI or Fe(II) to Fe(III); and this yields insoluble particles and scales, as per
Equations 5 and 6:
4Fe + 3O2 + 2H2O = 4 Fe(OOH) (5) 4Fe (OH)2 + O2 + 2H2O = 4Fe(OH)3 (6)
For a fresh iron surface, the amount of iron released is correlated to the rate of iron
corrosion; whereas after the surface has been covered by corrosion products such as FeOOH and
Fe(OH)3, the amount of iron released could depend upon such parameters as dissolved oxygen
concentration, pH, carbonate / hardness presence. It can also depend upon the rate of diffusion of
reactant and product species through the iron scale or colloids. Higher pH favors the formation of
an impervious passive iron (hydr)oxide layer on the ZVI (Baylis 1926); and this layer can retard
iron corrosion by blocking electron transfer and mass transfer through the layer, thus decreasing
iron release (Karalekas et al. 1983; Nishimura et al. 1996).
Aged carbon steel and gray cast iron pipes in water distribution systems are generally
covered by layers of scales that formed by corrosion of the iron pipes; and they are claimed to
release colored water that can contain iron above 0.3 mg/L (Sarin et al. 2003; Sarin et al. 2001;
79
Sarin et al. 2004). Such iron was released by (a) the corrosion of iron metal, (b) the dissolution of
ferrous components of the scales, and (c) hydraulic scouring of particles from the scales (Sarin et
al. 2003). Iron release could be greatly reduced as the corrosion deposits accumulated (Tongesayi
and Smart 2008).
The objectives of the study herein are as follows: (i) devise an innovative arsenic removal
system that employs both corroding iron plus iron – tailored GAC; (ii) test the arsenic removal and
iron release as a function of practical operating parameters such as pH, idle times, pre-corrosion,
aging, etc.; and, (iii) study the mechanism of iron release and arsenic removal during the column
operation period. The practical aim of this work has been to devise an arsenic removal
technology that would be useful for small water systems.
4.2 MATERIALS AND METHODS
4.2.1 Materials.
All chemicals were reagent grade. Metal ions employed include Ferric chloride
(FeCl3·6H2O) and Magnesium chloride (MgCl2·6H2O) from Fisher Scientific Company. Citric
acid (HOC(COOH)(CH2COOH)2) originated from J.T.Baker Company. The activated carbon
employed was Ultracarb from Siemens Water Technologies.
Three kinds of iron materials were applied. (1) Perforated steel sheets from McMaster.com,
which were low carbon plain steel, with a thickness of 0.5 mm; the steel sheets had holes with a
diameter of 0.6 mm, and total opening area of 23%. The sheets were cut into 0.5~0.6 (±0.2) mm ×
0.5~1.2 (±0.2) mm before use. (2) Galvanized Steel Fittings which had been corroded while
processing groundwater before this use. (3) Cast gray iron from McMaster.com, which was cut
80
into small cubes, of 0.5 (±0.2) mm × 0.5 (±0.2) mm × 1.0 (±0.2) mm.
4.2.2 Citrate-Fe preloaded carbon
To make the citrate-Fe preloaded carbon, Ultracarb was mixed with the citrate acid plus
ferric chloride ((FeCl3·6H2O)) solution, and then agitated on a shaker table at 100-120 RPM for
2~3 days, and this protocol facilitated reaching pseudo-equilibrium for sorption of the Fe-Citrate.
After this mixing, the adsorbent was filtered out and washed with distilled water until no color in
the washing water could be discerned. This iron – tailored carbon was dried at 104ºC overnight
and stored in desiccators before use.
4.2.3 Iron Pre-corrosion
The precorrosion of perforated iron sheets and cast gray irons was conducted by soaking the
iron materials in acid solutions for 1-6 days. We experimented with two protocols of acid
precorrosion: 0.3 M nitric acid + 4 % oxalic acid, or 1 M nitric acid + 8 % oxalic acid. After the
precorrosion, the iron pieces were washed with DI water until the washing water pH exceeded 5.5.
After precorrosion, the iron media was in some cases aged by soaking in 50 – 80 ml deionized
water that was open to the air for 1-12 days, as specifically identified in Table 3.
4.2.4 Column tests
Rapid small-scale column tests (RSSCT’s) were conducted to evaluate the GAC’s arsenic
adsorption capacity; and these columns were designed to simulate the adsorption conditions that
would occur in a full-scale bed. The RSSCT’s in this paper were designed to simulate a full scale
Column (employing #12×40 GAC) with an EBCT of 10 minutes, per proportional diffusivity
simulation; and the configuration of these were as per Chen et al. (2005).
81
Several columns were operated with Galvanized Steel Fittings, and in this case, the
Galvanized Steel Fittings attached the GAC columns to the connecting tubing, such that the water
ran past Galvanized Steel Fittings just before it entering the GAC columns.
The authors also designed a mini column test, which shortened the column operating time.
For this apparatus, the pre-corroded iron was added in the iron mini column (see Table 4.1). And
the water was passed through the iron mini column first and then pass through the GAC column
with a flow rate of 3 ml/min. The EBCT of the mini column was 0.31 mins, and this aimed to
mimic a full size (#12×40) column with an EBCT of 4.9 mins. Other parameters have been
detailed in Table 1. For these RSSCTs and mini-column tests, we assumed that arsenate and
arsenite sorbed per proportional diffusivity onto activated carbon, in which case the following
equation applies (Hand et al. 1989):
EBCTSC/EBCTLC=DSC/DLC
Where EBCT is the empty bed contact time (volume of vessel / flow rate), SC is the
small-scale column, LC is the large column (or full-scale column) being modeled, D is the
geometric average size of the GAC grains (millimeters). The mini columns employed GAC grains
of US mesh size #200×400 so as to mimic full-scale columns that have a 4.9 mins EBCT when
the GAC mesh size is #12×40. When EBCTLC= 4.9 mins,
EBCTSC= EBCTLC ×DSC/DLC=4.9×54.4/850=0.313 mins.
All column tests were carried out at room temperatures of 20-23°C. The ground water
originated from the well of the Cool Sandy Beach Community Water System of Rutland, MA
(identified as the Rutland water in subsequent discussion). The total Arsenic in this groundwater
was 47-55 ppb except as noted otherwise. Characteristics of the groundwater were as presented in
82
Table 4.2. This Rutland water had a native pH of 7.3-7.6. For columns operating at pH 6, we
adjusted pH with 1N hydrochloride acid. The water pH was monitored at least once every two
days (except as noted otherwise), and was readjusted by 1N hydrochloride acid again if pH
exceeded the target (i.e. when higher than 6.4).
Table 4.1: Configuration of Rapid Small Scale Column Tests (RSSCTs) and mini columns
EBCT (min)
Column length (cm)
Column diameter
(cm)
Column Volume
(cm3)
Particle size (US mesh)
mass (g)
Iron column
- 6 1.2 6.8 - 0-0.41
Carbon mini- column
0.31 4.5 0.5 0.94 200×400 (38-15 µm)
0.56
RSSCT carbon column
0.64 13.5 0.5 2.82 200×400 (38-15 µm)
1.67
4.2. 5 Chemical Analysis
In order to test the iron loading on tailored GAC, the authors ashed a portion of the fully
loaded GAC in a muffle furnace at 600°C for 24 hours. The ashed GAC was dissolved in 25 mL of
concentrated HCl. After a minimum contact time of 24 hours, the solution was filtered and the
filtrate was then diluted to 250 mL. Solutions were analyzed for iron by the inductively coupled
plasma mass spectrometry (ICP-MS) method or the atomic absorption spectrometry (AAS,
Shimadzu®) method.
During the column operation periods, water samples have been collected from the outlet of
both the GAC column and the iron column. Some of these samples were filtered through a 0.2
micrometer syringe filter right after collection. Arsenic concentration presented herein represent
total arsenic, unless specifically listed as filterable, or specifically as As(V) or As(III). All samples
were digested with 0.4% Ultrapure nitric acid and stored in room temperature before testing.
83
Arsenic concentrations were determined with ICP-MS method and also a
hydride-vapor-generation flame atomic absorption spectrometry (HVG-AAS, Shimadzu®) method,
as per Chen et al. (Chen et al. 2007). The iron concentration were analysed by inductively coupled
plasma mass spectrometry (ICP-MS) method or the atomic absorption spectrometry (AAS,
Shimadzu®) method.
Table 4.2. Water quality characteristics of Cool Sandy Beach Groundwater (Rutland, MA)
Cations Concentration Units Anions Concentration Units Calcium (Ca) 59 mg/l CaCO3 Bicarb (HCO3) 64.2 mg/l CaCO3
Magnesium (Mg) 11.3 mg/l CaCO3 Fluoride (F) 0.670 mg/l CaCO3 Sodium (Na) 27.5 mg/l CaCO3 Chloride (Cl) 9.32 mg/l CaCO3
Potassium (K) 4.2 mg/l CaCO3 Nitrate (NO3) 0.041 mg/l CaCO3 Iron (Fe) < 0.003 mg/l Phosphate (PO4) < 0.080 mg/l CaCO3
Manganese (Mn) 0.003 mg/l Sulfate (SO4) 26.4 mg/l CaCO3 Aluminum (Al) <0.006 mg/l Silica (SiO2) 12.5 mg/l CaCO3
Zinc (Zn) 0.004 mg/l Arsenic* 47-55 μg/l Other
parameters Units
Other parameters
Units
pH 7.4-7.6 Total Hardness 70.30 mg/l CaCO3 Turbidity 0.08 NTU TOC (C) 0.851 mg/l
Conductivity 165 μS Free (CO2) 3.6 mg/l CaCO3
* 25% of arsenic was As(III) and 75% of arsenic was As(V).
In order to discern the extent to which the arsenic was removed in the columns that
operated with galvanized steel fittings, we conducted digestion of all the possible arsenic
adsorbents. The glass wool, tubes and galvanized steel fittings were digested according the
Standard Method 3030E and 3030F (APHA, 1995). The carbon bed was digested by the sequential
leaching method. This procedure has been described in details elsewhere (Huggins et al. 2002). A
revised 3-step leaching protocol was adopted here, which included: (1) 1N ammonium acetate, (2)
3N hydrochloride (3) 3N nitric acid. Each step was performed for 24 hours at room temperature.
84
4.3 RESULTS AND DISCUSSION
RSSCTs and mini column tests were conducted to test the arsenic removal capacity of the
columns under different conditions, and detailed operation parameters were presented in Table 4.3.
4.3.1 Arsenic removal with and without precorroded iron
RSSCTs
Arsenic breakthrough behaviors for iron - tailored GAC with and without corroded iron
source have been presented in Figure 4.1. Results indicated that when coupled with corroded
galvanized steel fittings, the GAC’s bed life was extended by over 20 times.
Both columns have employed citrate acid-Fe preloaded Ultracarb, with an iron loading of
1.2 ~ 1.3%. To explore the iron corrosion effect, one column employed 316 stainless steel fittings
(i. e. as the non-corrosion control) and the other employed corroded galvanized steel fittings.
Without corroded iron, the iron – tailored GAC (IT GAC) column (#1) exhibited 10 ppb
breakthrough at 7000 bed volumes (BVs). When corrosion of alvanized steel fittings was coupled
with iron – tailored GAC (GS #1), the effluent arsenic remained below 10 ppb for 150,000 BVs,
and showed no sign of progressing to full exhaustion even at 250,000 BV. GS #1 had remained
idled at 26,000 BV for 6 days. A similar column (GS #2) was conducted with Citrate-Fe-Mg
preloaded GAC, which has an iron content of 0.97%. GS #2 brokethrough slightly earlier at
around 130,000 BVs, otherwise, the result were very similar to column GS #1.
85
Table 4.3. Column operating parameters and 10 ppb breakthrougha
a. All columns employed GAC that preloaded with Fe (1.0-1.35%) plus citric acid, except GSF #3, which has no
GAC. Run #1, Run GSF #1-4 were RSSCTs, while PS #1-7 are mini -column tests.
b. RSSCT of iron tailored GAC without corroded iron.
c. RSSCT of glass wool with galvanized steel fittings, no GAC.
d. GSF #4 had only been operated for 25,000 BVs.
e. PS #1 was running at pH 6 with a short period pH increase to around 6.5.
f. PS #4 was running at pH 6 mostly, but at 35, 500 – 54,500 BV, pH was increased to 6.7.
* BVs as measured through the GAC column
Arsenic and Fe accumulated in the system.
The authors suspected that the high arsenic removal in PS #1 could be attributed to: (1)
Arsenic adsorption and coprecipitation by the iron (hydr)oxides on the corroded galvanized steel
fittings surface, or (2) arsenic coprecipitation and adsorption by iron (hydr)oxides particles that are
released from galvanized steel fittings. Such As-Fe particles could be filtrated by the glass wool
plugs, the GAC column; and they may also deposite on the inside wall of the effluent tubings.
Iron materials Precorrosion
method Precorrosion
time (day)Age (day)
Mass after precorrosion
(gram)
Operation pH
7 days idle
BV*s to 10 ppb consistent breakthrough
#1b No - - - - 6 no 6000 GS #1 & #2
Galvanized Steel Fittings
- - -
- 6 22,000 - 26,000
135,000 - 150,000
GS #3c Galvanized
Steel Fittings -
- - - 6 no 600
GS #4d Galvanized
Steel Fittings -
- - - 6 no 24,000
PS #1 Perforated
Steel sheets 1 M nitric + 8% oxalic
3 5 3.3 6e 27100 248,000
PS #2 Perforated
Steel sheets 1 M nitric + 8% oxalic
6 0 4 6
16220 124400 180800
215,000
PS #3 Perforated
Steel sheets 1 M nitric + 8% oxalic
3 1 3.5 6 no 103,500
PS #4 Perforated
Steel sheets 1 M nitric + 8% oxalic
3 0 4 6-6.5f 27,300 70,000
PS #5 Perforated
Steel sheets 1 M nitric + 8% oxalic
1 2 3.9 7.3-7.6 18,700 20,000
PS #6 & #7
Perforated Steel sheets
1 M nitric + 8% oxalic
1 10 0.65-0.75 7.3-7.6 20,000 1000-1300
86
To appraise these possibilities, after GS #1 was stopped after 250,000 BV, the authors
analyzed the Fe and As distribution into each of the possible Fe and As deposits, namely the GAC
media, the glass wool plug, the galvanized steel fittings and the effluent tubings (see Table 4).
Over the course of the 250,000 BV, the glass wool plugs at the inlet and outlet had been replaced
three times; and the Fe and As analysis were appraised for the third plug.
From Table 4, the third plug of glass wool at the inlet port sorbed 8.3 mg arsenic, while the
carbon bed accumulated 11.5 mg arsenic. By comparing influent aqueous As to the effluent As, we
calculated that the total arsenic removed was 29.1 mg. So the carbon bed captured about 40% the
removed arsenic, while the third plug of inlet glass wool accumulated about 30% of the removed
arsenic. The galvanized steel fittings only offered a slight net arsenic accumulation. The glass
wool plug and the carbon bed also accumulated 80 to 110 mg Fe. The ratio of As/Fe was
0.068-0.076 moles As/mole Fe for either the inlet glass wool or the GAC media; and these ratios
highlight the high efficiency of these corrosion products for capturing As in the combined system.
The high amount of the iron accumulated on the outlet tubing came from the outlet galvanized
steel fittings.
Arsenic removal by glass wool after corroding galvanized steel fittings
The glass wool in the GS #2 accumulated a considerable amount of arsenic; and this might
lead someone to wonder what performance could be achieved with just glass wool. To test this, the
authors conducted a column test with glass wool and corroded galvanized steel fittings, but no
GAC (GS #3). In this case, the As effluent reached 20 ppb within a mere 600 BV, while the iron
effluent increased from around 0 to over 0.3 mg/L after 15,000 BVs. So the blank glass wool alone
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offered only limited capacity for capturing iron and arsenic. Thus, the activated carbon played an
import role in the arsenic removal process.
Table 4.4. Arsenic distribution in GS #1 (iron - tailored GAC coupled with corrosion of galvanized steel fittings) after 250,000 BV
Adsorbents Fe (μg) As (μg) As/Fe (g/g)
As/Fe (mole/mole)
Outlet tubing 26,200 760 0.029 0.022 Inlet Glass wool (third plug) 82,400 8,380 0.100 0.076
Outlet Glass wool (third plug) 1,480 230 0.159 0.12 HCl washed rust from Galvanized steel
fittings -- 1,600 -- --
0.75 g carbon* 40,730 3,724 0.091 0.068 Normalized to 1.67 g carbon 113,557 11588 0.101 0.075
Total (with 1.67 g carbon) 223,637 22558 0.096 0.072 *A representative 0.75 g activated carbon was evaluated, out of 1.67 g loaded activated carbon present at the start of this RSSCT run.
The authors also conducted a short run (GS #4), which has similar operating conditions but
was only operated for 25,000 BVs, this column (GS #4) reached 10 ppb breakthrough at 24,000
BV. After the column was stopped, we conducted the same digestion to the GAC media inside the
column. Results revealed that over 90% arsenic was removed by the GAC column. Comparision
of the results from GS #1, #2, and #4, with that from the glass wool column (GS #3) indicated that
the arsenic removal was initially preferably removed by the GAC media, and then subsequently,
more arsenic was removed by the inlet glass wool as time progressed and the GAC column
clogged up and the glass wool accumulated iron particles.
4.3.2 pH effect on Arsenic removal
The authors appraised how water pH influenced arsenic removal by running three mini
column systems: the first operated mostly at pH 6.0 (PS #1), the second at pH 6.0, with a time
span at pH 6.5 (PS #4), and the third at pH 7.5 (PS #5). For all three of the runs, a column of
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precorroded steel sheets (PSs), preceded the mini column of iron – tailored GAC. All three column
systems had been idled for 7 days, as indicated in Figure 4.2. The water pH clearly influenced the
system’s performance: the pH 6.0 run (PS #1) exhibited consistent 10 ppb As breakthrough at
248,000 BV, where the pH 7.5 run (PS #5) exhibited breakthrough at 20, 000 BV (Figure 2).
Indeed, even a short period of pH increase diminished not only the concurrent As removal, but
also the long term bed volume to continuous As breakthrough. This perspective comes from
evaluating the PS #4 data. This system was operated at pH 6.0±0.3 for 39,000 BV, then the water
pH was increased to 6.6-6.7 from 39,000 to 58,000 BV; after which, the water pH was returned
back to 6.0±0.3. As shown in Figure 2, after the pH increased, the arsenic effluent from PS #4
gradually increased from 8 to 33; then when the pH dropped back to 6.0, the arsenic effluent
concentration dropped back to below 10 ppb unitl 70,000 BV, when consistent breakthrough of >
10 ppb occurred. In comparison, when the pH was consistently maintained at 6.0 throughout a run,
As breakthrough above 10 ppb did not occur until 248,000 BVs (PS #1).
The PS #1 run did experience one short period arsenic breakthrough period from 45,000
to 70,000 BVs; and this could be attributed to a slight pH increase when pH was not monitored or
adjusted for 8 days (during a holiday). The authors had noticed that the Rutland water pH changed
with time after it received HCl dosing. Specifically, after we adjusted 19-22L of Rutland water to a
pH of 5.8±0.2, we observed that within one day, the water pH increased to 6.2±0.2, even when its
container was capped. When not capped, as the water was consumed by the arsenic removal
columns, the pH rise to 6.3~6.5 within 2 ~ 3days after a quarter of the 20L water was left. It’s
possible that when the water was left for 8 days without monitoring and adjusting pH, its pH value
increased to around 6.5 when the water – air partition reached equilibrium. Rutland water has a
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bicarbonate concentration of 64.2 mg/L (as CaCO3); and thus pH adjustment was somewhat
buffered by the bicarbonate. In order to keep the pH value stable between pH 5.8 to 6.2, the water
pH need to be monitored at least once a day and adjusted by 1 N HCl as needed. This protocol was
maintained in all subsequent runs, except in PS #4, where we aimed to temporary increase the pH.
pH effect on arsenic removal in iron column
Results from Figure 4.2 (A) and 4.2(B) suggested that arsenic was removed both by the Fe
column and the GAC column, with the GAC column as the major absorber. After 55,000 BV
operation, the arsenic accumulated in the Fe bed was calculated to be 0.1 mg, 0.46 mg and 0.54
mg for column running at pH 7.5, at pH 6-6.5 and at pH 6 respectively; and this highlights the
higher arsenic removal at lower pH.
These observations of better arsenic removal performance with lower pH is in agreement
with previous studies of arsenic removal by iron (hydr)oxides and zero valent iron (Bang et al.
2005; Jain et al. 1999; Lenoble et al. 2002). The pH effect on arsenic removal by iron
(hydr)oxides could be assigned to the arsenic speciation and differences in surface charge. Per the
literature, the pHzpc value of various kinds of iron (hydr)oxides ranges from 6.5 to 8.5 (Stumm,
1981), for examples, magnetite exhibited a pHzpc value of 6.3-6.5 (Stumm, 1981), while
amorphous iron hydroxide or HFO usually has a pHzpc value close to 8 (Dzombak and Morel,
1990). For a specific surface, when pH < pHzpc, the surface tends to be positively charged and
will attract anions such as HAsO42- and H2AsO4
-; in contrast, when pH > pHzpc, the surface tends
to be negatively charged and will repel anions.
75% of the arsenic species in Rutland water were in the As(V) state, while 25% of them
are in the As(III) state. The pKa values of arsenate are 2.2, 6.97, 11.53. For arsenate species, at pH
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7.5, 78% of the arsenate are H2AsO4-, and 22% are HAsO4
2-; while at pH 6, 90.2% are
H2AsO4-and 9.8% are HAsO4
2- (Calculated via. Visual MINTEQ). The pKa value of arsenite are
9,29 and 12.10 and 13.4, so it is almost 100% H3AsO3 at pH 6 - 7. 5. When considering the
arsenic speciation in Rutland water and pHzpc of iron (hydr)oxides surface, one can note that at
pH 6 the iron (hydr)oxides surface has more active sites for arsenic removal.
The authors note that there was a distinction between arsenic removal by iron column in
PS #1 and PS #4 in the first 20,000 BVs even thought the two runs were both operated at pH 6.0
(Figure 2 (B)). The PSs in PS #1 had been aged for 5 days, while those in PS #4 were not aged
(See Table 3). The authors suspected aging may have created a more extensive surface corrosion
layer which contain more fragile iron (hydr)oxides (Sarin et al. 2001, 2004); these fragile iron
(hydr)oxides would easily detach from the PS surface and transfer to the GAC column, where they
were filtered and captured more arsenic. This hypothesis is reasonable considering the fact that in
the first 20, 000 BVs, the total iron release amount, filterable iron release amount and the filterable
arsenic amount from iron columns were all higher with run PS #1.
These observations indicate that there may be other factors affecting arsenic removal. The
surface layer characteristics and composition may contribute to the arsenic removal in several
ways: (1) Different corrosion product will have different capacity toward arsenic removal. For
example, amorphous iron (hydr)oxides like GFO and HFO are known to have better arsenic
removal because of their higher and more reactive surface area. (2) The corrosion product and
morphology may control the mass transfer efficiency. Baylis (1926) observed that higher pH led to
the formation of an impervious iron membrane on the scale, whereas lower pH resulted in the
formation of a fibrous, porous scale structure. The later will be more beneficial to arsenic removal
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because the porous structure makes more active sites accessible to arsenic. (3) The amorphous iron
(hydr)oxides formed on the inter-surface of iron scales and water solution are fragile, and tend to
be released to water. When arsenic are adsorbed on those iron (hydr)oxides, the release of iron
particles resulted in arsenic release. Comparing the arsenic removal and iron release data from
Run PS #1 and #4, we noticed that in the first 20,000 BVs, the total iron release amount, filterable
iron release amount and the filterable arsenic amount from iron column were all higher with PS #1.
These observations suggested that the difference of arsenic removal in iron column from PS #1
and #4 as column just started could be attributed to the last reason listed above. The PSSs in
column #1 has been aged (Table 3), the authors suspected the higher amount of fragile iron
(hydr)oxides was created during the aging of the PSSs.
pH effect on iron release and arsenic removal in GAC column
Usually a low pH was accompanied by a low alkalinity and a high iron release (Karalekas et
al. 1983). The authors suspected that pH might also affect the arsenic removal by varying the
amount of Fe released from iron column and the amount of Fe accumulated in GAC column. In
order to test this hypothesis, the iron release from iron column and iron effluent from the GAC
column were monitored and the cumulative amount was calculated and presented in Figure 4.3 (A)
to (C). From Figure 4.3, the iron release amount from the three columns follow the order of
column #1 > column #5 > column #4. This observation seems contradict to the general rules that
the iron corrosion rate decrease as pH increases. But for a corrosion product covered iron surface,
the iron release amount is not in accordance with corrosion rate, it is controlled by the surface
corrosion layer (Sarin, et al. 2001; Sarin, et al. 2004). PSSs in Run PS #1, #4 and #5 had been
precorroded for 3 days, 3 days and 1 day, respectively (Table 4.3), PSSs in Run PS #1 and #5 had
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also been aged for 5 days and 2 days. 1 day of precorrosion may have formed a thinner corrosion
layer, which favors iron corrosion and release in Rutland water. As discussed before, the higher
filterable iron amount released from the iron column of PS #1 and #5 indicated that after aging,
fragile iron (hydr)oxides had formed and tended to detach from the corrosion layer. This result is
in agreement with the observation of higher arsenic removal in a fresh PS column (Figure 4.2 (B)).
We correlated the 10 ppb breakthrough bed volumes with several parameters, the results (Table
4.5) revealed that the 10 ppb breakthrough BVs are significantly correlated with the As removal in
Fe bed and the amount of Fe accumulated in GAC.
Table 4.5. Correlation of 10 ppb As breakthrough
Total Fe released from Fe bed
Fe accumulated in GAC
As removed in Fe bed
Filterable Fe released from Fe Bed
Correlation (R2) to BV at consistent 10 ppb As breakthrough
0.5349 0.9974 0.9793 0.4396
Note: all parameters were collected to 55,000 Bed Volumes, because the column running at pH 7.5 only have data
to 55,000 BVs. Data are collected from Run PS#1, PS #3 ~ 5.
* Filterable Fe here means the iron amount that can’t pass the 0.2 micrometer syringe filter.
4.3.3 Idle Effect on Arsenic Removal
Some of our initial results indicated that arsenic removal performance was affected when
the treatment system was idled such that no new water flowed through the beds for several days.
We aimed to appraise this affect with controlled experiments that compared no idle interval (PS
#3), one idle interval (PS #1), and three idle intervals (PS #2), results were presentedin Figure 4.4
- 4.5. All three runs were operated at pH 6 ± 0.3 except otherwise noted and each idle period
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lasted for 7 days.
From the results, we observed: (1) the column’s bed life was extended from 103,000 BV
(PS #3) to 248,000 BV when the column was idled for one time (PS #1), but 3 idle intervals (PS
#2) didn’t improve the column’s performance compared with one idle period. (2) Without any stop
or idle, the system exhibited a rather smooth breakthrough curves; On the contrary, the idling
intervals caused a short period arsenic breakthrough. Also, runs with idling intervals exhibited a
sharp final breakthrough curve. (3) The arsenic removal in Fe columns and filterable iron release
amount tends to be higher tends to be higher after idling.
Idle effect on arsenic removal in iron column
The high arsenic removal by Fe column as the column just started could be attributed to
the availability of active sites on the iron (hydr)oxides formed by pre-corrosion. Likewise, when
the column was restarted, the same trend could generally be observed, which may indicate the
active sites have developed during the idling period. As discussed below, this is attributed to the
reduction reaction.
The more active sites could be developed in two ways. First, iron corrosion will create more
active sites. During the idle period, after the dissolved oxygen in groundwater was consumed up
with iron corrosion, the column became anaerobic, and Fe corrosion may happen as reaction 1- 5
(Kuch 1988; Ritter et al. 2002; Ritter et al. 2003). Iron corrosion resulted in the local pH increase
at the cathode site on the iron surface, which precipitates ferrous iron as ferrous hydroxide.
Ferrous hydroxide is not the thermodynamically favorable state and will transform into magnetite
independent of water composition (Odziemkowski et al. 1998). Both reactions create fresh active
sites for arsenic removal. By reaction 2, some Fe (III) oxide is dissolved and the reaction sites that
94
previously covered by them becomes available for arsenic removal.
The arsenic removal by iron (hydr)oxides like GFH and HFO is generally a diffusion
controlled reaction. Previous research (Fuller et al. 1993; Yu et al. 2006) indicated that arsenic
diffusion into available active sites controls the reaction rate. In our case, the intra particle
diffusion controls the in-situ arsenic removal in the iron column; when the columns were idled, the
arsenic that had previously adsorbed on the outer surface of the iron oxide could diffuse into inner
pores of the iron scales, and this process would render more available active sites on the outer
surface for arsenic removal after the column restarted.
Idle effect on iron release from iron column
The results presented in Figure 4.5 showed that the more we idle the columns, the more the
iron released from Fe column; and as a result, the higher the iron effluent from the GAC column.
The filterable iron amount also increases with more idle times. This phenomena is more evident
when the column was just restarted, suggesting the formation of more dissolvable iron scales
during idling period. As discussed previously, Reaction 1-5 might happen during the idle period,
the dissolution of iron (hydr)oxides and the precipitation of ferrous hydroxide and magnetite
promoted the formation of new porous scales, with dissolved ferrous ions dispersed in those pores,
as proposed by Odziemkowski et al. (1998). After column restarted, the Rutland water with DO
concentration ranging from 4-6 mg/L entered iron column, at the iron scale and groundwater
inter-surface, reaction 6-8 might happen; as a result, the amorphous iron oxide formed on the scale
surface, these iron hydroxides are physically very weak and easy to detach from the iron surface
even with little hydraulic change. This explains the high total iron and filterable iron amount after
column restart. The amorphous iron hydroxide will transform into thermodynamically more stable
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state as magnetite and goethite, the later are stronger and the total and filterable iron release
amount became comparably stable.
Fe (OH)2 + O2 → Fe(OH)3 (6) Fe2+ + O2→ Fe3+ (7) Fe3+ + 3OH- → Fe(OH)3 (8)
It is worth mentioning that with idle periods, even at stable operation conditions, i.e. not
after the start and restart of the columns, the filterable iron amount was still fluctuating. This may
also indicate a rather porous and amorphous iron (hydr)oxides surface in the iron column.
Idle effect on arsenic removal in GAC column
As mentioned previously, the arsenic removal amount was highly related to the iron
accumulation amount in the GAC column. After idling, the iron release amount from the iron
column increased, so did the Fe accumulation in the GAC column, which directly resulted in more
arsenic removal by these iron (hydr)oxides in the column. However, this effect only manifest itself
after the column had recommenced flow for some time. Immediately after column restarted, there
was a short period of arsenic breakthrough, indicating that arsenic was released from the GAC
column.
Two reasons may have caused the arsenic release from GAC column. (i) Arsenic maybe
released with the transformation of amorphous iron (hydr)oxides to crystalline iron (hydr)oxides.
Ferrous iron is known to induce the solid-phase transformation of ferrihydrite to magnetite and
goethite (Benner et al. 2002; Hansel et al. 2003), which has lower arsenic removal capacity
because of lower surface area. (ii) The arsenic maybe released when arsenate was reduced to
arsenite. Sorbed ferrous iron is always much more reactive than dissolved iron(II), indeed, rates of
homogeneous reduction by dissolved Fe(II) are exceedingly slow, surface Fe(II) complexes
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formed with hydrous iron oxides, are on the other hand very efficient reductants from a
thermodynamic as well as from a kinetic point of view (Stumm and Sulzberger 1992). The
reduction of arsenate by adsorbed Fe(II) under anoxic conditions has been observed (Charlet et al.
2002; Johnston and Singer 2007). The release of As due to its reduction to the more weakly
adsorbed As(III) has also been observed by Burnol et al.(Burnol et al. 2007).
4.3.4 Precorrosion iron amount effect
The amount of total PSs in iron column may also have influence on arsenic removal. As
shown in Figure 4.6 (A), by increasing the precorroded iron amount from 0.75 gram to 4 gram, the
10 ppb arsenic breakthrough bed volumes was extended from 1300 BVs to 20,000 BVs at pH 7.5.
Unlike the previous cases, in which better arsenic removal in the systems also correspond
to a higher arsenic removal by the iron column, the column with 4 gram precorroded steel (PS #5)
have less arsenic removal in the iron column compared to the one with only 0.6 gram precorroded
steel (PS #6).
According to our previous experience, the better performance of Run PS #5 could be
related to the higher iron accumulation amount in GAC. As shown in Figure 4.7 (A) to (C), the
total iron release amount, the filterable iron amount and the Fe accumulation in GAC column are
all higher with 4 gram precorroded iron. This fact further proved that Fe accumulation in GAC
column is an important factor to the arsenic removal.
4.3.5 Iron release and arsenic removal in iron column – A summary
The iron release and arsenic removal from iron columns were calculated and summarized in Table 4.6.
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Table 4.6. Fe release amount and arsenic removal in iron column
Column PS #1 PS #2 PS #3 PS #4 PS #5 PS #6 PS #7 Operation pH 6 6 6 6-6.5 7.5 7.5 7.5
Precorroded PS amount (gram) 3.3 4 3.5 4 4 0.75 0.65
Precorroded time (day) 3 6 2 3 1 1 1
Aging time (day) 5 0 1 0 2 12 10
Total Fe (mg/g PSS) 12.42 22 6.66 7.25 8.72 10.27 13.2Filterable Fe (mg/g PSS) 3.333 5 1.2 0.825 2.44 8.667 10.3Non-filterable Fe (mg/g PSS) 9.091 17 5.46 6.425 6.28 1.6 3.23Fe in GAC (mg/g PSS) 10 15.75 6.49 7 2.56 10 11.7
0 - 24,500 BV
As in Fe (mg/g PSS) 0.027 0.05 0.1025 0.03 0.427 0.8
Total Fe (mg/g PSS) 22.73 33.75 12.2 9.75 14.2
Filterable Fe (mg/g PSS) 7.273 10 2.17 1.325 4.64 Non-filterable Fe (mg/g PSS) 15.45 23.75 10 8.425 9.51 Fe in GAC (mg/g PSS) 20.3 25.5 12 9.475 7.03
0 - 55,000 BV
As in Fe (mg/g PSS) 0.058 0.15 0.115 0.03
From 0 ~ 24,500 BV, the filterable iron release amount follows the order of #7 > #6 > #2 >
#1 > #5 > # 3 > #4. This series indicated that aging is the most important factor for controlling
filterable iron release, more aggravated precorrosion also improves filterable iron release amount.
The non filterable iron release amount follows the order of #2 > #1 > #4 > # 5 > #3 > #7 > #6,
indicating that the precorrosion condition is the most important factor controlling non-filterable
iron release. Moreover, idling and lower pH also improve the amount of non-filterable iron release,
long aging time of more than 10 days retarded the non-filterable iron release. While aging for less
than 5 days showed no big difference. The Fe accumulation amount in GAC column follows the
order of #2 > #7 > #6 = #1 > # 4 > #3 > #5, suggesting that precorrosion condition and aging are
both important for increasing the Fe accumulation in GAC, Idling and lower pH also have minor
effects. The arsenic removal amount in Fe column follows the order of #7 > #6 > #4 > #3 > #5 >
#1.
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Results from 0 ~ 55,000 BV only have data for Run PS #1, and #3 to #5. Since Run PS #4
has a pH increase to 6.5, the non-filterable iron release amount, Fe accumulation amount in GAC
and arsenic removal in iron column are all lower for PS #4. This difference indicated that after the
column had been operated for some time, the effect of precorrosion and aging became allevated
while pH effect became important.
4.4 CONCLUSIONS
In general, the arsenic removal by the Fe-preloaded GAC plus precorroded iron system
offers promise for practical applications. To get desired performance, pH need to be slightly
adjusted to 6, the 10 ppb arsenic breakthrough occurs from 100,000 BVs to 250,000 BVs under
this pH; the iron effluent could be controlled under 0.2 ppm when column that employed fresher
PSSs was operated continuously. The study of the idle effect is beneficial relative to the practical
applications of the system, because many groundwater systems are operated in on-off mode.
Stop the column and idle for several days extended the column’s bed life by 2 times. But a short
period arsenic breakthrough will happen after column restarted; and this represents a limitation of
a system such as this, where a bed of corroding iron precedes a bed of GAC or other media that
can participate in redox reactions.
The arsenic was removed in both the iron column and the GAC column, with GAC column
as the major absorber. Arsenic removal in the GAC column was proportional to the iron amount
accumulated in the GAC column. The iron amount accumulated in the GAC column was generally
controlled by the operating pH, but also affected by the precorrosion conditions of the iron and the
99
idling of the whole system. The precorrosion conditions is suspected to control the formation of a
surface corrosion layer of the iron, which in turn will affect how the iron was released and
accumulated in the GAC column, especially when the column just restarted. Idling the columns
for 7 days is suspected to promote the dissolution of Fe(III) oxide/hydroxide and the precipitation
of Fe(II) oxide/hydroxide, the dissolution/precipitation process favors the formation of a porous
scales structure, which resulted in high iron release after column restart.
The arsenic removal in the iron column is generally higher with the lower pH of 6.0,
however, as the column just started, it is more likely to be controlled by the iron pre-corrosion
condition. Longer precorrosion period seems to have promoted arsenic removal in the iron column.
The arsenic removal is generally lower with aged PSSs as the column just started; and this was
attributed to the release of iron (hydr)oxides particles from the iron column.
The arsenic removal mechanisms in the precorroded iron plus Fe preloaded GAC system is
a complex process involving corrosion, adsorption, co-precipitation, transport and redox reaction.
By far, it is still not clear to us what kind of iron (hydr)oxides formed on the precorroded PSSs
surface, how did they transform with the process of arsenic removal under different pH, what kind
of arsenic reaction happeded during the idling period in both columns, how did that affect the iron
and arsenic release. To get a fundamental understanding of Fe release, arsenic and Fe interaction
and arsenic removal mechanisms, we still need to further study the following area: the
composition and structure of the iron corrosion product, especially the very surface layer, with the
change of pH, precorrosion condition and idling; the iron corrosion, iron and arsenic redox
reaction during the idle period; and the arsenic and Fe interaction in the GAC column. And these
subjects will be discussed in the next chapter.
100
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103
As removal with and without corroded iron
Without iron
With
galvanized
steel fitting
0
10
20
30
0 50000 100000 150000 200000 250000
Bed Volumes
As effluent (ug/L)
Figure 4.1 RSSCT of iron tailored GAC with (solid triangle, GS #1) and without (hollow square, #1) corroded iron, both columns operated at pH 6±0.3. Rutland groundwater as influent (As 47~55 ppb, Fe < 3 ppb). Dashed line indicated where the column (solid triangle) was stopped and ceased for 6 days.
104
pH effect on As removal
0
10
20
30
40
50
60
0 80000 160000 240000 320000
Bed Volumes
As effluent
(ug/L)
p H 7.5 (PS #5)
pH 6-6.5 (PS #4)
pH 6 (PS #1)
As removal by Fe bed at different pH
0
5
10
15
20
25
0 20000 40000 60000 80000 100000 120000 140000
Bed Volumes
As removed by Fe
bed (ug/L)
p H 6 (PS #1)
pH 6-6.5 (PS #4)
pH 7.5 (PS #5)
Filterable arsenic from iron column
0
10
20
30
40
50
0 50000 100000 150000 200000
Bed Volumes
Filterable As from
iron column (ug/L)
pH 6 (PS #1)
pH 6-6.5 (PS # 4)
Figure 4.2 pH effect on Mini column performance (A) Arsenic effluent from GAC column. (B) Arsenic removed by iron column. (C) Filterable arsenic (arsenic that didn’t pass the 0.2 μm syringe filter) after iron column. Lines indicated where the columns were idled for 7 days (Dotted line - pH 7.5, dashed - pH 6, Solid line - pH 6-6.5). PS #4 was operated at pH 6 most of the time, except from 35500 BV to 54500BV, where the water pH was increased to 6.5-6.7.
B
A
C
105
Total Fe Release from iron column
0
50
100
150
200
250
300
0 50000 100000 150000 200000 250000 300000
Bed Volumes
total iron
released (mg) p H 6 (PS #1)
pH 6-6.5 (PS #4)
pH 7.5 (PS #5)
Filtrable Fe Release from iron column
0
20
40
60
80
0 50000 100000 150000 200000 250000 300000
Bed Volumes
Filtrable iron
released (mg) p H 6 (PS #1)
pH 6-6.5 (PS #4)
pH 7.5 (PS #5)
Fe accumulated in GAC columns
0
100
200
300
400
0 50000 100000 150000 200000 250000 300000
Bed Volumes
Fe accumulated in
GAC (mg/g)
p H 6 (PS #1)pH 6-6.5 (PS #4)pH 7.5 (PS #5)
Figure 4.3 pH effect on (A) Total Fe release. (B) Filterable Fe release. (C) Fe accumulated in GAC column.
C
B
A
106
Total As effluent from GAC column
0
20
40
60
80
0 50000 100000 150000 200000 250000 300000 350000
Bed Volumes
As effluent (ug/L)
no idle (PS #3)idle once (PS #1)idle 3 times (PS #2)
As removed by Fe column
0
10
20
30
40
50
0 50000 100000 150000 200000
Bed Volumes
As in Fe (ug/L)
no idle (PS #3)
idle once (PS #1)
idle 3 times (PS #2)
Filterable As from iron column
0
10
20
30
40
50
0 50000 100000 150000 200000
Bed Volumes
Filterable As from
iron column (ug/L)
Idle once (PS #1)
no idle (PS #3)
Figure 4.4 Arsenic removal with no idle (open diamond, PS #3), one idle (solid reactangle, PS #1) and 3 idle (solid triangle, PS #2). (A) As effluent from GAC column. (B) As removal in Fe column. (C) Filterable arsenic from Fe column. Solid line indicate where PS #2 was stopped for 7 days, dashed line indicate where PS #1 was stopped for 7 days. All columns were operated at pH 6±0.3.
C
B
A
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Total Fe from Fe column
0
0.5
1
1.5
2
2.5
3
3.5
4
4.5
0 50000 100000 150000 200000 250000 300000 350000 400000
Bed Volumes
Fe conc. (mg/L) i d le 3 times (PS #2)
idle once (PS #1)
no idle (PS #3)
Filtrable Fe released from Fe column
0
0.5
1
1.5
2
2.5
3
0 50000 100000 150000 200000 250000 300000 350000
Bed Volumes
Fe conc. (mg/L) i dle 3 times (PS #2)
idle once (PS #1)
no idle (PS #3)
Total Fe effluent from GAC column
0
0.2
0.4
0.6
0.8
1
1.2
1.4
0 50000 100000 150000 200000 250000 300000 350000 400000
Bed Volumes
Fe conc. (mg/L) i dle 3 times (PS #2)
idle once (PS #1)
no idle (PS #3)
Figure 4.5 Idle time effect on Fe release. (A) Total Fe released from iron column. (B) Filterable iron (Iron that can’t pass the 0.2 micrometer syringe filter) released from Fe column. (C) Fe effluent from GAC column. Solid line indicates where the column with 3 idle times was stopped and idled for 7 days each time. Dashed line indicates where the column with 1 idle time was stopped and idled for 7 days.
A
B
C
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Total As effluent from GAC column
0
5
10
15
20
25
30
35
40
0 20000 40000 60000 80000
Bed Volumes
As effluent (ug/L)
P S S 0.6g (PS #6)
PSS 4g (PS#5)
As removed by iron column
0
5
10
15
20
0 10000 20000 30000 40000 50000 60000
Bed Volumes
As removal in Fe
(ug/L)
P S S 0.6 g (PS #6)
PSS 4 g (PS #5)
Figure 4.6 The effect of precorroded iron amount on arsenic removal. (A) Arsenic breakthrough curve. (B) Arsenic removal by Fe column. Both columns were operated at pH 7.5. Dashed line indicated where Run #6 was idled for 7 days, solid line indicated where PS #5 was idled for 7 days.
B
A
109
Total Fe released from Fe column
0
10
20
30
40
50
60
0 10000 20000 30000 40000 50000 60000
Bed Volumes
Total Fe
released(mg)
P S S 4 g (PS#5)
PSS 0.6 g (PS#6)
Filtrable Fe released from Fe column
0
5
10
15
20
0 10000 20000 30000 40000 50000 60000
Bed Volumes
Filtrable Fe (mg)
P S S 4g (PS #5)
PSS 0.6 g (PS #6)
Fe accumulation in GAC column
05
101520
2530
3540
0 10000 20000 30000 40000 50000 60000
Bed Volumes
Fe in GAC (mg)
P S S 4g (PS #5)
PSS 0.6g (PS #6)
Figure 4.7 The effect of precorroded iron amount on (A) Total Fe release, (B) Filtrable Fe release, (C) Fe accumulation in GAC column. Both columns operated at pH 7.5.
B
A
C
110
CHAPTER 5
Arsenic removal mechanisms in a precorroded steel sheets plus iron
preloaded activated carbon column systems
ABSTRACT
In our previous study, precorroded steel sheets plus iron preloaded activated carbon had showed
considerable capacity for arsenic removal. In this study, the morphology and structure of surface
corrosion products on precorroded steel sheets had been studied via Scanning Electron
Microscopy (SEM), X-ray Diffraction (XRD) and X-ray Photoelectron Spectroscopy (XPS)
method. The results indicated that iron release and arsenic removal are highly related to surface
corrosion products. Fresh precorroded steel sheets have a uniform surface, while aged precorroded
steel sheets exhibited a heterogeneous surface with some area covered with thick, porous scales.
Lepidocrocite (γ- FeOOH), humboditine (FeC2O4(H2O)) and clinoferrosilite (Fe1.5Mg0.5Si2O6) are
the main component on the fresh precorroded steel sheets, while goethite (α- FeOOH),
lepidocrocite (γ- FeOOH) and magnetite (Fe3O4) are the primary component of the aged
precorroded steel sheets surface. After precorroded steel sheets were employed in the column for
arsenic removal, the primary phase changed to goethite (α- FeOOH ) and magnetite (Fe3O4),
calcite (CaCO3) was also detected. Arsenic extraction from precorroded steel sheets in the iron
columns that had been idled and were operated at pH less than 7 showed the existence of only
As(III). X-ray absorption fine structure (XAFS) study of the GAC in pH 7.5 columns showed the
presence of reduced iron phases such as wustite (FeO) and green rust, the arsenic edge tests of
selected GAC samples indicated the reduction of As(V) occured. All these observations indicated
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that reduction reaction happened in both Fe column and GAC column.
5.1 INTRODUCTION
5.1.1 Surface corrosion layer and its effect on contaminant removal
Oxidation of Fe0 can proceed along several reaction pathways. In reduced environments
at low temperature, Fe(OH)2 is stable, but is predicted thermodynamically to convert to either
magnetite or intermediate products (green rusts).
Iron metal can be transformed into green rusts such as [Fe4(OH)8Cln·H2O],
[Fe6(OH)12][CO3·nH2O] in moderately neutral solutions ( 6.5 < pH < 8.0). Green rust as an
intermediate product during the hydrolytic oxidation of Fe2+ solutions often transform to goethite
(α-FeOOH), lepidocrocite (γ-FeOOH), maghemite (Fe2O3), or magnetite (Fe3O4), depending on
rate of oxidation and dehydration (Myneni et al. 1997). Rapid oxidation of ferrous hydroxides
forms lepidocrocite (γ-FeOOH) on top layers of iron surface (Cornell and Schewertmann, 1996).
Goethite (α-FeOOH) formation predominanted over lepidocrocite formation at a slow oxidation
rate (Benjamin et al. 1996). A high pH and a slower rate of oxidation favored formation of
magnetite (Fe3O4) over lepidocrocite (Cornell and Schewertmann, 1996). Iron corrosion produces
OH- ions that increase pH and react with dissolved carbonic acid and bicarbonate species in the
groundwater to produce carbonate ions, build up of carbonate ions eventually results in the
precipitation of carbonate solid species, such as calcite (CaCO3) and siderite (FeCO3). Buildup of
precipitates could reduce the reactivity of Fe0 by blocking the surface to further reaction, they
could also reduce water flow rate by blocking the pore spaces between iron particles.
Surface Fe(II) complexes formed with hydrous iron oxides, silicates and sulfides, are very
112
efficient reductants from a thermodynamic as well as from a kinetic point of view (Stumm and
Sulzberger 1992). Surface Fe(II) has been shown to effectively remove a variety of pollutant
including nitrite, nitrate, chromium, selenate, uranium, vanadate, and nitrobenzene from aqueous
solution (Buerge and Hug 1999; Cui and Eriksen 1996; Klausen et al. 1995; Liger et al. 1999;
Myneni et al. 1997; Ottley et al. 1997; Schwertmann and Pfab 1994). The reaction rate is not a
direct function of the total sorbed Fe(II) concentration. Instead, it has been shown by Charlet et al.
(1998b) to be proportional to the concentration of the >FeOFeIIOH0 hydroxylated surface complex
of Fe(II).
Sorption of Fe(II) to Fe-oxyhydroxide surfaces enhances the oxygenation of sorbed Fe(II)
species (Tamura et al., 1980). This reaction can lead to autocatalytic processes. Upon oxidation of
a Fe(II) surface species, Fe-oxyhydroxide precipitates, which serves as a substrate for further Fe(II)
sorption. Therefore, as the reaction proceeds, the heterogeneous precipitation of Fe-oxyhydroxides
leads to the formation of new sites for Fe(II) sorption, and the reaction rate speeds up.
The sorption of Fe(II) has also been proposed to inhibit microbial dissolution of crystalline
iron oxides (Liu et al. 2001; Roden and Urrutia 1999)
5.1.2 Arsenic – iron redox reaction and As removal by ZVI
Corrosion products of ZVI are a mixture of amorphous Fe(III) oxide/hydroxide, magnetite
and/or maghemite and lepidocrocite (γ-FeOOH) (Farrell et al. 2001; Kanel et al. 2005; Manning et
al. 2002; Melitas et al. 2002). The formation of various corrosion products on the surface of ZVI
results in the creation of adsorption sites for both As(III) and As(V). The suggested mechanism of
arsenic removal is the formation of inner-sphere bidentate As(III) and As(V) complexes with iron
corrosion products (Farrell, et al. 2001; Manning, et al. 2002 ).
113
Abiotic As(V) reduction by Fe0 is still debatable. Using X-ray Photoelectron Spectroscopy
(XPS), Su and Puls (2001) found reduction of As(V) to As(III) on Fe0 where a steady distribution
of 73-76% As(V) and 22-25% As(III) was achieved over 30-60 days in the solid phase corrosion
products. Reduction of As(V) to As(III) and As(III) to As(0) by iron fillings have also been
observed in batch experiments conducted in nitrogen purged solutions (Bang, et al. 2005).
Lackovic et al. (2000) found no evidence of As(V) reduction or As(III) oxidation in leachates from
As(V)- and As(III)- treated Fe0:sand column experiments over shorter time periods. Recent work
by Melitas et al. (2002) has proposed that reduction of As(V) adsorbed on ZVI is more favorable
than As(V) in solution, they also concluded that the electrochemical potential required to reduce
As(V) to As(III) is lower than the potential produced at the corroding Fe0 surface in aqueous
solution, therefore, in the absence of biological reduction, there will be little conversion of As(V)
to As(III) in zero-valent iron filter media.
Relatively little work has been done on the As(V)–Fe(II) reaction. Charlet et al. (2002)
showed that Fe(II) adsorbed to a clay surface was oxidized by As(V), producing a Fe(III) coating
along surface defects. Johnston et al. (2007) observed the reduction of As(V) to As(III) with the
presence of goethite and Fe(II), suggesting that an adsorbed Fe(II) species is the active reductant.
Direct homogenous reduction of As(V) by Fe(II) is thermodynamically feasible but is kinetically
limited. In the presence of goethite, the reaction is catalyzed to some degree, but the kinetics are
still relatively slow..
Under aerobic conditions, the Fe0 corrosion did not cause As(V) reduction to As(III) but
did cause As(III) oxidation to As(V). Water reduction and release of OH- to solution on the surface
of corroding Fe0 may also promote As(III) oxidation. A previous study (Devitre et al. 1991) found
114
that iron oxides and oxyhydroxides synthesized from Fe(II) were capable of rapid As(III)
oxidation. Co-oxidation of As(III) was observed during oxygenation of Fe(II), but the relative
extents of As(V) production and Fe(II) consumption were highly dependent on buffer type and
concentration.
Although As redox transformations may be influenced by abiotic reactions,
microorganisms appear to commonly dominate the redox chemistry of As and are capable to
reducing As(V) in solution or adsorbed on the surface of Fe (hydr)oxides (Zobrist et al. 2000).
5.1.3 As release
The principal mechanisms of arsenic mobilization associated with geochemical
conditions have been identified as desorption in alkaline conditions, competitive sorption, and
reductive release, especially as associated with the dissolution of iron oxides. Of these, the
reductive release of arsenic and/or arsenic-bearing minerals from such deposites as iron(III)
(hydr)-oxides, appears to be the primary cause of elevated arsenic levels in groundwater
(Cummings et al. 1999; Nickson et al. 2000; Pfeifer et al. 2004). The dissolution and
transformation of the iron (hydr)oxides will impart a pronounced effect on As partitioning.
Ferrihydrite, a short-range order material common in soils and sediments, is transforming to lower
surface area minerals such as goethite and magnetite in the presence of aqueous Fe(II) (Benner et
al. 2002; Hansel et al. 2003). Thus, iron reduction should be expected to induce As release
(desorption) when Fe(III) (hydr)oxides are dissolved or are transformed to lower surface area
minerals. As(III) binds to Fe(III) (hydr)oxides more extensively than As(V) under circumneutral
conditions (Dixit and Hering 2003), but was contrarily shown to be more mobile under flow
conditions than As(V) (Gulens and Champ 1978). Thus, the reduction of As(V) to As(III) will also
115
cause arsenic release (Jenne, 1979). Arsenic associated with poorly crystalline iron oxides can also
be mobilized as a result of dissimilatory iron reduction by microorganisms (Cummings et al. 1999;
Nickson et al. 2000; Pfeifer et al. 2004; Van Geen et al. 2004; Zobrist et al. 2000).
In drinking water distribution systems, arsenic released could be related to iron based
solids. It was reported that solids released from cast iron pipes could have an arsenic content of 83
ug As/g solid, while hydrant flushed solid contain nearly 2000 ug As/g solid (Lytle et al. 2004).
Those iron oxide solids are loosely deposited at the pipe surface and can become re-suspended by
hydraulic flow.
The objectives of this study was to (i) Study the morphology and structure of the surface
corrosion layer of precorroded steel sheets (ii) investigate how the surface structure of precorroded
steel sheets affect arsenic removal. (iii) Study the arsenic speciation in precorroded steel sheets
and iron – tailored GAC column and explore the redox reaction in the system.
5.2 MATERIALS and METHODS
5.2.1 Precorroded steel sheets.
Perforated steel sheets originated from McMaster.com, which were low carbon plain steel,
with a thickness of 0.5 mm; the steel sheets had holes with a diameter of 0.6 mm, and total
opening area of 23%. The sheets were cut into 0.5~0.6 (±0.2) mm × 0.5~1.2 (±0.2) mm before
use.
The precorrosion of perforated steel sheets was conducted by soaking the steel materials in
acid solution (1 M nitric acid + 8 % oxalic acid) for 1-6 days. After the precorrosion, the steel
pieces were taken out and washed with deionized (DI) water until the washing water pH exceeded
116
5.5. The precorroded steel sheets was then soaked in DI water and stored in a glove bag before it
was tested by Scanning Electron Microscopy (SEM), X-ray Diffraction (XRD) and X-ray
Photoelectron Spectroscopy (XPS) method, these samples are designated as fresh precorroded
steel sheets herein. For a second set of samples, the precorroded steel sheets were soaked in DI
water that was exposed to air for two weeks, and then were stored in a glove bag until analylsis.
These samples have been identified as aged precorroded steel sheets herein. Yet a third set of
samples were collected from the iron column that preceded an iron – tailored GAC bed in arsenic
removal mini – column tests (See Chapter 4). These samples have been identified as PSSs #1 - #7
according to the Runs they have been served. The detailed pre-treatment and column operating
parameters were listed in Table 5.1.
Table 5.1 The pretreatment precorroded steel sheets and columnoperating conditions
Pre-Corrosion
time (days) Pre-Aged
(day) Operation
pH Idle
BVs operated before column
stop
BVs to 10 µg/L As consistent breakthrough
Fresh precorroded steel sheets
3 - - -
- -
Aged precorroded steel sheets
3 6 - -
- -
PSS #1 3 5 6 1 370,000 248,000
PSS #2
6
0 6 3
350,000 215,000
PSS #3 3 1 6 no 150,000 103,500 PSS #4 3 0 6-6.5 1 117,000 70,000 PSS #5 1 2 7.3-7.6 1 70,000 20,000
PSS #6 & #7 1 10 7.3-7.6 1 44,000 1000-1300
5.2.2 Scanning Electron Microscopy and energy-dispersive X-ray spectroscopy
(SEM-EDS) tests.
The morphology of the precorroded steel sheets was examined using an Phillips FEI
117
Quanta 200 scanning electron microscopes (SEM), equipped with an energy-dispersive X-ray
spectroscopy (EDS) (Oxford Instruments INCA X-sight system) at an accelerating voltage of
15 kV. SEM images were collected at a beam potential of 20 kV.
5.2.3 X-ray Diffraction (XRD) Measurements.
The XRD patterns of the precorroded steel sheets samples were obtained on a Scintag X2
theta-theta powder diffractometer equipped with a copper target (Cu Kα1 radiation, λ=1.54059 Å),
and a Si(Li) peltier detector. The equipment was run at 45 kV and 40 mA by step-scanning from
10°to 60°2θ with increments of 0.02° 2θ and a counting time of 0.1 second at each step.
5.2.4 X-ray Diffraction (XRD) Measurements of the powders collected from
PSS surface.
The surface precipitates from the precorroded steel sheet surface were scrapped off, then
grounded in an agate mortar and pestle, and sieved to 200 mesh. The sieved sample was then
mounted into micro cavity powder holders (400 μm in diameter and 200μm in depth) for reflection
mode XRD. The XRD spectrum was collected from Rigaku DMAX RAPID microfocus XRD
using Cu Kα radiation at 50 kV and 40 mA. The diffractions were monitored in the 2θ range of
5–90° at a step of 0.03 degree.
5.2.5 X-ray Photoelectron Spectroscopy (XPS) analysis.
The composition and chemical states of the corrosion films were studied using XPS
(PHI-5300/ESCA ?). The vacuum of the analysis chamber was less than 3×10-7 Pa. All binding
energies have been corrected for sample charging effect with reference to the C 1s line at 284.6
eV.
118
5.2.6 Digestion of precorroded steel sheets for arsenic speciation.
As on precorroded steel sheets that have served in arsenic removal runs was extracted by
1.5 ml 5 N sodium hydroxide for 24 hours. Then 0.75 ml sodium hydroxide extraction solution
was centrifugated followed by neutralization with 0.75 ml 5N hydrochloride. As(V) was measured
using a modified molybdenum-blue method (Johnson 1971; Johnson and Pilson 1972; Murphy
and Riley 1962; Ray and Johnson 1972), in which 0.84 volume of sample was mixed with 0.16
volume of colorimetric reagent. As(III) was oxidized to As(V) using 0.2 mM KIO3 to obtain total
As.
5.3 RESULTS and DISCUSSION
5.3.1 SEM result
The morphology of the precorroded steel sheets has been studied via a scanning electronic
microscopy (SEM) method.
Scanning electron micrographs of a freshly precorroded steel sheet showed a quite “clean”
and uniform surface (Figure 5.1 A), it is mostly composed of dense, compact corrosion film with
small amount of amorphous aggregates of nano-scale crystals scattering on top of the dense layer.
The EDS result indicated the presence of Fe, O and little amount of C.
After aging for 2 weeks, the precorroded steel sheets had developed a rough and
heterogenous surface, indicating that the precorroded steel sheets surface had been corroded
unevenly. The edges and hole area of the steel sheets tended to grow a thick layer of porous
corrosion products, while the remaining area are comparably clean, just like the freshly
precorroded steel sheet. The rusty area of the aged precorroded steel sheets is mainly constituted
119
of coalesced flower shaped aggregates (diameter between 10 and 20 μm) of micro-scale pedal
shaped crystals. For the aged samples, the rusty, porous scale structure were indicative of more
active surface, which is in agreement with the higher arsenic removal in the steel chamber of Runs
PS #6 and PS #7. The loose structure of the rust indicated that it could be easily released; and this
is also in agreement with our previous observation that aged precorroded steel sheets have
released more filterable iron.
After being employed in the iron columns for arsenic removal, the surfaces had become
fully covered by corrosion products, as shown in Figure 5.1 C – 5.1 H. This observation indicated
that further corrosion and precipitation occurred during the arsenic removal process. The EDS
analysis of selected spots indicated the presence of Fe, O, C, Si and Ca on the precorroded steel
sheets surface. The Calcium containing spot (2 μm×2 μm×2 μm) generally exhibited a
higher carbon content (by EDS measurement), inferring the formation of calcium carbonate .
SEM images of the precorroded steel sheets employed in Run PS #6 and PS #7 exhibited a
comparably uniform corrosion surface, it is mostly amorphous structured aggregates composed of
very fine particles. Small crystals may sprinkle on the amorphous surface as shown in Figure 5.1
D. These observations indicated that transformation of the iron phase occurred during the
treatment process，as discussed in section 5.3.3.
Precorroded steel sheets employed in columns that operated at pH 6 exhibited a rougher
and very complex surface structure, with rusty corrosion surfaces at the edges. As shown in Figure
5.1 E and 5.1 F, these rusty areas exhibited a layered scale structure with large amount of
randomly precipitated crystals (mostly in the size range of 5-20 μm) on top of a more compact
corrosion film. Various kinds of crystal structures appeared on those chips. Some are listed in
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Figure 5.2. The EDS analysis indicated that the crystals in Figure 5.2A to Figure 5.2E were all
iron oxide. Figure 5.2 F contains same amount of calcium as the Fe (atomic percentage),
indicating calcium precipites and iron oxides. The less rusty surface of precorroded steel sheets
employed at pH 6 runs (i.e. PS #1 – PS #4) has an amorphous structure with considerable
porosities, as shown in Figure 5.1 G and 5.1 H. The observation of more corrosive surface
and large amount of precipitates with precorroded steel sheets in pH 6 runs indicated that more
extensive corrosion at pH 6, and this helps to explain the higher iron release, especially higher
non-filterable iron release and better arsenic removal at pH 6.
5.3.2 XPS results
XPS survey results indicated that fresh precorroded steel sheets surface contained O, Fe, C
and Si (Figure 5.3), After the precorroded steel sheets were employed for arsenic removal, most of
them collected some nitrogen and calcium. This is in general agreement with the SEM-EDS
results, although EDS didn’t detected any Si on fresh precorroded steel sheets, nor any N on PSS
#1 to #7. This could be because XPS is only able to detect up to 10nm in depth from the surface.
The Fe2p, O 1s peak from XPS survey were presented in Figure 5.4 and 5.5. The
observation of Fe2p 3/2 peak at 711.5 eV and 2p 3/2 satellite at 719.8 eV, indicated the Fe on the
very surface of fresh precorroded steel sheets are mainly FeOOH (Su and Puls 2004). This result
is in agreement with the O1s peak at 531.5 ev, which is characteristic of hydroxide
(http://srdata.nist.gov/xps). The O 1s region typically exhibited two resolvable peaks that
correspond to metal oxide, predominately iron oxides (529.5 – 530.3 eV), FeOOH (531 ~ 531.5
eV) and carbonate at 530.5~531 eV (Su and Puls 2004). The main hydrocarbon peak at 284.6 eV
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corresponds to C-C, C-H, and C=C type bonding (http://srdata.nist.gov/xps), shoulder observed at
286 and 288-289 ev are associated with C-O, C=O, HOC=O and carbonate
(http://srdata.nist.gov/xps).
The Fe2p 3/2 peak from all the precorroded steel sheets that served for arsenic removal
shifted slightly to lower energy, indicated that some reduction might have happened in the iron
columns. Run PS #1 and PS #2 are operated in similar condition, sample PSSs #1 are particles that
detached from the PSS surface during the arsenic treatment process. Sample PSSs #2 are powders
ultrasounded from the corroded iron sheet surface. The Fe and O are bonded at slightly higher
energy state with sample PSSs #1, which may indicate that the particles detached from
precorroded steel sheets surface are more likely to be in the Fe(III) hydroxide form, while those
remained on the precorroded steel sheets are probably iron oxide such as hematite or magnetite.
This is in agreement with Sarin’s (2001) result, the very surface of the corroded iron scale are very
loosely attached iron oxide or hydroxide, which may be easily peeled off the scale by change of
hydrolic station (Sarin et al. 2001).
A major peak at 347.1 ~ 347.5 eV was observed for Ca 2p spectrum, revealed that PSS #1,
#4 and #6 and 7 has precipitation of CaCO3 (http://srdata.nist.gov/xps). Si 2p3/2 spectrum for all
samples has a major peak range from 101.7 to 102.4 eV, that’s in good agreement with silicate
(http://srdata.nist.gov/xps). N1s spectrum showed a peak at 400 eV, which is the binding energy of
N2, indicated that the precorroded steel sheets has adsorbed some N2 while stored in the glove bag.
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Table 5.2 Quatitative analysis of precorroded steel sheets – atomic percentage of each element
Name Fresh PSSs
PSSs #1a
PSSs #2b
PSSs #3
PSSs #4
PSSs #6
PSSs #7
O 1s 55.19 48.82 50.56 60.08 56.24 55.16 53.63Fe 2p 12.92 13.67 12.57 12.63 11.06 11.33 12.57C 1s 11.65 19.73 16.81 21.71 28.08 28.61 27.87Si 2p 20.24 15.48 18.51 4.45 1.73 2.74 5.94Ca 2p - 0.15 1.4 1.13 2.37 1.88 - N 1s - 2.15 0.15 - 0.51 0.28 -
a— The Sample #1 tested here are powders from iron particles which had deteached from the PSSs #1.
b— The sample #2 tested here are ultrasounded from the PSSs #2
The quantitative analyses of all the elements detected were presented in Table 5.2. The
surface of precorroded steel sheets employed in arsenic removal columns has comparable Fe and
O percentage as fresh precorroded steel sheets, however, surfaces of PSSs #4, #6 and #7 had
higher carbon content than fresh precorroded steel sheets. This observation indicated that
carbonate precipitates such as CaCO3 or carbonate iron (hydr)oxides developed on the PSS
surface during the arsenic treatment process. The high silica amount from fresh precorroded steel
sheets comes from the steel component, while the silica content from PSSs #1 to #7 was silicate
adsorbed from Rutland water. The detached particles has lower Ca content compared to the
precorroded steel sheets (PSS#2, #3, #4, #6, #7), this could be because Ca came from precipitation
of calcium carbonate on the precorroded steel sheets surface where pH rise as a result of iron
corrosion. No arsenic was detected from precorroded steel sheets employed in any column,
indicating the arsenic concentration on the very surface of precorroded steel sheets was very low.
This could be attributed to the competition from silicate in Rutland water considering the high
silica content on steel surface and the fact that Rutland water has a silica concentration of 12.5
mg/L. The removal of competitive silica anions from water before it reached the GAC column
also helps explain the high arsenic removal in the GAC column.
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5.3.3 XRD result
The X-ray diffraction patterns of precorroded steel sheets were listed in Figure 5.6 and
Figure 5.7. The results revealed that the diffraction peaks of fresh precorroded steel sheets are
attributed to magnetite (Fe3O4), lepidocrocite (γ-FeOOH), Humboldtine (FeC2O4(H2O)2) and
Clinoferrosilite (FeSiO3). The presence of silica phase crystals is in agreement with the high silica
content detected by XPS measurement. Aged PSS showed quite different X-ray diffraction
patterns compared to fresh precorroded steel sheets. The Fe(II) phase, i.e, Humboldtine
(FeC2O4(H2O)2) and Clinoferrosilite (Fe1.5Mg0.5Si2O6), were attenuated whereas peaks of Fe(III)
phases, goethite (α-FeOOH) appeared and dominated, indicate that the major components of rust
on aged precorroded steel sheets surface were goethite (α-FeOOH), magnetite (Fe3O4) and
lepidocrocite (γ-FeOOH). The goethite and lepidocrocite corresponded to the high filterable iron
release observed during the operation of Run PS #6 and PS #7. The transformation from
lepidocrocite to goethite by aging is in accordance with the observation of Cornell and
Schwertmann (1996), who had reported that rapid oxidation favors formation of lepidocrocite
while slow oxidation forms goethite.
No elemental iron phase was detected from either the fresh precorroded steel sheets or
aged precorroded steel sheets; indicating the formation of corrosion layer throughout the steel
surface, the further corrosion of fresh/aged precorroded steel sheets, the iron release from
precorroded steel sheets, and the arsenic adsorption on precorroded steel sheets are all highly
related to the corrosive layers.
After precorroded steel sheets were employed in arsenic removal columns, the strong
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peaks of Humboldtine (FeC2O4(H2O)2), Clinoferrosilite (Fe1.5Mg0.5Si2O6), goethite (α-FeOOH),
magnetite (Fe3O4) and lepidocrocite (γ-FeOOH) that had appeared on fresh precorroded steel
sheets or aged precorroded steel sheets were attenuated, the iron surfaces became amorphous and
no major crystal phase can be discerned (Figure 5.7). Several reasons may be attributed to this fact.
First, the loosely attached iron precipitates from aged column detached from the precorroded steel
sheets surface. Secondly, as the groundwater pass through the precorroded steel sheets, previous
iron phase got dissolved and new phases such as amorphous ferric hydroxide and calcium
carbonate formed. Moreover, a phase transformation might have happened in the iron column,
especially during the idling of the column, as discussed in our previous work.
Its hard for us to tell the iron phase from Figure 5.7, which may because of low content of
each phase in the steel sheets. So in the next step, we collected the powder from the steel surface
and conducted the XRD test again, results showed that the major component of PSSs #2 and PSSs
#3 surfaces are magnetite and goethite (Figure 5.8), PSSs #2 also contain some wustite (FeO). The
peaks of magnetite and goethite are stronger with PSSs #3; these observations may indicate that
reduction reaction happen during the operation of Run PS #2.
Figure 5.8 and 5.9 revealed that the surface layers of precorroded steel sheets after
serving in columns are mainly composed of goethite and magnetite, and some calcium carbonate
phase was also detected. EDX test was conducted with these samples to check elemental
composition so as to help identify the XRD phase, results indicated the existence of Fe, O, Ca and
also some Si. Si containing phases are not as good a match to the XRD patterns, indicated that Si
containing phase has a more amorphous nature. Humboltine, an Fe oxalate phase was also
identified with PSSs #3, this phase comes from the procorrosion of the steel sheets, indicated that
125
less precipitates developed on PSS #3.
We had mentioned that goethite (α-FeOOH), magnetite (Fe3O4) and lepidocrocite
(γ-FeOOH) are the major iron phases on aged precorroded steel sheets, while after serving in
columns, the major components on PSS#6 and PSS#7 are Magnetite, Calcite and Goethite. Fe(III)
phase dominated on PSS surface before and after the column operation. This observation,
combined with the high filterable iron release data (chapter 4), indicated that Fe(III) particles had
entered and accumulated in the GAC column.
5.3.4 Arsenic extraction from precorroded steel sheets in iron column
The arsenic extraction result indicated that As(V) reduction happened in columns that had
been idled and operated at pH below 7 (i.e. Run PS #1 and #4). The Rutland water had an
As(V)/As(III) ratio of 3:1. The As(V)/As(III) ratio on PSSs #1 and #4 are < 0.01, indicating that
all As(V) had been reduced. For PS #6 and #7, which was operated at pH 7.5 and had been idled
for 7 days, the As(V)/As(III) ratio was 1.2, indicating that As(V) was partly reduced. In
comparison, for column that operated continuously at pH 6 (PS #3), As(V)/As(III) ratio was 4, no
As(V) had been reduced, rather, some As(III) had been oxidized to As(V).
All of these observation indicated that idling the column had created a reducing
environment. This is in agreement with the observation of more reduced iron oxide phase in PSSs
#2, which had been idled for 3 times. The reduction reaction reactivated the passivated steel
surface, renewed iron corrosion and thus extended the system’s bed lives.
5.3.5 XAFS result
After the columns stopped, the GAC was taken out and stored in glove bag before they are
126
analyzed by the XAFS method. The arsenic edge of selected GAC were listed in Figure 5.10 ~
5.13. All samples have exhibited both As(V) and As(III) peaks. The As(V)/As(III) ratio for all the
GAC samples ranges from around 0 to approximately 2, while the ratio for Rutland water is 3,
indicating that As(V) reduction could happen in the GAC column.
The Fe edge results (Figure 5.14) of PS#6 and PS#7 GAC indicated that iron in GAC are
mostly wustite and green rust. In section 5.3.4, we had mentioned that Fe(III) particles should
have accumulated in the GAC column. So reduction of Fe(III) must have occurred during the
operation of PS#6 and PS#7.
The reduction of Fe(III) phase dissolved Fe(III) containing particles, arsenic that complexed
or adsorbed on the particles could be released to the effluent water during this process. When large
Fe(III) particles broke into smaller particles, they could move inside GAC pores, leaving more
space available for capturing fresh Fe(III) particles and arsenic contaminant. Thus, the arsenic
removal capacity of GAC column was recovered by the idling process.
5.4 Conclusions
SEM, XRD and XPS studies of the surface layer on PSS indicated that the aged PSS have a
porous and layered scale structure. Amorphouse iron hydroxide comprised of the very top layer,
and they are loosely attached and easily peel off, releasing considerable high iron during column
operation. The precorroded steel sheets that employed in steel chambers showed that different
surface morphology formed at pH 6 and pH 7.5, indicated that more corrosion occurred in
columns operated at pH 6.
127
XRD results and XAFS edge tests results revealed that reduction reaction occurred in both
the steel chamber and the GAC columns for columns that idled. By idling, the passivated PSS
surface was reactivated and the arsenic removal capacity of GAC column was recovered.
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130
C
A B
D
E F
10 μm
50 μm 10 μm
5 μm 20 μm
5 μm
131
Figure 5.1 SEM of precorroded steel sheets (A) Fresh precorroded steel sheets – clean surface (B) Aged precorroded steel sheets – rough and rusty (C) Steel surface in PS # 6 (pH 7.5, idle once) – amorphous and uniform. (D) Steel surface in PS # 7(pH 7.5, idle once) – amorphous and uniform. (E) Steel surface in PS # 4 (pH 6-6.5, idle once) – rough with lots of precipitates. (F) Steel surface in PS #2 (pH 6, idle 3 times) – rough with lots of precipitates. (G) Steel surface in PS #2 (pH 6, idle 3 times) – porous (H) Steel surface in PS #4 (pH 6-6.5, idle once) – porous
20 μm 20 μm
HG
132
Figure 5.2 Various crystals on surfaces of PSSs # 2 and PSSs #4. (A) to (E) Iron oxides, (F) Calcium oxides and iron oxides.
10 μm 50 μm
20 μm 5 μm
5 μm 10 μm
C
A B
D
E F
133
Figure 5.3 Elements identification on precorroded steel sheets by XPS survey. Top spectrum– iron particles ultrasounded from PSS #2 (pH 6, idle three times); 2nd spectrum–iron particles detached from PSS #1 (pH 6, idle once); 3rd spectrum – fresh precorroded steel sheets; 4th spectrum – PSS #3 (pH 6, no idle); 5th spectrum – PSS #4 (pH 6 ~ 6.5, idle once).
Figure 5.4 XPS survey of Fe 2p peak. Note that Fe are FeOOH or iron oxide (Fe2O3 & Fe3O4). Top spectrum– iron particles ultrasounded from PSS #2 (pH 6, idle three times); 2nd spectrum–iron particles detached from PSS #1 (pH 6, idle once);3rd spectrum – fresh precorroded steel sheets; 4th spectrum – PSS #3 (pH 6, no idle); 5th spectrum – PSS #4 (pH 6 ~ 6.5, idle once).
2004006008001000 1200
2
4
6
8
10
12
14 ×104
Binding Energy (eV)Fe2p
O1s
C1s C
a2pN
1s
Si2s
Cou
nts p
er S
econ
d
134
Figure 5.5 XPS survey of O 1s peak. Note that O is mainly hydroxide or iron oxide. 1st spectrum– iron particles ultrasounded from PSS #2 (pH 6, idle three times); 2nd spectrum–iron particles detached from PSS #1 (pH 6, idle once);3rd spectrum – fresh precorroded steel sheets; 4th spectrum – PSS #3 (pH 6, no idle); 5th spectrum – PSS #4 (pH 6 ~ 6.5, idle once).
Figure 5.6 Typical X-ray diffraction patterns of fresh precorroded steel sheets and aged precorroded steel sheet. Peak designations: G = Goethite (α-FeOOH); L = Lepidocrocite (γ-FeOOH); M = Magnitite (Fe3O4); Cl = Clinoferrosilite (Fe1.5Mg0.5Si2O6); H = Humboltine (FeC2O4(H2O)2).
135
Figure 5.7 XRD result of the corroded steel sheets after serving the columns. Sample from top to bottom are PSSs #1, #2 and #6, all samples showed amorphous structure, no major phase could be discerned.
0
5
10
15
20
25
30
35
40
45
50
10 20 30 40 50 60 70
2 theta
M
M
M M
WW
G GG
M
G M
Figure 5.8 X-ray diffraction patterns of the powdered rust collected from steel chamber after runs PS #3 (no idling-top pattern); PS #2 ( thrice-idled-bottom pattern). Peak designations: G = Goethite-α-FeOOH; M = Magnitite Fe3O4; W = Wustite FeO; H = Humboltine (hydrous ferrous oxalate).
PSSs #1
PSSs #2
PSSs# 6
136
0
20
40
60
80
100
120
140
160
15 25 35 45 55 65
Two - theta (degree)
MM MM
C CC CC
G G G GH
HH
Figure 5.9 X-ray diffraction pattern of powders collected from PSSs #4 (pH 6-6.5, idle once), PSSs #3 (pH 6, no idle), PSSs #1 (pH 6, idle once), PSSs #7 (pH 7.5, idle once) and PSSs #6 (pH 6, idle once). Samples are in series from top to bottom. G = Goethite-α-FeOOH; M = Magnitite Fe3O4; W = Wustite FeO; H = Humboltine (hydrous ferrous oxalate).
137
Figure 5.10 Arsenic edges of GAC collected from Run PS#1 (pH 6 & idle once) after column stopped. Note that As(V)/As(III) ratio dropped from 3 (Rutland water) to 0.43 indicating As(V) reduction occurred.
0
0.5
1
1.5
2
2.5
11860 11870 11880 11890E (ev)
μ(E)
As(V)
As(III)
Figure 5.11 Arsenic edges of GAC collected from Run PS#2 (pH 6 & idle three times) after column stopped. Note that As(V)/As(III) is less than 1 indicating As(V) reduction occurred.
138
Figure 5.12 Arsenic edges of GAC collected from Run PS#4 (pH 6 ~6.5 & idle once) after column stopped. Note that As(V)/As(III) ratio dropped from 3 (Rutland water) to 2 indicating As(V) reduction occurred.
Figure 5.13 Arsenic edges of GAC collected from Run PS#5 (pH 7.5 & idle once) after column stopped. Note that As(V)/As(III) ratio dropped from 3 (Rutland water) to 1.3 indicating As(V) reduction occurred.
As2O3 standard
As2O5 standard
Linear combination
As edge
139
0
0.2
0.4
0.6
0.8
1
1.2
1.4
1.6
1.8
7110 7120 7130 7140 7150 7160 7170
E (ev)
u(E)
hematite ferrihydrite
goethite green rust
magnatite PS #5 GAC
PS #6 GAC FeO
Figure 5.14 Fe edge from XAFS result of GAC from Run PS #5 and PS #6, both column were operated at pH 7.5 and idle once. Note that iron are best fit with FeO and green rust, indicating Fe(III) reduction occurred.
Vita: Jiying Zou
EDUCATION Ph.D. Environmental Engineering (expected 05/2009) The Pennsylvania State University, University Park, PA Thesis: “Arsenic removal from groundwater with iron tailored granular activated carbon preceded by pre-corroded steel” Advisor: Fred S. Cannon M.S. Environmental Engineering (7/2002) Peking University, Beijing, China Thesis: “The sustainable development of Guangzhou City – The application of industry ecology to the layout of population, industry, and natural resources.” Advisor: Zhifeng Mao B.S. Textile Engineering (7/1999) Qingdao University, Qingdao, China
PUBLICATIONS AND PRESENTATIONS
Chen WF, Parette R, Zou JY, et al. Arsenic removal by iron-modified activated carbon. WATER RESEARCH 41 (9): 1851-1858 MAY 2007 Zou, Y. and Fred. S. Cannon (2005) Arsenic removal from groundwater with iron tailored activated carbon. AWWA annual conference 2005, SanFransisco, CA, June 12 – 16.
