ASSESSMENT OF THE EFFECTS OF SOIL AMENDMENTS ON THE LEACHING  
 
OF LEAD AND ARSENIC FROM CONTAMINATED SOIL  
 
 
 
 
 
 
 
Except where reference is made to the work of others, the work described in this thesis is 
my own or was done in collaboration with my advisory committee.  This thesis does not 
include proprietary or classified information. 
 
 
 
 
 
___________________________________________ 
Yu Wang 
 
 
 
 
 
Certificate of Approval: 
 
 
 
______________________________ ______________________________ 
Dongye Zhao         Mark O. Barnett, Chair 
Associate Professor     Associate Professor 
Civil Engineering     Civil Engineering 
 
 
 
_______________________________  ______________________________ 
T. Prabhakar Clement     Stephen L. McFarland 
Associate Professor     Acting Dean    
Civil Engineering     Graduate School
 
ASSESSMENT OF THE EFFECTS OF SOIL AMENDMENTS ON THE LEACHING  
 
OF LEAD AND ARSENIC FROM CONTAMINATED SOIL  
 
 
Yu Wang 
 
 
 
A Thesis 
Submitted to 
the Graduate Faculty of 
Auburn University 
in Partial Fulfillment of the  
Requirements for the 
Degree of 
Master of Science 
 
 
Auburn, AL 
December 15, 2006
iii 
ASSESSMENT OF THE EFFECTS OF SOIL AMENDMENTS ON THE LEACHING  
 
OF LEAD AND ARSENIC FROM CONTAMINATED SOIL  
 
 
 
 
Yu Wang 
 
 
Permission is granted to Auburn University to make copies of this thesis at its discretion, 
upon request of individuals or institutions at their expense.  The author reserves all 
publication rights. 
 
 
 
       ______________________________ 
       Signature of Author 
        
 
       ______________________________ 
       Date of Graduation 
 
 
 
 
iv 
 
 
 
 
 
THESIS ABSTRACT 
 
ASSESSMENT OF THE EFFECTS OF SOIL AMENDMENTS ON THE LEACHING  
 
OF LEAD AND ARSENIC FROM CONTAMINATED SOIL  
 
 
 
 
Yu Wang 
Master of Science, December 15, 2006 
(B.S., Nanjing University, 1999) 
 
 
99 Typed Pages 
 
Directed by Mark O. Barnett 
 
 
Lead has a long history in the life of human beings. Unfortunately, only recently did 
people start to realize that lead is a toxin. Because of increasing human activity, soil lead 
pollution has become a serious environmental problem. Phosphorus is an important 
amendment for the in situ immobilization of lead in contaminated soils. To evaluate the 
effect of phosphorous on the bioaccessibility of lead in contaminated soil, a 
Physiologically Based Extraction Test (PBET) was used in this thesis research. 
Three kinds of lead contaminated solids were used in the in situ immobilization test. In 
the in situ process, enough phosphate was added to form chloropyromorphite (the least 
soluble lead-phosphate salt). In the PBET process, chloropyromorphite was oversaturated 
in the PBET supernatant for lead-contaminated soil, potentially indicating that it could 
v 
precipitate from the PBET solution itself, resulting in an experimental artifact. To better 
understand the lead reactions occurring in the process, the formation of 
chloropyromorphite was studied in HCl- and HNO
3
- based PBET solution. In the 
presence of phosphate, lead was not efficiently removed from the PBET solution under 
either homogeneous or heterogeneous conditions. In homogeneous chloropyromorphite 
formation tests under different conditions, the lead solubility in the PBET solution was 
generally greater than predicted for chloropyromorphite. The results indicate that the 
PBET is an acceptable method to evaluate the bioaccessibility of lead contaminated soils 
that are treated by phosphate amendments, as almost no lead precipitation occurred in the 
PBET process. 
Arsenic (As) is a toxic metalloid that has caused widespread soil contamination 
problems. While various iron-based amendments have been studied for immobilizing As 
in contaminated soils, the feasibility of stabilized iron-based nanoparticles has not been 
reported. This study investigated the effectiveness of using three types of iron-based 
nanoparticles, including zero-valent iron (ZVI), ferrous sulfide (FeS), and magnetite 
(Fe
3
O
4
), for the immobilization of arsenic in two kind contaminated soils.  
Different Fe/As molar ratio (5:1 ~100:1) and treatment times (3 and 7 days) were 
conducted to evaluate the effect of these nanoparticles. Higher amounts of iron particles 
were more effective in reducing As bioaccessibility and leachability. In addition, 
magnetite nanoparticles generally worked a little better than FeS and ZVI nanoparticles 
in reducing the bioaccessibility. For two different types of soils, treatment was more 
effective on the soil with the lowest iron content and highest As concentration. These 
results suggest that the immobilization of As and reduction of As bioaccessibility in soil 
vi 
by adding iron-base nanoscale particles maybe an effective method to remediate As 
contaminated soils. 
 
vii 
ACKNOWLEDGEMENTS 
The author would like to acknowledge the excellent instruction, support, and 
guidance of his advisor, Dr. Mark O. Barnett and his co-advisor Dr. Dongye Zhao. 
Additional thanks to Dr. Prabhakar Clement for his instruction, knowledge, and support 
during the author?s term at Auburn University.  The author would like to thank Jinling 
Zhuang for technical assistance in the environmental laboratory at Auburn University. 
Thanks to Dr. Daniel E. Giammar in the Civil and Environmental Engineering 
Department at Washington University at St. Louis for the instruction during the author?s 
visiting term. Thanks also to Thomas Albrecht-Schmitt in the Chemistry Department at 
Auburn University for the powder X-ray diffractometer measurements.  The author 
would also like to thank his loving wife Jie Gao for her support and understanding, as 
well as his parents, Shusen Wang and Yingli Wan for their encouragement during the 
author?s term at Auburn University. 
viii 
Style manual or journal used  Guide to Preparation and Submission of Theses and 
Dissertations                                _ 
Computer software used: 
Microsoft Office XP: Excel, PowerPoint, Word,  
Endnote 7.0 
Visual MINTEQ 2.40 
 
ix 
TABLE OF CONTENTS 
LIST OF TABLES?????????????????????????.....  xi
LIST OF FIGURES????????????????????????..?. xii
CHAPTER I.  INTRODUCTION???????????????????.... 1
          1.1 Lead and Human Beings???????????????...???..... 1
          1.2 Lead in the Soil Environment??????????.?????..??. 2
          1.3 In situ Lead Immobilization with Phosphorus???????.???? 2
          1.4 Chloropyromorphite Formation and Dissolution???????..??... 4
          1.5 Bioaccessibility and Bioavailability of Lead?????????.............. 6
          1.6 Physiologically Based Extraction Test??????????????.. 8
          1.7 Objectives?????????????????...???????? 9
          1.8 Organization????????????????????????? 11
CHAPTER II. ASSESSMENT PBET PROCEDURE FOR PHOSPAHTE 
AMENDMENT LEAD-CONTAMINATED SOIL SAMPLES ???12
          2.1 Introduction???????...????????????????... 12
          2.2 Materials and Methods????????????????????. 14
2.2.1 Materials???..????????.????????????. 14
2.2.2 Amendments??????????????????????. 15
2.2.3 PBET Extractions???????????????????? 16
2.2.4 Additional Wet Chemical Experiments???????????? 16
x 
2.2.5 Heterogenous Chloropyromorphite Formation Test??????.?. 18
2.2.6 Chloropyromorphite Synthesis and Dissolution Test?????.?.. 18
2.2.7 Measurement??????????..????????.???. 18
2.3 Results and Discussions????????????????????. 19
2.3.1 Effects of Soil Moisture on Amendment Addition???................. 19
2.3.2 Lead and Phosphate Uptake in PBET Process?????.???... 22
2.3.3 Homogeneous Chloropyromorphite Formation Test in PBET???. 30
2.3.4 Homogeneous Chloropyromorphite Formation Test in Different 
Condition???????????????????????? 33
2.3.5 Heterogeneous Chloropyromorphite Formation????????... 38
2.3.6 Chloropyromorphite Dissolution????????????..?? 41
2.3.7 Conclusion???????????????????????. 43
CHAPTER III. IN SITU SOIL IMMOBILIZATION OF ARSENIC BY 
NANOSCALE ZERO-VALENT IRON, FERROUS SULFIDE 
(FeS), MAGNETITE (Fe
3
O
4
) PARTICLES?????????. 44
3. 1 Introduction??????????????????????.??. 44
3.2 Materials and Method???..?????????????????. 45
3.2.1 Materials???????????????????????.. 47
3.2.2 Preparation of Nanoparticles????????????????. 47
3.2.3 Treatments for As-Contaminated Soils????????????.. 49
3.2.4 TCLP and PBET Extraction Test????????????.?? 49
3.2.5 Analytical Method???????????????????? 50
3.3 Results and Discussion????????????????????.. 51
3.3.1 Treatment Effects on Soil Chemistry?????????????. 51
xi 
3.3.2 Reduction of PBET Bioaccessibility????????????? 57
3.3.3 Reduction of TCLP Leachability???..??????????? 58
3.3.4 Effect of Treatment Time????????????????? 63
3.3.5 Mechanisms of Arsenic Immobilization by Iron-based Nanoparticles. 63
3. 4 Conclusion????????????????????????? 65
CHAPTER IV    CONCLUSIONS???????????????????... 67
REFERENCES??????????????????.?????????. 70
 
  
xii 
LIST OF TABLES 
Table 1.1.    Summary of Parameters Used by Different PBET ???????... 10
Table 2.1.    Reactions Included in PBET solution Modeling????????? 32
Table 3.1      pH Change for Different Molar ratio Treatments????????.... 52
 
 
 
 
 
 
 
 
 
 
 
xiii 
 
 
 
LIST OF FIGURES 
Figure 1.1  XRD Patterns for Amended Soil. The Curve consisting of Circles 
Represents a 1% Diluted Chloropyromorphite Standard...????. 7
Figure 2.1   Effect of Initial Moisture Content on the Resulting Bioaccessibility of 
Amended and Unamended  Pb-spiked Soil, Pb-contaminated Soil, 
and Pb-spiked Sand?????.?????????????.... 21
Figure 2.2a  Concentration of Dissolved Pb During the PBET Extraction of the Pb-
Contaminated Soil????...????????????.?.. 24
Figure 2.2b  Concentration of Dissolved P During the PBET Extraction of the Pb-
Contaminated Soil?...???????????..????..??. 25
Figure 2.2c   Concentration of Dissolved K During the PBET Extraction of the Pb-
Contaminated Soil??????????????????...? 26
Figure 2.3a   Saturation Indices and Percentage of Added Pb(II) and P in the 
Dissolved Phase as a Function of Aging Time for HCl-based  PBET 
Solutions????????????????????..?...? 28
Figure 2.3b   Saturation Indices and Percentage of Added Pb(II) and P in the 
Dissolved Phase as a Function of Aging Time for HNO
3
-based 
PBET Solutions????.?.............................................................. 29
 
xiv 
Figure 2.4      Percentage of Dissolved Lead Versus Time in Different pH Buffer 
Solutions???????................................................................. 34
Figure 2.5    Percentage of Dissolved Lead Versus Time in Different Glycine 
Solutions???????????????????????.. 35
Figure 2.6a   Percentage of Dissolved Lead Versus Time (7 days) in the Presence 
of Different Glycine Concentrations Solution????????.... 36
Figure 2.6b   Percentage of Dissolved Lead Versus Time (1 hour) in the Presence of 
Different Glycine Concentrations Solution??................................ 37
Figure 2.7     Percentage of Pb in Aqueous Phase Versus Time with Solid 
Phosphate Rock Added to the PBET Solution????????.... 40
Figure 2.8     Pb Release Chloropyromorphite in the PBET as a Function of Time... 42
Figure 3.1     Comparison of As Bioaccessibility (PBET) of WAOS Soil Sample 
by Different Fe/As Ratio Iron Based Nanoparticles Treatment..?.... 53
Figure 3.2     Comparison of As Bioaccessibility (PBET) of As-spiked Soil Sample 
by Different Fe/As Ratio Iron Based Nanoparticles Treatment..?? 54
Figure 3.3      Comparison of As TCLP leachability of WAOS Soil Samples by 
Series Fe/As Ratio Iron Based Nanoparticles Treatment????.. 55
Figure 3.4      Comparison of As TCLP leachability of As-spiked Soil Samples by 
Series Fe/As Ratio Iron Based Nanoparticles Treatment??..?? 56
Figure 3.5     Comparison of As Bioaccessibility of WAOS Soil Samples for 3 and 
7 days Treatment by Different Iron Based Nanoparticles????.. 59
 
xv 
Figure 3.6     Comparison of As Leachability of WAOS Soil Samples for 3 and 7 
days Treatment by Different Iron Based Nanoparticles ??...?? 60
Figure 3.7     Comparison of As Bioaccessibility of As-spiked Soil Samples for 3 
and 7 days Treatment by Different Iron Based Nanoparticles??.... 61
Figure 3.8     Comparison of As Leachability of As-spiked Soil Samples for 3 and 
7 days Treatment by Different Iron Based Nanoparticles................... 62
 
  
 
1 
 
 
 
CHAPTER I 
INTRODUCTION 
 
1.1 Lead and Human Beings 
Lead is the heaviest member of the carbon family (Group  ) with a molecular weight 
of 207.20. Lead has such a long history in the life of human beings that its first use can 
be dated back to ancient Egypt (3400 B.C.) [1]. Unfortunately, only recently did people 
start to realize that lead is a neurotoxin that can harm the brains and nervous systems of 
young children temporarily or permanently [2]. High amounts of lead ingestion are also 
fatal for adult [2]. Some researchers even attributed the fall of Rome to lead poisoning 
[2]. Nowadays lead is being used in many industrial products, such as lead-acid batteries 
used by cars, lead solder which is used to seal cans, and as a metal alloy which is used 
for military (e. g. bullets and shells) [2]. Because of its multiple use, lead has many 
pathways to enter human body and damage human health. Generally, people can get lead 
in their body under the following three circumstances: ?1) ingest objects covered with 
lead; 2) ingest paint chips or soils that contain lead; 3) breathe in lead dust? [3].  
 
1.2 Lead in the Soil Environment 
The abundance of lead in the Earth's crust ranges from 0.0013% to 0.002% [2]. 
Almost no pure lead element exists on the earth [4], and all pure lead came from human 
 
2 
activities. In nature, galena (PbS) is the most common ore of lead minerals [5]. Other 
solid forms of lead are anglesite, or lead sulfate (PbSO
4
); cerussite, or lead carbonate 
(PbCO
3
); and mimetite (PbCl
2
? Pb
3
(AsO
4
)
2
) [4].  
Lead is an reactive metal and can form many compounds with other elements. The 
increasing impacts of human activities have caused large quantities of lead to be released 
into the environment, especially into the soil environment. After entering soil, lead will 
then cause soil pollution. Different target criteria have been set to define soil pollution. 
The lead limit was established at 400 ppm (residential) and 750 ppm (industrial) for 
human health criteria; for the ecological receptor criteria, the lead limit was 740 ppm 
(mammalian) and 40.5 ppm (avian) [6]. According to report of Salatas et al., lead was 
the main pollutant at DOD (Department of Defense) sites [6]. Therefore, some data for 
lead soil pollution were published, and different methods to clean lead-contaminated soil 
were studied. Cao et al. reported that lead concentrations in soils in Florida fire ranges 
increased due to the weathering of lead bullets [7, 8].  
There are two ways to remediate lead soil pollution, one is a ?capping strategy?, and the 
other is in situ immobilization.
 
1.3 In situ Lead Immobilization with Phosphorus 
Phosphorus is an important amendment for in situ lead immobilization in 
contaminated soil, and most remediation techniques use phosphate [9-20]. It has been 
suggested that phosphate minerals, such as phosphate rock and apatite, can significantly 
reduce lead bioavailability in lead contaminated soils [4]. For example, Ma et al. studied 
in situ immobilization of lead with apatite (Ca
5
(PO
4
)
3
OH(s)) as an amendment and 
 
3 
showed reduction of lead bioavailability by phosphate amendment [21].  Ruby et al. used 
in situ formation of lead phosphate in soils as a method to immobilize lead [22]. The 
phosphate amendment method was shown to be the most cost-effective option for 
reducing the leaching and bioavailability potential of soil phase lead [23].  
Researcher observed that lead and phosphate could form precipitates very easily and 
quickly in the aqueous phase. In 1932, Jowett and Price measured the solubility of two 
kinds of lead phosphate salts, Pb
3
(PO
4
)
2
 and PbHPO
4
 [24]. Their study demonstrated 
that chloropyromorphite (Pb
5
(PO
4
)
3
Cl(s)) is the main constituent in the formation of 
solids or minerals [24]. Previous research has revealed that chloropyromorphite  
formation and stability of lead are the most important mechanisms [25-27]. Thus, the 
formation of solid Pb-PO
4
 complexes (especially for chloropyromorphite) in lead 
contaminated soils is considered as a final step in many remediation strategies [4]. 
Hettiarachchi and Pierzynski and Cotter-Howells et al. studied lead phosphate 
formation in soils, and proposed that phosphate amendments could induce further 
pyromorphite (Pb
5
(PO
4
)
3
X(s), where X can be Cl
-
, OH
-
, or Br
-
) formation [4, 28]. 
Chloropyromorphite  plays an important role in the immobilization process because it is 
least soluble among all the lead phosphate minerals, which is the most desired 
characteristics for lead immobilization amendment [4]. Direct evidence for the formation 
of pyromorphite in lead-contaminated soils with P amendment has been reported [29, 
30].  
  
 
 
 
4 
1. 4 Chloropyromorphite Formation and Dissolution 
For soil remediation, the chloropyromorphite formation is an important step in lead 
immobilization. Because of the importance of chloropyromorphite in the in situ lead soil 
remediation, most previous studies for chloropyromorphite formation were done under 
heterogeneous conditions [22, 26, 31]. Other solid lead species could also be 
transformed to chloropyromorphite by a dissolution-precipitation mechanism to 
immobilize soluble soil Pb in situ by simply adding phosphate or phosphate salt to the 
soil. Cotter et al. observed chloropyromorphite formation with soil lead and added 
phosphate [28], and Laperche et al. studied the effect of amendment apatite on soil lead 
and chloropyromorphite formation [30, 32]. Zhang and Ryan reported the formation of 
chloropyromorphite from galena in the presence of hydroxyapatite, and attributed this 
result to hydroxyapatite dissolution and consequent phosphate release. [29]. 
This reaction primarily occurs under heterogeneous conditions, though the solubility 
of both Pb and P could affect the effectiveness of P-induced Pb remediation [33]. The 
main factor in the remediation is soil pH. Cao et al. and Yang et al. concluded that low 
pH will lead to rapid chloropyromorphite formation [15, 19]. Zhang et al studied the 
chloropyromorphite formation on the surface of goethite, and showed that 
chloropyromorphite can even be formed from adsorbed lead [34]. 
Odyakov and Zhizhina studied chloropyromorphite formation under homogenous 
conditions by adding Pb
2+
 into H
2
PO
4
-
 and Cl
- 
system, and their results showed that at 
their experimental pH (3.0 - 6.0), the reaction was very quick [35]. They concluded that 
the ratio of H
2
PO
4
-
 and Pb
2+
 in precipitation crystals will change with different aqueous 
reactant ratio. Scheckel et al. conducted similar experiment in soft-drinks (phosphoric 
 
5 
acid) and observed significant lead removal during the first several minutes [36]. XRD 
result and SEM photo of chloropyromorphite precipitation were further used to provide 
direct evidence for chloropyromorphite formation [36, 37]. Figure 1.1 is the XRD 
pattern of standard chloropyromorphite. 
Effects of the presence of different cations and anions on the pyromorphite formation 
were also studied. Anions (NO
3
-
, Cl
-
, F
-
, SO
4
2-
, and CO
3
2-
) can increase the pyromorphite 
formation while cations (Al, Cd, Cu, Fe(II), Ni, and Zn) will inhibit the process [38, 39]. 
Dissolved organic matter can affect chloropyromorphite formation, and this is explained 
by Pb complexation [40]. Organic matter inhibited the formation of chloropyromorphite 
at pH 3 and 4, but showed no effect at pH 5, 6 and 7 [41]. Besides, organic coating 
around the nuclei of chloropyromorphite may inhibit the reaction as well [41]. The study 
done by Martinez et al. indicated that lead will be over-saturated in some aqueous phases 
[20, 42, 43]. 
Previous studies [5, 8, 19, 44-46] observed a quick kinetic formation of 
chloropyromorphite by the reaction of lead and phosphate. In addition, the reverse 
reaction - dissolution of chloropyromorphite, was also studied under different 
conditions. The effect of pH on solubility of chloropyromorphite has been investigated 
[47]. Increasing pH was associated with a decreasing lead uptake, indicating positive 
effect on lead remediation effectiveness. On the other hand, low pH results in no obvious 
remediation effect. However, this result conflicts with some other studies [15, 19], where 
low pH was related to the increase in both formation and dissolution rates of 
chloropyromorphite. The formation of chloropyromorphite or hydroxypyromorphite 
were barely observed in alkaline condition [48]. 
 
6 
1.5 Bioaccessibility and Bioavailability of Lead 
Bioaccessibility and bioavailability are two important aspects of contaminated soils 
and sediments. In 2002, the National Research Council published the report 
?Bioavailability of contaminants in soils and sediments: processes, tools, and 
applications? which defined the word ?bioavailability? [49]. Hettiarachchi and 
Pierzynski referred  bioavailability to ?the portion of a substance or element in soil that 
is available for absorption into living organisms, such as humans, animals, or plants? [4, 
23]. In the review paper written by Ruby et al., bioaccessibility was also defined as ?the 
fraction that is soluble in the gastrointestinal environment and is available for 
adsorption?. [50]. 
In general, there are currently two methods to assess bioavailability and 
bioaccessibility respectively: one is in vivo (conducted with living animals), and the 
other is in vitro (conducted outside the living organism in the laboratory) [4, 23]. The 
well known Physiologically Based Extraction Test (PBET) has been proposed to quickly 
measure bioaccessibility [51]. 
 
7 
 
 
Figure 1.1 XRD Patterns for Amended Soil. The Curve consisting of Circles 
Represents a 1% Diluted Chloropyromorphite Standard. [52] 
 
 
 
8 
 1.6 Physiologically Based Extraction Test  
In early 1990s, the relationship between lead content and lead uptake by human was 
studied, and work was published on the geochemistry of lead in a simulated stomach 
environment [53, 54]. In 1996, Ruby et al. first brought up the idea of ?Physiologically 
Based Extraction Test? (PBET) [51]. It is ?an in vitro test system for predicting the 
bioavailability of metals from a solid matrix and incorporates gastrointestinal tract 
parameters representative of a human (including stomach and small intestinal pH and 
chemistry, soil-to-solution ratio, stomach mixing, and stomach emptying rates)? [51]. 
Basically, the PBET is used to simulate in vivo processes for Pb and As bioavailability 
in order to estimate the changes induced by soil amendments [4, 23]. In that report, Ruby 
et al. [51] specified that the absolute bioavailability of Pb is the fraction of ingested Pb 
that is absorbed into systemic circulation. The idea of PBET excited many 
environmental scientists since the PBET is such a simple chemical procedure that it 
could potentially replace the complicated animal experiments, such as immature swine 
and rat feeding studies. 
After developing the PBET, Ruby et al. studied the relationship between the PBET 
test and traditional animal test. They compared the results of the PBET extraction with 
data from a Sprague-Dawley rat model and found that there is a linear correlation 
between in vitro estimated bioaccessibility and in vivo measured relative bioavailability 
[23, 51]. Pu et al. evaluated the relationship between the rat model and PBET model for 
assessing toxicity from soil for phenanthrene [55]. Oliver et al. observed a significant 
positive relationship between in vitro measurements of Pb in dust and blood Pb level of 
children [56]. Recently, Marschner compared the soil lead in vitro bioaccessibility and in 
 
9 
vivo bioavailability and concluded that there is a correlation between intestine 
bioaccessibility and in vivo bioavailability [57]. 
Several control factors, is very important in PBET, and can affect the extraction 
results, such as pH, temperature, and stomach mixing and emptying rates [60]. Based on 
the Ruby et al.?s PBET extraction method [51], some modified methods were designed 
to simulate the human GI tract by different researchers [55, 58, 59] (Table 1.2). In this 
thesis, we report results from a  modified, streamlined version of the PBET extraction 
[60, 61].  
 
1.7 Objectives 
The objective of this study was to determine whether or not reductions in Pb 
bioaccessibility when adding P amendments to soil were occurring in situ or as an 
experimental artifact in the PBET itself. Amendments (KH
2
PO
4
) were applied in situ to 
lead contaminated soils, and the immobilization of lead in the amendment process was 
monitored. The formation of chloropyromorphite was studied in the PBET process, and 
the lead immobilization rate was evaluated in PBET solution under different conditions. 
In addition, the dissolution of chloropyromorphite was studied to examine the lead 
release rate in PBET solution. 
 
 
10 
Table 1.1 Summary of parameters which used by different PBET [58] 
Parameter PBET 
model [51] 
 
PBET 
[60-62] 
In Vitro 
model [56] 
PBET 
[63] 
PBET 
(used in this 
study) 
Target 
organism 
human human human human human 
pH 1.3, 2.5 and 
4.0 
1.5 1.3 2.2 2.3 
Na
+
 none none none none none 
Cl
- 
none 0.406 M 0.1 M Near 
0.25M 
0.242 M 
Glycine none 0.4 M none 0.4M 0.4 M 
Pepsin 1.0% none 0.1% none none 
Citrate 0.05% none 1.0% none none 
Malate 0.05% none 0.1% none none 
Lactic acid 0.5% none none none none 
Acetic acid 0.5% none none none none 
Fluid solution 40 mL 10 mL 140 mL 35 mL 100 mL 
Amount of 
soil  
0.4g 0.1g 10 g 3.5 g 1g 
Temperature 
37?  37?  37?  37 ?  37?  
Food added none none none none none 
Extraction 
time 
60 mins 60 mins 120 mins 60mins 60 mins 
Soil/solution 
ratio 
1/160 1/100 1/14 1/10 1/100 
 
 
11 
1.8 Organization 
The organization of this thesis follows the guidelines of a publication style thesis 
as outlined in the Guide to Preparation and Submission of Theses and Dissertations 
printed by the Auburn University Graduate School..  The results of the investigation are 
provided in chapter two and chapter three of this thesis, which are prepared as two draft 
manuscripts for journal submission.   
 
12 
 
 
 
CHAPTER  II 
 ASSESSMENT PBET PROCEDURE FOR PHOSPHATE AMENDMENT  
LEAD-CONTAMINATED SOIL SAMPLES 
 
2.1 Introduction 
The ingestion of soils is typically the major human-health risk pathway, especially to 
children?s health, at lead (Pb)-contaminated sites. As such, the in situ immobilization of 
Pb in contaminated soils is a potentially cost-effective way to reduce the risk of Pb-
contaminated soils. Accordingly, in situ soil remediation of lead via phosphate 
amendments to transfer lead to insoluble phosphate minerals has been extensively studied 
[5, 15, 19, 22, 26, 44-46, 64].  
Among the lead phosphate minerals, chloropyromorphite [Pb
5
(PO
4
)
3
Cl(s)] is the least 
soluble and therefore the most important lead phosphate mineral in soil lead remediation 
[4]. Cotter-Howells [28] studied lead phosphate formation in soils, and proposed that 
phosphate amendments could induce further chloropyromorphite formation. Direct 
evidence for the formation of chloropyromorphite in lead contaminated soils with P 
amendment has been reported [30, 31, 65]. From the published literature [31, 33, 34, 42, 
47, 66, 67], all Pb sources and minerals including anglesite [PbSO
4
(s)], cerrusite 
[PbCO
3
(s)] and Pb adsorbed on goethite, can be transformed to chloropyromorphite 
under the right conditions. Odyakov and Zhizhina [35] studied the formation of 
chloropyromorphite in aqueous phase, and they concluded the formation is very quick 
 
13 
and the ratio of H
2
PO
4
-
 and Pb
2+
 in precipitation crystal can change in different aqueous 
reactant ratio.  Scheckel and Ryan [36] studied the formation of chloropyromorphite in 
soft-drink (phosphoric acid) and found most lead removal happened in the first several 
minutes. Other studies [21, 29, 38, 39, 68] showed that the reaction rate of 
chloropyromorphite formation was kinetically rapid from dissolved lead and other 
phosphate sources in a heterogeneous system. 
The method used to assess the bioavailability of lead in soil is an important issue. The 
bioavailability of Pb has been estimated using the physiologically based extraction test 
(PBET) [5, 45, 51], an in vitro leachability test that was designed to mimic the solubility-
limiting conditions in a child?s digestive tract. For lead (Pb), the results of the in vitro 
PBET have been shown to be linearly correlated with results from an in vivo Sprague-
Dawley rat model ( r
2 
= 0.93) [51].  
However, Scheckel et al. [69] warned that the results of a sequential extraction test of 
identical materials were compromised by the rapid formation chloropyromorphite in the 
extraction procedure itself. Scheckel et al. [52] also warned of the possibility of 
experimental artifacts confounding the results of soil amendment experiments when 
assessing their effectiveness with extraction tests like the PBET.  
The objective of this study was to evaluate the potential formation of 
chloropyromorphite in soil as an experimental artifact during the PBET extraction tests. 
Phosphate amendments were added to three different kinds of Pb-containing solids, and 
the concentration of Pb, phosphate and potassium (K
+
, a tracer) were measured during the 
PBET process. The saturation indices of chloropyromorphite and other minerals were 
 
14 
monitored along with Pb removal in the PBET solution and various control solutions as a 
function of time during the process. 
 
2. 2 Methods and Materials 
2. 2. 1 Materials  
All chemicals employed in this study were analytical grade or above and obtained from 
Fisher Scientific (Pittsburgh, PA), and all solutions were prepared with deionized water 
(18 M??cm) from an ion exchange apparatus. Experiments were conducted on a Pb-
contaminated soil collected from the field and on two solid materials spiked with Pb in 
the laboratory.  The Pb-contaminated soil sample was collected from a small-arms firing 
range in east-central Alabama.  An uncontaminated sample of the same type of soil was 
collected within a few hundred meters of the firing range. Both of these soil samples were 
ground and sieved to less than 250 ?m.  The uncontaminated soil sample and a sample of 
pure quartz sand were spiked with a small volume of concentrated Pb(NO
3
)
2
 solution to a 
1:10 g solid per mL of 10
-3
 M CaCl
2
 solution sufficient to achieve a target solid-phase Pb 
concentration of 1000 mg kg
-1
.  A small volume of 1 M NaOH solution was added to the 
soil/sand slurry at the same time as the Pb spike to neutralize the acid added with the Pb 
spike. After 48 hr of mixing, the soil/sand suspensions were centrifuged and the 
supernatant was decanted. The decanted supernatants were filtered through a 0.45-?m 
membrane filter, and the concentration of Pb in the filtrate was measured. After rinsing 
by pure water three times, the soil/sand was then dried. The total Pb recovery for the Pb-
spiked soil and sand were 100?10%. The concentration of Pb in the Pb-contaminated 
 
15 
soil, the Pb-spiked soil, and the Pb-spiked sand were 3900, 700, and 780 mg kg
-1
, 
respectively. 
2. 2. 2 Amendments 
A 2.0 gram sample of each of the three solid materials was weighed and placed in a 20 
mL HDPE bottle. A phosphate amendment [potassium dihydrogen phosphate, 
KH
2
PO
4
(s)] was added to each to achieve a total P content of 0% (unamended control) 
and 5% by mass. The soil and amendments were dry mixed for five minutes. Then 0%, 
30%, 100%, or 500% (e.g., 5 mL water per 1 g solid) deionized water was to each 
material. The samples without free-standing water (0% and 30% moisture) were aged 
immediately for seven days as described below. The samples with free-standing water 
(100% and 500% moisture) were shaken for four hours. After the shaking process, the 
samples with 100% and 500% moisture were centrifuged and the supernatant was 
decanted and filtered using a 0.45 ?m syringe filter. The wet solid samples were then 
placed in an oven at approximately 70 ?C for four hours to evaporate additional moisture 
to achieve a moisture content of 30%.  All samples were then aged in a covered apparatus 
receiving a constant flow of 100% relative humidity air. After seven days, the samples 
were removed from the aging apparatus and air-dried for 24 hours and then oven dried at 
55 ?C for 24 hours. The concentration of lead, phosphate and potassium in the 
supernatant of the 100% and 500% moisture samples were measured as described below. 
 
2. 2. 3 PBET Extractions 
The bioaccessibility of lead in the solid materials was monitored by a modified 
streamlined version  of the PBET method. The PBET extraction solution was made using 
 
16 
a 0.4 M glycine solution adjusted to a pH of 2.3 with concentrated HCl. The solution pH 
was adjusted at a temperature of 37 ? 2 ?C using a pH meter calibrated with buffer 
solutions adjusted to 37 ? 2 ?C. A 0.1 gram soil sample, 1:100 solid to solution ratio, 
which has previously been shown to be consistent with a 1.0 g sample at the same solid to 
solution ratio , was used for running the PBET. During the 1-hour extraction, the water 
temperature in the bath was maintained at body temperature (37?2 ?C). The samples were 
removed at 60 minutes, centrifuged and filtered. Finally Pb, K and P concentration in the 
PBET supernatant were analyzed as described below. A NIST soil sample was run as 
QA/QC control at the same time. 
 
2. 2. 4 Additional Wet Chemical Experiments 
A series of homogeneous (solution phase only) and heterogenous (both solid and 
solution phase) wet chemistry experiments were conducted to better understand the 
dynamics of chloropyromorphite formation and dissolution.  These experiments were 
conducted using a variety of sources and concentration of Pb (II) and P and auxillary 
chemicals (to adjust ionic strength and pH) and the aqueous concentrations of Pb and P 
were monitored over time as described below.  Many of these experiments were 
conducted at room temperature (~22 ?C) rather than body temperature (37 ?C) as in the 
PBET solution so that the results would be more directly comparable to thermodynamic 
calculations, where the majority of the data is tabulated at 25 ?C.  Tests were conducted 
to simulate the aqueous conditions in the PBET test in the absence of a solid phase. 
In addition to the standard PBET chemical solution, 16.1 mM aqueous phosphate and 
241 ?M Pb(II) via the addition of 0.25 mL 676.2 mM KH
2
PO
4
 and 0.25 mL 10.12 mM 
 
17 
Pb(NO
3
)
2 
were added to the PBET solution. These P and Pb concentrations were 
equivalent to the total (dissolved plus solid) amount of phosphate and Pb in 10 mL of 
PBET extraction solution of a 6400 mg/kg Pb-contaminated soil plus a 5% P amendment. 
Because Cl
-
 is a constituent of chloropyromorphite, both HCl-based and HNO
3
- based 
(using HNO
3
 in place of HCl) PBET solutions were made. The glycine concentration was 
maintained at 0.4 M in both, but the final Cl- concentration of HCl-based PBET solution 
was 0.242 M and NO
3
-
 concentration in HNO
3
-based PBET solution was 0.236 M. In 
addition, 0.4, 0.1, 0.01, 0.001 M glycine solution were made with 0.2 M NaCl and 0.3, 
0.2, 0.1, 0.01, 0.001 M glycine with 0.2 M NaCl solution were adjusted pH to 2.3 with 
same method used to make the PBET solution. Also, pH 2.3 (NaCl/HCl), 4.5 
(CH
3
COONa/ CH
3
COOH), 7.0 (Pipe/NaOH) buffer solution with 0.2 M NaCl PBET 
solutions were tested.  
Batch experiments were conducted at 37 ?C (310 K) and room temperature, ~22 ?C 
(295 K). Additional solutions of just 241 ?M Pb (II) and 16.1 mM H
2
PO
4
-
 were added in 
these solutions and DI water, respectively. A white precipitate was observed immediately 
in most solutions except PBET and 0.3, 0.2, 0.1 M glycine (pH=2.3). Each sample was 
shaken for 10 min, 20 min, 30 min, 45 min, 60 min, 2 hour, 4 hour, 8 hour, 12 hour, 24 
hour, 2 day, 3 day, 4 day, 5 day, and 7 day, respectively. After shaking, the samples were 
opened, and an aliquot of the supernatant was withdrawn and immediately filtered with a 
0.45 ?m syringe filter, and the pH of the remaining solution was immediately measured 
with a pH meter and combination electrode. The filtrate was used to measure the aqueous 
chloride, phosphate and lead concentrations. After 1 hour of shaking, 0.1 g sand was 
added into PBET solution as potential precipitation nuclei. Pb, Cl, P concentration were 
 
18 
measured after adding the nuclei. All saturation indexes for chloropyromorphite in 
solution were calculated with Visual MINTEQ [70]. 
 
2. 2. 5 Heterogenous Chloropyromorphite Formation Test 
0.254 g Phosphate Rock (about 19% P, equivalent to 0.0156 M P) was added into 10 
ml PBET solution as P source with 241 ?M Pb concentration, and the sample was shaken 
1 hour at 37?C. The samples were shaken for 1 hour, then centrifuged and filtered and the 
final Pb concentration was measured. 
 
2. 2. 6 Chloropyromorphite Synthesis and Dissolution Test 
Precipitation of various aged samples of chloropyromorphite was carried out in HDPE 
bottle similar to the method described by Scheckel and Ryan [47]. A solution of 0.25 M 
Pb(NO
3
)
2
, 0.15 M KH
2
PO
4
 and 0.1 M NaCl were mixed at room temperature (~22 ?C) 
and aged three days with the pH controlled at 7.0 by additional NaOH while purging with 
nitrogen. The aged material was collected by centrifugation and washed 5 times with DI 
water to remove excess lead, phosphate, and chloride. After washing, the 
chloropyromorphite was air-dried. Dissolution experiments were carried out in batch 
method with 0.1 g chloropyromorphite shaken with 10 mL PBET solution with 0 or 
0.0161 M phosphate. 
2. 2. 7 Measurements 
    Aqueous lead concentrations were measured with a flame Atomic Absorption 
Spectrometer (SpectrAA 220FS, Varian).  Phosphate and chloride concentrations were 
 
19 
measured with an ion chromatograph (Model DX-120, AS14 column, Dionex, 
Sunnyvale, CA).  
 
2. 3 Results and Discussions 
2. 3. 1 Effects of Soil Moisture on Amendment Addition 
Since the amendment we used in this study, potassium dihydrogen phosphate, KH
2
PO
4
 
is relatively soluble in the aqueous phase, it will dissolve and release phosphate ions 
quickly in soil solution . There are a series of steps required for phosphate amendments to 
reduce lead bioaccessibility. First phosphate and/or Pb must dissolve into the aqueous 
phase, the transport media for the reactants.  Secondly, phosphate, Cl, and Pb(II) must 
collide and react (in one step or a series of steps), resulting in the formation of 
chloropyromorphite. Other lead minerals will transfer to chloropyromorphite quickly in 
the presence of aqueous phosphate [32]. In our experiments where KH
2
PO
4
(s) was added 
to dry Pb-contaminated solids in the presence or absence of varying amounts of soil 
moisture, the rate controlling step of phosphate entering the aqueous phase is the water 
content of sample. In experiments with free-standing water (100% and 500%) moisture, 
the water content was sufficient to rapidly dissolve and mix the dissolution products 
throughout the solution.  In experiments without any moisture, the formation of 
chloropyromorphite would necessarily be limited because of the absence of a transport 
media and the resulting lack of contact between Cl, Pb (II), and PO
4
-containing 
molecules and ions.  The amendment with 30% moisture corresponds to an intermediate 
condition of soil moisture condition with no free-standing water [71], typical of field 
amendment applications (e.g., at Joplin, MO, the amendments were hand applied to tilled 
 
20 
soil and then rototilled into the soil ).  Soil moisture experiments were conducted to 
evaluate the effect of water content on reducing bioaccessibility, and whether the 
increasing water content can reduce the bioaccessibility of lead in soil.  
In order to determine the progress of different reactions in the amendment and PBET 
process, concentrations of lead, potassium (as a tracer to monitor amendment dissolution) 
and phosphate were monitored in PBET solution and, where possible, the supernatant of 
the soil solution (100% and 500% moisture only). 
The K measurements in the initial soil solution for the samples with free-standing 
water (supernatant of 100%, 500%) indicated that KH
2
PO
4
 was rapidly dissolved in the 
100% and 500% moisture. After 4 hours, the amount of K released to the soil solution 
ranged from 34.0-89.0% for the 100% moisture sample and from 47-88.1% for 500% 
moisture sample. In contrast, only 6.7 - 19% of the phosphate appeared in the supernatant 
of the 100% and 500% moisture samples respectively.  This phenomena indicates either 
incongruent dissolution (more K than P dissolved), or, more likely, the greater reactivity 
of phosphate with the solid phase compared to K
+
.  In either case, at the end of the 4 hour 
period, there was ample P available to react with the Pb on a stoichiometric basis.  
 
 
21 
 
 
0
10
20
30
40
50
60
70
80
90
100
0 30 100 500
Moisture (%)
B
ioa
cces
sib
ili
ty (
%)
 
Figure 2. 1 Effect of initial moisture content on the resulting bioaccessibility of 
amended (closed symbol) and unamended (control, open symbol) Pb-spiked soil 
(diamonds), Pb-contaminated soil (triangles), and Pb-spiked sand (squares) (T = 37 ?C). 
 
 
22 
Figure 2.1 shows the bioaccessibility of the control (no phosphate added) and amended 
samples after seven days of aging as a function of the initial moisture content.  For all 
samples and all moisture contents, the control (unamended, open symbols) samples 
exhibited higher bioaccessibility than the amended samples, indicating that the addition 
of phosphate to the samples markedly reduced their bioaccessibility. In addition, for the 
unamended samples, for all moisture contents, the bioaccessibility increased in the order 
expected (i.e., weathered soil has lower bioaccessibility than spiked soil and spiked soil 
has lower bioaccessibility than spiked sand).  The bioaccessibility of the unamended Pb-
contaminated soil and Pb-spiked sand was relatively independent of initial soil moisture 
content.  However, the bioaccessibility of the Pb-spiked soil decreased as initial moisture 
content increased, potentially reflecting the impact of additional moisture on the aging of 
soils spiked with labile metals.  In contrast, all three materials at all soil moistures 
exhibited relatively consistent bioaccessibility, indicating that the phosphate source was 
likely controlling the bioaccessibility of Pb for all amended samples, regardless of Pb 
source or moisture added with the amendment. 
 
2. 3. 2 Lead and Phosphate in the PBET  
Since the initial soil moisture content did not markedly affect the resulting lead 
bioaccessibility in the amended samples, we were concerned that the observed reduction 
in Pb bioaccessibility may have been occurring in the PBET solution itself instead of in 
situ [52, 69], at least for the 0% moisture sample. To further understand the dynamics 
occurring in the PBET solution, we monitored the aqueous Pb, P, and K concentrations 
for these samples in PBET supernatant versus time over the nominally one hour 
 
23 
extraction. The concentration profiles for the Pb-contaminated soil are shown in Figure 
2.2 (a, b, c).  For all four initial moisture contents and all three kinds of solid media (the 
concentration profiles for the Pb-spiked soil and Pb-spiked sand were similar and are not 
shown), as the extraction time increased, the concentration of Pb, K, and P increased until 
reaching an approximately constant concentration.  The K concentrations reached an 
approximately constant concentration within ~10 mins.  Mass balance calculations of the 
K decanted in the original soil solution (100% and 500% initial moisture content only) 
and the K in the PBET indicated that all of the original KH
2
PO
4
(s) amendment had been 
completely solubilized within at least the first ten minutes in the PBET solution.   
In contrast, the Pb and P concentrations did not reach a constant concentration until 
approximately 20-30 minutes.  This delay relative to K indicates either slower release 
into the PBET solution or reactions in addition to simple dissolution occurring in the 
PBET.  For the sample that didn?t receive any initial moisture with the KH
2
PO
4
(s), this 
delay in the concentration of P relative to K almost certainly indicates additional 
reactions of P other than simple dissolution. 
Mass balance calculations were performed on aqueous phosphate measurements in 
PBET solution, by subtracting the phosphate concentration in the control samples (no P 
amendment) sample and phosphate removed with the decanted supernatant (for the 100% 
and 500% moisture only). The corrected phosphate percent in PBET of total amendment 
was near 15% for the two soil samples and near 25% for the Pb-spiked sand, indicating 
greater phosphate attraction for the soil than for the sand. 
 
24 
0
0.02
0.04
0.06
0.08
0.1
0.12
0.14
0 102030405060
Time (mins)
Pb concentration in aqueous phase (mM)   .
0%
30%
100%
500%
control
 
Figure 2.2 (a)  Concentration of dissolved Pb during the PBET extraction of the Pb-
contaminated soil (T = 37 ?C).  Unamended (open symbol, 30% moisture) and amended 
(closed symbol) results at various initial moisture contents shown.  
 
 
25 
 
0
0.5
1
1.5
2
2.5
3
3.5
0204060
Time (min)
P concentration in aqueous phase(mM)
.
0%
30%
100%
500%
control
 
Figure 2.2 (b)  Concentration of dissolved P during the PBET extraction of the Pb-
contaminated soil (T = 37 ?C).  Unamended (open symbol, 30% moisture) and amended 
(closed symbol) results at various initial moisture contents shown.   
 
26 
 
0
2
4
6
8
10
12
14
16
18
20
0 102030405060
Time (min)
K concentration in aqueous phase (mM)   .
0% 30% 100%
500% Control
 
Figure 2.2 (c)  Concentration of dissolved K during the PBET extraction of the Pb-
contaminated soil (T = 37 ?C).  Unamended (open symbol, 30% moisture) and amended 
(closed symbol) results at various initial moisture contents shown.  
 
27 
All moisture samples resulted in an initially low Pb concentration followed by a slow 
increase to a steady-state Pb concentration.  The concentration of Pb reached a constant 
value in 30-40 minutes, indicating either slower release than P or additional reactions not 
involving aqueous P after the P concentration had reached steady-state.  The saturation 
indices of lead minerals for all the solid sample and all sampling time (0, 10, 20, 30, 45, 
60 minutes) were calculated with the different Pb, P, Cl concentration by Visual 
MINTEQ.  For the Pb-spiked soil and the Pb-spiked sand, the saturation indices of all Pb- 
and P-containing solids were all well below 0, with the saturation indices of 
chloropyromorphite ranging from -2.99 to -1.45.  It indicates that no potential lead 
mineral precipitation in the PBET solution for these two solids. However, for the Pb-
contaminated soil, the final chloropyromorphite SI for all points were above 0, ranging 
from 0.56 to 1.11.  The SI of all other Pb- and P-containing minerals were all well below 
-2.0.  Thus thermodynamically, no known pure Pb- or P-containing solids could be 
controlling the solubility, and thus the bioaccessibility, of Pb for the Pb-spiked soil and 
sand, and only chloropyromorphite could be controlling the bioaccessibility of the Pb-
contaminated soil. 
 
 
28 
 
HCl-Bace PBET
0
20
40
60
80
100
0 50 100 150 200
Time (hour)
Dissolved %
4
5
6
7
8
9
10
SI for chloropyromorphite
Dissolved Cl
Dissolved PO4
Dissolved Pb
SI for chloropyromorphite
 
Figure 2. 3(a) Saturation indices and percentage of added Pb(II) and P in the dissolved 
phase as a function of aging time for HCl-based  PBET solutions (~22 ?C).  Initial P 
concentration was 0.0161 M, Cl
-
 concentration was 0.242 M, and Pb concentration was 
0.000241 M. 
 
 
29 
 
0
10
20
30
40
50
60
70
80
90
100
0 50 100 150
Time (hour)
Dissolved %   
.
-12
-10
-8
-6
-4
-2
0
SI for hydroxypyromorphite   .
Dissolve PO4 %
Dissolved Pb %
SI for
hydroxypyromorphite
 
Figure 2. 3(b) Saturation indices and percentage of added Pb(II) and P in the dissolved 
phase as a function of aging time for HNO
3
-based PBET solutions (~22 ?C).  Initial P 
concentration was 0.0161 M, Cl
-
 concentration was 0.242 M, and Pb concentration was 
0.000241 M. 
 
30 
2. 3. 3 Homogeneous Chloropyromorphite Formation Test in PBET 
The concentrations of Pb and P in the PBET solutions indicated that, regardless of the 
amount of moisture added initially with the P amendment, the Pb concentration reached a 
constant concentration before the end of the PBET.  However, the wide range of resulting 
SI values did not suggest that the samples had consistently reached equilibrium with 
respect to chloropyromorphite, the thermodynamically stable solid phase.  To better 
understand the potential reactions occurring between lead and phosphate in the PBET, a 
homogenous test (i.e., no solid present) was performed to measure the reaction of lead 
and phosphate in two kinds of PBET solution, both at 37 ?C, the normal PBET 
temperature, and at 25 ?C in an effort to more accurately model the data 
thermodynamically.  In addition, although the PBET protocol specifies the use of HCl as 
a stomach acid simulation, we conducted additional experiments with HNO
3
 rather than 
HCl, as Cl is a potential reactant with Pb and P in the formation of chloropyromorphite in 
the PBET solution. Finally, the experiment at 25 was conducted for up to seven days to 
see where the reactions would end up.  Under the normal PBET conditions (HCl, 1 hour 
extraction, 37 ?C) less than 5% of the Pb was removed from the PBET solution (not 
shown), clearly indicating the homogeneous precipitation of chloropyromorphite from the 
PBET was not occurring, at least not to the extent predicted thermodynamically.  
Similarly very little Pb (<10%, Fig. 2. 3a) was removed from the normal PBET solution 
at 25 ?C, even after seven days.  In addition, replacing the HCl in the PBET solution with 
HNO
3
 resulted in almost exactly the same degree of Pb removal (c. 10%, Fig. 2. 3b), 
even after seven days.  At the end of the experiment shown in Fig. 2. 3a, the calculated SI 
for chloropyromorphite was 6.49, extremely oversaturated with respect to 
 
31 
chloropyromorphite.  In fact, these solutions were initially close to saturation with respect 
to PbHPO
4
(s), with SI values ranging of -0.72 and 0.61 respectively.  Over the seven day 
period, approximately the same degree of Pb removal occurred in both solutions, 
decreasing the SI for PbHPO
4
(s) to -0.76 and -0.66 respectively.  The final SI for 
chloropyromorphite in the HCl-based PBET solution was 6.49, indicating that the 
solution was significantly supersaturated with respect to chloropyromorphite, even after 
seven days.  In an attempt to potentially increase the reaction rate, pure quartz sand was 
added as a precipitation nuclei to some samples after 1 hour aging time, however, no 
difference was observed. It sum, this information clearly indicates the precipitation 
kinetics of this reaction are slow with comparison other chloropyromorphite formation 
test in homogeneous condition [35, 36]. 
The difference between our study and Odyakov and Zhizhina?s [35] is the glycine 
concentration and pH. Prior works studied the complexation reaction of Pb ion and 
glycine [72-74]. Pb ion was easy to form complexes with high concentration glycine at 
low pH. Table 2.1 is the reaction and constant in the PBET solution. So a series of 
homogeneous chloropyromorphite formation experiments with different pH and glycine 
concentrations were conducted to explain the phenomena. 
 
32 
 
 
Table 2.1 Reactions Included in PBET Modeling (25 ?C, I= 0.1, from VM) 
Chemical Reaction Log K 
Chloropyromorphite (Pb
5
(PO
4
)
3
Cl, s) 5Pb
2+
 +3PO
4
3-
 +Cl
-
? Pb
5
(PO
4
)
3
Cl
 
-84.4 
Hydroxylpyromorphite (Pb
5
(PO
4
)
3
OH, s) 
5Pb
2+
 +3PO
4
3-
 +OH
-
? Pb
5
(PO
4
)
3
OH -76.8 
Pb(OH)
2
(s) Pb
2+
 +2OH
-
 ? Pb(OH)
2
  8.15 
Pb
3
(PO
4
)
2
(s) 3Pb
2+
 +2PO
4
3-
 ? Pb
3
(PO
4
)
2
  -43.53 
PbHPO
4
(s) Pb
2+
+ H
2
PO
4
-
?  PbHPO
4
+ H
+
 -23.8 
PbH-(glycine)
2+ 
Pb
2+ 
+H
+
 + glycine
-
?  PbH-
(glycine)
2+
 
11.7 
Pb-(glycine)
+ 
Pb
2+ 
+ glycine
-
? Pb-(glycine)
+
 5.25 
Pb(glycine)
2 
Pb
2+ 
+ 2glycine
-
? Pb-(glycine)
2
 8.35 
Pb(H-glycine)
2
2+
 Pb
2+ 
+2H
+
 + 2glycine
-
? Pb(H-
glycine)
2
2+
 
21.3 
H-glycine H
+
+glycine
-
? H-glycine 9.77 
H
2
-glycine H
+
+H-glycine? H
2
-glycine
+
 12.24 
PbCl
+ 
Pb
2+
+Cl
-
? PbCl
+
 1.67 
PbCl
2
 Pb
2+
+2Cl
-
? PbCl
2
 1.89 
PbCl
3
-
 Pb
2+
+3Cl
-
? PbCl
3
-
 1.91 
PbCl
4
2- 
Pb
2+
+4Cl
-
? PbCl
4
2-
 1.81 
 
  
 
 
33 
2. 3. 4 Homogeneous Chloropyromorphite Formation Test in Different Condition 
Other precipitation tests were performed with different pH (2.3, 4.5 and 7.0) buffer 
solution and different glycine concentration (0.4, 0.3, 0.2, 0.1, 0.01 and 0.001M) with or 
without adjusting pH to 2.3. Precipitation tests were performed in three pH buffers (2.3, 
4.5 and 7.0). An interesting visual observation was noted when the Pb solution was added 
to the solution with P. A white precipitate formed immediately in two kinds of buffer 
solution (pH 4.5 and 7.0), similar with Scheckel et al.?s result in soft-drink, compare to 
that nothing was observed in PBET solution. The pHs in solution were monitored during 
the reaction, and they were kept close to the original value. From the Pb concentration 
versus time curves (Figure 2. 4), it seem that the reaction is very quick and above 90% Pb 
precipitate in the first hour. For different glycine concentration solution but without 
adjusting pH samples, the same white precipitations were observed. The Pb concentration 
dropped quickly (Figure 2. 5) in different glycine concentration solution. The pH of all of 
these solutions, however, remained >3.9. 
 
34 
 
0
20
40
60
80
100
0 1020304050
Time (hour)
Pb in aqueous phase %    
.
-5
0
5
10
15
20
25
30
35
40
SI for Chloropyromorphite    .
pH= 2.3
pH= 4.5
pH= 7.0
SI for Clp(pH2.3)
SI for Clp(pH4.5)
SI for Clp(pH7.0)
 
Figure 2. 4 Percentage of dissolved lead versus time in different pH buffer solutions 
(pH 2.3, 4.5, 7.0; 0.2 M NaCl; 22 ?C). No glycine in the solution. Initial P concentration 
was 0.0161 M, and Pb concentration was 0.000241 M. Total Cl
-
 was 0.2 M.  
 
35 
 
0
20
40
60
80
100
0 0.2 0.4 0.6 0.8 1
Time (hour)
Pb in aqueous phase %       .
0.4M glycine
0.1M glycine
0.01 M glycine
0.001 M glycine
DI water
 
 
 
Figure 2. 5 Percentage of dissolved lead versus time in different glycine solutions (pH 
3.9-6.8; 0.2 M NaCl; 22 ?C). Initial P concentration was 0.0161 M, and Pb concentration 
was 0.000241 M. After 1 hour, the concentration didn?t change appreciably over the next 
7 days. 
 
36 
 
0
20
40
60
80
100
120
140
0 50 100 150
Time (hour)
Pb in aqueous phase %      .
0.4M glycine pH2.3
0.3M glycine pH2.3
0.2M glycine pH2.3
0.1M glycine pH2.3
0.01M glycine pH2.3
0.001M glycine pH2.3
 Figure 2. 6 (a) Percentage of dissolved lead versus time (7 days) in the presence of 
different glycine concentrations solution (T=~22 ?C, pH=2.3). Initial P concentration was 
0.0161 M, Cl
-
 concentration was 0.242 M, and Pb concentration was 0.000241 M.   
 
 
 
37 
 
0
10
20
30
40
50
60
70
80
90
100
0 0.2 0.4 0.6 0.8 1
Time (hour)
Pb in aqueous phase %      
  .
0.4M
0.3M 
0.2M
0.1M 
0.01M
0.001M
 
Figure 2. 6 (b) Percentage of dissolved lead versus time (1 hour) in the presence of 
different glycine concentrations solution (T=~22 ?C, pH=2.3). Initial P concentration was 
0.0161 M, Cl
-
 concentration was 0.242 M, and Pb concentration was 0.000241 M.  . 
 
 
 
38 
The experiments above show that neither low pH, the presence of glycine, nor ionic 
strength along interferes with the bulk precipitation of chloropyromorphite.  However, as 
demonstrated in Figure 2. 6, at pH 2.3, as the concentration of glycine increases, the rate 
of precipitation of chloropyromorphite decreases dramatically.  In the presence of only 
0.001 and 0.01 M glycine, the Pb concentration dropped very rapidly (<15 min) and had 
dropped to less than 5% of its initial concentration within one hour.  However, when the 
glycine concentration was increased to 0.1 M, the white precipitate was not observed 
after just 1 hour, at which time there was still near 70% Pb for 0.1 M and >90% Pb for 
0.2 and 0.3 M in the solution.  
The saturation index of chloropyromorphite in 0.1 M glycine solution before reaction 
was 7.8, and after 7 days reaction the saturation index of chloropyromorphite in 0.1 M 
glycine is 3.95. And the saturation index of chloropyromorphite in 0.4 M glycine solution 
before reaction was 6.7, and after 7 days reaction the saturation index of 
chloropyromorphite in 0.4 M glycine is 6.49. It seems that the high concentration glycine 
will block the precipitation in low pH and keep the solution over-saturated. 
 
2. 3. 5 Heterogeneous Chloropyromorphite Formation  
Rock phosphate was added into Pb-PBET solution as a P source. Figure 2. 7 is the 
aqueous lead percent in PBET solution after adding the rock phosphate. After 1 hour 
aging at 37 ?C, 89.1% Pb was still in aqueous phase, this phenomenon indicate in 
heterogeneous condition, the transformation reaction of lead to solid phase in the present 
of phosphate in PBET environment is slow, only 11% Pb was transfer to solid phase . 
This illuminates that the lead removal reaction (chloropyromorphite formation) is slow in 
 
39 
PBET solution under both homogenous and heterogeneous conditions. Based on the 
results from heterogeneous and homogeneous condition, we propose that the high 
concentration of glycine will block the reaction among Pb
2+
, Cl
- 
and PO
4
3-
, because 
chloropyromorphite formation is a multi-ion compound, glycine can affect the collision 
of ions.  
 
40 
 
0
20
40
60
80
100
0 102030405060
Time (min)
Pb percent in aqueous phase %
 
Figure 2. 7  Percentage of Pb in aqueous phase vs. time with solid phosphate rock added 
to the PBET solution (37 ?C). Total Pb in PBET solution was 0.000241M and total P 
concentration was similar 0.0161M. 
 
41 
2. 3. 6 Chloropyromorphite Dissolution 
Figure 2. 8 show the Pb
2+
 concentration change during the dissolution of 
chloropyromorphite in 1 hour. In the present of P, only 0.3% Pb of chloropyromorphite 
dissolved from solid. This rate of dissolution was slower than Scheckel and Ryan?s result 
[47]. The other hand, the saturation indices of chloropyromorphite of the PBET solution 
during 1 hour were 5.168~5.81 without P and 6.34~7.04 for with P. It seems that 
chloropyromorphite is over-saturated in the PBET solution.  Although Pb solubility 
sometimes does not conform to thermodynamic predictions, this specific observation, the 
apparent supersaturation of chloropyromorphite, has been noted in a number of other 
studies.  For example, adding phosphate and chloride to soil reduced Pb solubility but not 
to the level predicted thermodynamically, suggesting potential inaccuracies in the 
thermodynamic data [41, 75].  Mart?nez et al.  continually observed dissolved Pb
2+
 
activity in suspension one to two orders of magnitude higher than predicted 
thermodynamically, even in samples aged up to three years, which was consistent with 
several other published data sets.  Perhaps most relevantly, Lang and Kaupenjohann  
consistently observed solutions apparently oversaturated with respect to 
chloropyromorphite, particularly at low pH in the presence of organic ligands, which they 
potentially attributed to kinetic limitations, colloids, and crystallization inhibition.  
Whatever the cause, our data clearly do not support the formation of chloropyromorphite 
in PBET solution. 
 
42 
0
0.2
0.4
0.6
0.8
1
1.2
1.4
0204060
Time (min)
Pb in Aqueous phase  (mM)
 ...
3
4
5
6
7
8
9
10
SI for Chloropyromorphite      .
  
Figure 2.8 Pb release chloropyromorphite in the PBET as a function of time (37?C). 0.1 
g chloropyromorphite was added into 10 ml PBET solution with 0.0161M phosphate (?) 
and without phosphate (? ). Saturation indices were calculated by visual MINTEQ with 
P (?) and without P (? ) 
 
43 
2. 3. 7 Conclusion 
In the PBET process for the three amended solids, the concentration of K
+
, Pb
2+
 and 
H
2
PO
4
-
 increased in the one hour PBET time. For both Pb-spiked soil and Pb-spiked 
sand, the saturation indices of lead-phosphate minerals were below 0. This result 
indicated no potential lead mineral precipitation in the PBET solution for these two 
solids. However, for Pb-contaminated soil, due to the high initial soil Pb concentration, 
after 10 minutes in the PBET process, the SI for all points were above 0 (0.56-1.12). 
These numbers indicated that the PBET solutions were oversaturated.  
In the homogeneous chloropyromorphite formation test in PBET solution, under the 
normal PBET conditions (HCl, 1 hour extraction, 37 ?C) less than 5% of the Pb was 
removed from the aqueous phase, clearly indicating that the homogeneous precipitation 
of chloropyromorphite from the PBET was not occurring to any significant extent.  Under 
heterogeneous conditions, after one hour aging at 37 ?C, 89.1% Pb was still in aqueous 
phase. It showed that the transformation reaction of lead to solid phase was very slow in 
the present of phosphate in PBET solution.  
These results illuminate that the lead removal reaction (chloropyromorphite formation) 
is very slow in PBET solution under both homogenous and heterogeneous conditions. 
Based on this conclusion, we speculate that high concentrations of glycine may block the 
reaction among Pb
2+
, Cl
- 
and PO
4
3- 
by affecting the collision of ions since 
chloropyromorphite formation involves multiple ions. This result also indicates that the 
immobilization of lead occurs more under in situ conditions than in the PBET process. 
 
44 
 
 
 
CHAPTER  ?  
IN SITU SOIL IMMOBILIZATION OF ARSENIC BY NANOSCALE ZERO-VALENT 
IRON, FERROUS SULFIDE (FeS), MAGNETITE (Fe
3
O
4
) PARTICLES 
 
3. 1 Introduction 
Arsenic in soils and groundwater results from natural sources (e.g. natural geochemical 
reactions) as well as anthropogenic activities, such as mining, discharges of industrial 
wastes, military activities, and application of agricultural pesticides. Arsenic is ranked the 
second most common inorganic pollutant in the U.S. superfund sites [76]. Arsenic 
contaminated soils, sediments and waste slurry are major sources of arsenic in food and 
water. To mitigate the toxic effect on human health, the maximum contaminant level 
(MCL) for arsenic in drinking water was lowered from the previous 50 ppb to 10 ppb, 
effective in January 2006.  
Arsenic is a redox active element, with As(V) or (III) being the two most common 
stable oxidation states in soils [77]. In general, inorganic arsenic is more toxic than 
organic arsenic [78], and arsenic in soils is less bioavailable and less bioaccessible than 
arsenic in water due to soil adsorption effect [79]. 
Arsenate can strongly interact with soils, especially, iron (hydr)oxides. Adsorption of 
arsenate by iron (hydr)oxides have been widely studied [80-88]. These studies have 
focused on the adsorption and surface complexation of arsenic on the amorphous and 
 
45 
crystalline iron oxide structures, such as ferrihydrite and goethite. The complexation 
between arsenate and iron (hydr)oxide surfaces has been known to be inner-sphere 
surface complexation as either mono-dentate core sharing, bi-dentate core sharing, or bi-
dentate edge sharing complexes [80, 89].  
Laboratory-scale and field-scale studies have been reported on in situ remediation of 
As-contaminated groundwater by zero-valent iron (ZVI) [90-92] and iron oxides [93]. 
They observed that ZVI can reduce the concentration of As in aqueous phase. Recently, 
nanoscale iron-based media (such as zero-valent iron) have been studied for potential 
uses in environmental remediation [94-96]. Because of the small particle size, large 
surface area, and high reactivity, these nanoscale materials have showed great potential 
for treatment of contaminated soil and groundwater [97-99]. Cumbal and Sengupta [100] 
studied arsenic removal from water by hydrated iron oxides nanoparticles loaded on 
polymer-matrix, and the immobilized nanoscale iron oxides displayed high sorption 
capacity for both arsenite and arsenate. For the arsenic removal in groundwater by iron-
based nanoparticles, surface adsorption appears to an important mechanism [101]. 
Compared to commercial iron powder or granular iron particles, ZVI nanoparticles offer 
much faster sorption kinetics and are more deliverable in the subsurface. Consequently, 
iron nanoparticles hold great potential to immobilize arsenic in situ in contaminated soil 
and groundwater [94].  
However, because of the high reactivity and inter-particle interactions, ZVI 
nanoparticles tend to agglomerate rapidly, resulting in the formation of much large 
(>microscale) particles and loss of reactivity and soil mobility. To prevent iron 
nanoparticle agglomeration, various particle stabilization strategies were reported [102-
 
46 
104]. He and Zhao [104] reported a new method for synthesizing stabilized iron 
nanoparticles by using some low-cost and environmentally benign starch and cellulose as 
a stabilizer. The stabilized nanoparticles displayed much improved physical stability, soil 
mobility, and reactivity compared to non-stabilized iron particles.  
To quantify relative As mobility and leachability in soil, two operationally defined 
measures, bioaccessibility and TCLP (toxicity characteristic leaching procedure) 
leachability, have been commonly used. Bioaccessibility is quantified by a 
physiologically based extraction test (PBET), which mimics the conditions in human 
stomach and essentially reflects in vivo accessibility of As [50]. TCLP is an EPA-defined 
standard method for measuring extractability of various chemicals from solid wastes. 
Earlier, a number of researchers [105-107] used TCLP tests to evaluate the leachability of 
As in contaminated soils. 
Akhter et al. [107] concluded that higher iron content in soil reduces the leachability of 
arsenic. Yang et al. [62] observed that high iron content reduced the bioaccessibility of 
arsenic in soil. 
The objective of this study was to test the effectiveness of stabilized nanoparticles for 
reducing the bioaccessibility and TCLP leachability of arsenic in soils. Three types (ZVI, 
FeS, and Fe
3
O
4
) of stabilized nanoparticles were prepared using a water-soluble starch as 
a stabilizer, and then used for treating two representative soils in batch experiments. 
Effects of the Fe-to-As molar ratio were examined on the treatment effectiveness. 
 
 
 
 
47 
3. 2 Materials and Method 
3. 2. 1 Materials  
All chemicals used in this study were of analytical grade or above, and were obtained 
from Fisher Scientific (Pittsburgh, PA). All solutions were prepared with deionized water 
(18 M??cm).  
An As-contaminated sandy soil (As concentration: 315 mg/kg and denoted as WAOS) 
was collected from Washington Orchard, an orchard contaminated from application of 
As-based pesticides. In addition, a relatively clean clay soil was collected near a small 
police firing range in east-central, Alabama, USA. Both soils were first fractionated using 
standard sieves, and soil fractions of <250 ?m were used in all experiments. 
The WAOS soil has an iron content of nearly 5.24% and a soil pH of 6.75. The range 
soil has a higher iron content soil (12.2%) and a soil pH of 4.83. For subsequent testing, 
the range soil was first spiked with arsenic following the procedures by Yang et al. [62], 
resulting in an arsenic concentration of 89 mg/Kg.   
 
3. 2. 2 Preparation of Nanoparticles  
The method developed by [104] He and Zhao was adopted for preparing ZVI 
Nanoparticles. In brief, a water-soluble starch (Alfa Aesar, Wall Hill, MA) was used as a 
stabilizer in the preparation. The preparation was carried out in a 250 mL flask. Before 
use, deionized (DI) water and starch solution were purged with N
2
 for 2 h to remove 
dissolved oxygen (DO). FeCl
3
 stock solution was added to a starch solution (2.4%) 
through a buret, to give a final Fe concentration of 2.35 g/L and a starch concentration of 
1.2%. The final pH was 8.1. Then, Fe
3+
 was reduced to Fe
0
 using stoichiometric amounts 
 
48 
of sodium borohydride (equation 3.1). To ensure efficient use of the reducing agent BH
4
-
, 
the reactor system was operated in the absence of DO. The flask was shaken via hands 
during the reaction.  
 
Fe(H
6
O
3
)
3+
 + 3 BH
4
-
 +3 H
2
O ?  Fe
0
 +10 (0.5)H
2
 +3 B(OH)
3
 (3.1) 
A method used by Si et al. [108] was modified for synthesizing magnetite (Fe
3
O
4
) 
nanoparticles. First, 50 mL of an aqueous solution of FeCl
2
?4H
2
O (5.0 g/L as Fe) was 
added dropwise to a 50 mL aqueous solution of 2.55% (w/v) starch solution under 
continuous shaking. The mixture was shaken for 30 minutes to allow for formation Fe
2+
-
starch complex. Then, the pH of the solution was then increased slowly to 11 by adding 
0.5 M NaOH solution. The reaction mixture was subsequently aged for 1 h with constant 
shaking, give a final Fe concentration of 2.35 g/L and a starch concentration of 1.2%. 
(equation 3.2).  
6 Fe
2+ 
+ O
2
 +12 OH
- 
?  2 Fe
3
O
4 
+ 6 H
2
O    (3. 2) 
 The method by Xu and Akins [109]  for preparing CdS nanoparticles was modified for 
preparing our starch-stabilized FeS nanoparticles. First, deionized water and starch 
solution (3.6%) were purged with pure nitrogen to remove the dissolved oxygen. The, 
FeCl
2
 solution was prepared and added into starch solution to form Fe-starch complex 
(Fe: 3.525g/L, Starch: 1.8% (w/v)). Na
2
S solution (4.03g/L as S) was added dropwisely 
to Fe-starch solution to form the FeS nano-particles, give a final Fe concentration 2.35g/L 
and 1.2% starch concentration. Final pH was 6.8. Equation 3.3 gives is the reaction 
stoichiometry. 
 
 
49 
FeCl
2 
+ Na
2
S ?  FeS +2 NaCl     (3.3) 
 
3. 2. 3 Treatments for As-Contaminated Soils  
A series of soil treatment tests were performed in 15 mL centrifuge tubes (Fisher, 
polypropylene tube), where 2 g of two kinds of an As-laden soils sample was mixed with 
a nanoscale particles (Fe
3
O
4
, FeS, or NVI) suspension. To test the effect of iron dosage 
on arsenic immobilization effectiveness, the range of Fe:As molar ratio (5:1, 10:1, 25:1, 
50:1, 75:1, 100:1) was tested. In all cases, the soil-to-solution ration was either 1g: 2mL 
solid-aqueous ratio (for Fe:As molar ratio of 5:1, 10:1, or 25:1) or and 1g: 10mL solid-
water ratio (for Fe:As molar ratio of 50:1, 75:1, 100:1). In addition, control tests were 
carried out in parallel with 2 g of a soil with 4 mL and 10 mL, respectively, of 1.2% 
starch solution. After the mixtures were shaken thoroughly for 5 minutes, the tubes were 
placed on a rotator for 3 days or 7 days. After the treatments, all samples were 
centrifuged with 6000 g force (Fisher, Accuspri 400 centrifuger). Arsenic and iron 
concentrations in the supernatants were monitored after centrifuging. Upon removal of 
the supernatant, each soil sample was oven-dried at 70?C for one day. 0.1 g of treated 
soils was sampled and used for PBET, 0.5 g for TCLP tests, and 1.0 g for soil pH 
measurements. To ensure data quality, all tests were performed in duplicates. 
 
3. 2. 4 TCLP and PBET Extraction Test  
The bioaccessibility of lead was monitored by a modified PBET method [61, 62]. 
PBET extraction solution was made using a 0.4 M glycine solution adjusted to a pH of 
1.5 using HCl solution. In each PBET test, 0.1 gram of a soil sample is mixed with 10 mL 
 
50 
of the extraction solution, i.e. a, 1:100 solid-to-solution ratio of 1:100. During the 1-h 
extraction, water temperature was maintained at body temperature (37?2 ?C) with a water 
bath. After the extraction, the samples were centrifuged at 1000 g force. The supernatant 
was then filtered using 0.45?m filter (Fisher, DISPNR 25mm 0.45?m filter), and then 
analyzed for arsenic extracted. To ensure QA/QC, NIST soil samples were also subjected 
to the same procedure. 
TCLP tests were performed to evaluate the leaching potential of arsenic in the 
untreated and treated As-contaminated soils following the US. EPA protocol (Method 
SW-846). In brief, 0.5 g of an air-dried soil sample was mixed with the TCLP extraction 
solution at a solid-to-liquid ratio of 1:20. The mixtures were placed on a rotating shaker 
operated at 30 rpm. After 18 hours of extraction, the samples were centrifuged at 1000 g 
force, and the supernatants were separated by 0.45 ?m filter. The soluble arsenic 
concentration in the filtrate was analyzed with AAS. 
 
3. 2. 5 Analytical Method  
Aqueous samples were diluted as necessary and analyzed for aqueous As and Fe 
concentrations. A graphite-furnace atomic absorption spectrometer (GFAA) was used to 
analyze As concentration. Aqueous Fe concentrations in samples were analyzed using a 
flame atomic adsorption spectrometer (FLAA). Solution pH was measured with a pH 
meter (Thermo Orion, pH meter 410).  
 
 
 
 
51 
3. 3 Results and Discussion 
3. 3. 1 Treatment Effects on Soil Chemistry  
As mentioned earlier, treated soil samples were first centrifuged. To ensure mass 
balances for both the nanoparticles and arsenic in the systems, total Fe and As in the 
supernatants were also analyzed. For the range soil samples treated with the three 
nanoparticles, iron concentration was less than 1% of total Fe amount added, and arsenic 
concentration was less than 0.5% total arsenic initially in the soils. These observations 
indicated that upon the high speed centrifuging, virtually all of the nanoparticles were 
removed from the aqueous phase. For the WAOS soil samples, 0.1-1.1% of total iron and 
1.1-2.4% of arsenic stayed in the supernatant when treated with Fe
3
O
4
 nanoparticles, 
whereas 1.5-2.6% of total iron and 0.9-2.7% of arsenic remained in the supernatants with 
nanoscale NVI particles; and 1.3-3.9% of total iron and 0.3-2.7% arsenic were in the 
supernatants with FeS nanoparticles. These results again indicated that most of iron 
treatments were removed from the solution.  
The soil pH was also measured after the treatments. All results were showed in Table 
3.1. For two type?s soils, pHs of control samples were similar with the initial soil pH, and 
after NVI and FeS Nanoparticles treatments, the soil pHs didn?t change. However, 
because Fe
3
O
4
 nanoparticles solution has a high pH (~11), the pH of soil samples 
increase a little after nanoscale Fe
3
O
4
 particles treatment. 
 
52 
 
 
Table 3.1 pH Change for Different Molar ratio Treatments 
Fe/As 
molar 
ratio 
Initial 
pH 
Control 
samples* 
(0) 
5 10 25 50 75 100 
WAOS- 
NVI* 
6.74 6.71 6.73 6.82 6.79 6.89 
WAOS- 
Fe
3
O
4 
7.03 7.15 7.17 7.19 7.26 7.37 
WAOS- 
FeS 
6.75 6.73-6.89
6.64 6.71 6.67 6.78 6.86 6.95 
As-
spiked- 
NVI 
4.78 4.86 4.82 4.91 4.89 5.02 
As-
spiked- 
Fe
3
O
4
 
4.96 5.03 5.04 5.12 5.17 5.35 
As-
spiked- 
FeS 
4.83 4.83-4.93
4.81 4.91 4.89 4.87 4.94 5.03 
*Control samples: for different soil/solution ratio and starch solution. 
*WAOS-NVI: It is WAOS soil and NVI treatment. 
 
 
 
53 
 
0
10
20
30
40
50
60
70
80
0 5 10 25 50 75 100
Fe:As ratio
B
i
o
accessi
b
i
l
i
t
y
 o
f
 A
s
 W
A
O
S
 so
i
l
 (
%
) 
 .
magnetite nanoparticles
zero valent iron nanoparticles
ferrous sulfide nanoparticles
 
Figure 3. 1 Comparison of As bioaccessibility (PBET) of WAOS soil sample by different 
Fe/As ratio iron based nanoparticles treatments. 
 
 
 
54 
 
0
5
10
15
20
25
30
0 5 10 25 50 75 100
Fe/As molar ratio
B
i
o
accessi
b
i
l
i
t
y
 o
f
 A
s
 o
f
 A
s
-
s
p
i
k
e
d
 so
i
l
 (
%
)
  
.
magnetite
nanoparticles
zero valent iron
naoparticles
ferrous sulfide
nanoparticles
 
Figure 3. 2 Comparison of As bioaccessibility (PBET) of As-spiked soil sample by 
different Fe/As ratio iron based nanoparticles treatments. 
 
 
55 
0
0.5
1
1.5
2
2.5
3
3.5
0 5 10 25 50 75 100
Fe:As ratio
TC
L
P
 l
e
a
c
ha
bi
l
i
t
y
 of
 A
s
 f
o
r
 W
A
O
S
 s
o
i
l
(
%
)
  
.
magnetite nanoparticles
zero valent iron
nanoparticles
ferrous sulfide
nanoparticles
 
Figure 3.3 Comparison of As TCLP leachability of WAOS soil samples by different 
Fe/As ratio iron based nanoparticles treatments. 
 
56 
 
0
0.1
0.2
0.3
0.4
0.5
0.6
0 5 10 25 50 75 100
Fe/As molar ratio
T
C
L
P
 L
each
ab
i
l
i
t
y
 o
f
 A
s
 f
o
r
 A
s
-
s
p
i
k
e
d
 so
i
l
 (
%
)
 
.
magnetite
nanoparticles
zero valent iron
nanoparticles
ferrous sulfide
nanoparticles
 
Figure 3. 4 Comparison of As TCLP leachability of As-spiked soil samples by 
different Fe/As ratio iron based nanoparticles treatments. 
 
 
57 
 
3. 3. 2 Reduction of PBET Bioaccessibility 
The PBET-based bioaccessibility of arsenic for WAOS soil samples were measured 
after the treatments, and is given in Figure 3.1. Figure 3.1 shows that with the increasing 
Fe/As molar ratio, the bioaccessibility decreased progressively. After three days of the 
treatments, the bioaccessibility of As decreased from an initial 71.3?3.1% to 29.8?3.1%, 
30.9?3.2%, 37.6?1.2% for Fe/As ratio 100:1 Fe
3
O
4
, NVI, FeS nanoparticles, 
respectively. Earlier, Subacz [110] studied the reduction of bioaccessibility of the same 
soil with Fe/As ratio 100:1 FeCl
3
 amendment, and the bioaccessibility decreased to 33% 
after treatment. Compare to Subacz?s result, magnetite nanoparticles treatment appears a 
little better (29.8% vs. 33%) than normal iron amendments. 
For the As-spiked soil, similar results were observed (Figure 3.2). After three days of 
the treatments, the bioaccessibility of As decreased from an initial 23.2?2.8% to 
10.4?1.5%, 11.7?1.4%, 16.1?0.8% for Fe/As ratio 100:1 Fe
3
O
4
, NVI, FeS nanoparticles, 
respectively. For the two soils, Fe
3
O
4 
nanoparticles appear to be most effective for As 
immobilization. The better performance of Fe
3
O
4 
nanoparticles treatment is at least 
partially due to the elevated soil pH upon the treatment. Earlier, Yang et al. [62] reported 
that bioaccessibility decreases with increasing soil pH. The treatment was more effective 
for the range soil, which has a >2 times greater iron content. Yang et al. [62] concluded 
that bioaccessibility of soil arsenic decrease with increasing iron content. Akhter et al. 
[107] also studied relationship between TCLP leachability of As-contaminated soils and 
iron content, and he correlated the leachability of As in soils and with iron content. 
 
 
58 
3.3.3 Reduction of  TCLP Leachability  
The initial TCLP leachability for both of the soils was quite low. The results were 
showed by Figure 3.3 and 3. 4. TCLP leachability is the As percent in the leachate vs. the 
total As content. The initial TCLP leachability for the untreated range soil was 0.51% and 
3.28% for the untreated WAOS soil. This result was in accord with those reported by 
Akhter et al. [107], they studied As-contaminated soil from industrial sites. No TCLP 
leachates showed arsenic concentrations as high as 5 mg/L, which is the EPA benchmark 
value for a hazardous waste. When treated at the 100 Fe/As ratio, the leachability of 
arsenic in the WAOS soils decreased from the initial 3.28?0.78% to 1.63?0.16%, 
1.83?0.06%, 1.73?0.17% by Fe
3
O
4
, NVI, FeS nanoparticles, respectively; whereas the 
leachability for the range soil decreased from the initial 0.51?0.11% to 0.17?0.04%, 
0.24?0.03%, 0.27?0.04% by Fe
3
O
4
, NVI, FeS nanoparticles, respectively. For both soils, 
the As leachability was the lowest when Fe
3
O
4
 nanoparticles were applied. Miller et al. 
also observed that As in TCLP liquid was reduced for an As-contaminated soil from 1.42 
to 0.26 mg/L by adding iron treatments (added FeSO
4
) [105]. 
 
 
 
59 
 
0
10
20
30
40
50
60
Magnetite
nanoparticles
Zero valent iron
nanoparticles
Ferrous sulfide
nanoparticles
Different nanoparticles
B
i
o
accessi
b
i
l
i
t
y
 o
f
 A
s
 W
A
O
S
 so
i
l
(
%
)
  
.
3 days treatment
7 days treatment
 
Figure 3. 5 Comparison of As bioaccessibility of WAOS soil samples for 3 and 7days 
treatment by different iron based nanoparticles 
 
60 
 
0
0.5
1
1.5
2
2.5
3
3.5
4
4.5
5
Magnetite
nanoparticles
Zero valent iron
nanoparticles
Ferrous sulfide
nanoparticles
Different nanoparticles treatment
T
C
L
P
 le
a
c
h
a
b
ility
 o
f
 A
s
 fo
r W
A
O
S
 s
o
il (%
)  
 .
3 days treatment
7 days treatment
 
Figure 3. 6 Comparison of As leachability of WAOS soil samples for 3 and 7days 
treatment by different iron based nanoparticles 
 
61 
 
0
5
10
15
20
25
Magnetite
nanoparticles
Zero valent iron
nanoparticles
Ferrous sulfide
nanoparticles
Different nanoparticles treatment
B
i
oa
c
c
e
s
s
i
bi
l
i
t
y
 of
 A
s
 f
o
r
 A
s
-
s
pi
ke
d s
o
i
l
 (
%
)
  
.
3 days treatment
7 dayse treatment
 
Figure 3. 7 Comparison of As bioaccessibility of As-spiked soil samples for 3 and 7days 
treatment by different iron based nanoparticles 
 
62 
 
0
0.05
0.1
0.15
0.2
0.25
0.3
0.35
0.4
0.45
0.5
Magnetite
nanoparticles
Zero valent iron
nanoparticles
Ferrous sulfide
nanoparticles
Different nanoparticles treatment
T
C
L
P
 le
a
c
h
a
b
i
lity
 o
f
 A
s
 fo
r
 A
s
-s
p
i
k
e
d
 s
o
il (%
)   
. 
3 days treatment
7 days treatment
 
Figure 3. 8 Comparison of As leachability of As-spiked soil samples for 3 and 7days 
treatment by different iron based nanoparticles 
 
63 
3. 3. 4 Effect of Treatment Time 
Figure 3.5 -3.8 compares the TCLP and PBET results when the soils were treated for 3 
days and 7 days. From these figures, the bioaccessibility and leachability of As for 100:1 
Fe/As molar ratio by 3 or 7 days treatment are comparable. A student t-test revealed no 
significant difference between the results from the two treatment times. Earlier, Subacz 
(2004) observed that the bioaccessibility of contaminated soils when amended with FeCl
3
 
for 3 days and 7days differed significantly. These results suggest that the stabilized 
nanoparticles offer rather fast mass transfer and reaction kinetics. Most of adsorption of 
arsenate occurred in one hour for nano zero valent iron groundwater treatments (Kanel et 
al. 2006). And the As adsorption reaction on nanoscale NVI were more quick than on 
micron ZVI. 
 
3. 3. 5 Mechanisms of Arsenic Immobilization by Iron-based Nanoparticles 
The mechanisms for arsenic sorption by iron oxides have been studied extensively [85, 
86, 88, 93]. Waychunas et al. [111] described the adsorption mechanism of nanoscale 
iron oxides in soils and sediments, they concluded that nanoscale iron oxides have bigger 
surface area than microscale iron oxides to occur surface complexation reaction. Gao and 
Mucci stated that surface complexation is the main reaction mechanism for arsenic 
uptake by iron oxides (equations 3.4-3.8), [112].  
 
 
 
 
 
64 
>FeOH + H
+
 ?  >FeOH
2
+ 
    (3.4) 
>FeOH ? >FeO
-
 +H
+    
 (3.5) 
>FeOH + H
2
AsO
4
- 
?
 
>FeAsO
4
2-
 +H
+
 +H
2
O     (3.6) 
>FeOH + H
2
AsO
4
- 
?
 
>FeAsO
4
H
-
 +H
2
O        (3.7) 
>FeOH + H
2
AsO
4
- 
+H
+
?
 
>FeAsO
4
2-
H
2
 +H
2
O (3.8) 
Previous work reported that zero-valent iron nanoparticles react with both 
contaminants and dissolved oxygen as well as water [103]. After the oxidation of 
zerovalent iron, an iron oxides layer forms at the surface of zero-valent iron particles, 
whereas some Fe
2+
 or Fe
3+
 ions can release into the aqueous phase. Cornell and 
Schwertmann [113] proposed that iron oxides form a passivation of the surface if the 
surface sites become saturated with iron oxides (equation 3.9-3.12). Arsenic, both 
arsenite and arsenate can be adsorbed on the surface of iron oxide [90-92]. Jegadeesan et 
al. [94] also proposed adsorption of arsenic on the corroded iron surface is the main 
mechanism of arsenite removal by nanoscale NVI from groundwater. Nanoscale zero-
valent iron has a structure which 19% were in zero valent state with a coat of 81% iron 
oxides [101]. Kanel et al. also confirmed that nanoscale zero-valent iron and arsenate 
forms an inner-sphere surface complexation, 99% arsenate was adsorbed by nanoscale 
zero-valent iron in one hour. Bang et al. [114] claimed that arsenic can also be removed 
by Fe
0
 through reducing arsenite and arsenate to zerovalent arsenic, which is insoluble in 
water.   
Fe
0 
+2 H
2
O?  Fe
2+
 +H
2
 (g)
 
+ 2 OH
-
    (3.9) 
4 Fe
2+
 +4H
2
O +O
2
 ? 4 Fe
3+ 
+ 8 OH
-
 (3.10) 
 
65 
Fe
2+
 ? e ? Fe
3+
        (3.11) 
Fe
3+
 +2H
2
O?  FeOOH +3H
+
       (3.12) 
Arsenic concentrations typically decrease under anoxic conditions by sulfide minerals 
[115]. Arsenic sorption on FeS was studied by [116] with X-ray absorption spectroscopy. 
They proposed eqn (3.13) as the main reaction for arsenite removal by FeS, which was 
supported by their XRD results. Nanoscale FeS particles have greater surface area, the 
sorption of arsenic on FeS nanoparticles can explain the immobilization of arsenic. 
 
3FeS + H
3
AsO
3
 ?  FeS
2
 +FeAsS +Fe(OH)
3
 (3.13) 
             
3. 4 Conclusion 
The results of this investigation suggest that iron-based nanoparticles can be added to 
soils to decrease As bioaccessibility, leachability and the potential of bioavailability. The 
starch-stabilized nanoparticles were found effective to reduce both TCLP-leachability 
and PBET-bioaccessibility of As in As-contaminated soils. The bioaccessibility and 
leachability decrease with increasing Fe/As molar ratio. After three days treatments, the 
bioaccessibility of As decreased from an initial 71.3?3.1% to 29.8?3.1%, 30.9?3.2%, 
37.6?1.2% for Fe/As ratio 100:1 Fe
3
O
4
, NVI, FeS nanoparticles, respectively, and for the 
100 Fe/As ratio, the leachability of arsenic of in a range soil decreased from an initial 
0.51?0.11% to 0.17?0.04%, 0.24?0.03%, 0.27?0.04% by Fe
3
O
4
, NVI, FeS 
Nanoparticles, respectively. Fe
3
O
4
 nanoparticles worked better than the other two 
nanoparticles in reducing the bioaccessibility and leachability. No significant difference 
in the effectiveness was evident between 3 days and 7 days treatments. Compare to two 
 
66 
soils, the treatment was more effective for the range soil which has much lower iron 
content. These results suggest that stabilized nanoparticles may serve as alternative media 
for in situ immobilization of arsenic in soils, especially soils with high As concentration 
and low Fe content. 
 
 
67 
 
 
 
 
CHAPTER      IV 
CONCLUSIONS 
 
Conclusion 
In Chapter 2 of this study, the Pb bioaccessibility of three types of Pb-contaminated 
solids was reduced by the use of phosphate amendment KH
2
PO
4
 (potassium dihydrogen 
phosphate) regardless of the initial moisture contents. Rapid potassium release into 
supernatants indicated that the amendment KH
2
PO
4
 dissolved and released potassium as 
well as phosphate ions quickly in free water (moisture) during soil remediation process. It  
also showed that there was enough phosphate in the in situ process to form lead-
phosphate salt. 
In the PBET process for the three amended solids, the concentration of K
+
, Pb
2+
 and 
H
2
PO
4
-
 increased in the one hour PBET time. For both Pb-spiked soil and Pb-spiked 
sand, the saturation indices of lead-phosphate minerals were below 0 (saturation indices 
of chloropyromorphite ranged from -2.99 to -1.45; saturation indices of 
hydroxypyromorphite ranged from -21.58 to -20.05; saturation indices of PbHPO
4
 ranged 
from -2.91 to -2.49). This result indicated no potential lead mineral precipitation in the 
PBET solution for these two solids. However, for the Pb-contaminated soil, due to the 
high initial soil Pb concentration, after 10 minutes in the PBET, the SI for all points were 
 
68 
above 0 (0.56-1.12).  These numbers indicated that the PBET solutions were 
oversaturated.  
In the homogeneous chloropyromorphite formation test in PBET solution, under the 
normal PBET condition (HCl, 1 hour extraction, 37 ?C) less than 5% of the Pb was 
removed from the aqueous phase, indicating that the homogeneous precipitation of 
chloropyromorphite from the PBET was not occurring to any significant extent.  
Similarly, even after seven days, very little Pb (10%) was removed from the normal 
PBET solution at 25 ?C, and the SI for chloropyromorphite was 6.49. In the HNO
3
- based 
PBET solution, Pb removal was also less than 10% after seven days. For homogeneous 
chloropyromorphite formation experiments under different conditions, lead solubility was 
also greater than predicted for chloropyromorphite in most solutions. Under 
heterogeneous conditions, after one hour aging at 37 ?C, 89.1% Pb was still in aqueous 
phase. It showed that the transformation reaction of lead to solid phase was very slow in 
the present of phosphate in PBET solution.  
These results illuminate that the lead removal reaction (chloropyromorphite formation) 
is very slow in PBET solution under both homogenous and heterogeneous conditions. 
Based on this conclusion, we speculate that high concentrations of glycine may block the 
reaction among Pb
2+
, Cl
- 
and PO
4
3- 
by affecting the collision of ions since 
chloropyromorphite formation involves multiple ions. This result also indicates that the 
immobilization of lead occurs more under in situ conditions than in the PBET process. 
Further research work is needed to monitor chloropyromorphite formation in amended 
soils by methods to definitively prove metal speciation (e. g. XAS, XRD). We also found 
lead solubility was greater than predicted for chloropyromorphite in the PBET condition, 
 
69 
suggesting further studies should be conducted at different pH values (e. g. 4.0 and 7.0) 
to indentify the cause for the higher than predicted solubility of lead in PBET solution.  
The ion activity products of chloropyromorphite should be calculated with equilibrium 
Pb
2+
, Cl
-
 and H
2
PO
4
-
 concentrations at different pH conditions in PBET solutions, to 
compare the data to the reported K
SO
 of chloropyromorphite.  
In the chapter 3 of this study, iron based nanoparticles were added to soils in order to 
reduce arsenic bioaccessibility and leachability. Using a soluble starch as stabilizer, most 
iron based nanoparticles attached to the solid phase after three days of treatment. 
Nanoparticles treatments were effective in reducing the bioaccessibility and leachability 
of arsenic from contaminated soils. As bioaccessibility and leachability decreased with 
increasing iron content. Among the three types of nanoparticles (100:1 Fe/As molar ratio) 
tested, the bioaccessibility and leachability of magnetite (Fe
3
O
4
) nanoparticle-treated 
soils were lower than the other two nanoparticle treatments, ferrous sulfide (FeS) and 
zero-valent iron (ZVI). No significant difference between three days and seven days of 
treatment was observed, indicating that most immobilization occur during first three 
days.  
With regard to comparison between two soil types, treatment was more effective on 
the lower iron content, a high arsenic concentration soil. These results suggest that such 
laboratory-scale testing may be worthwhile, especially at sites that have high arsenic 
concentrations and low Fe contents. Further investigation is needed to understand the 
mechanism of arsenic immobilization by FeS and ZVI. The potential immobilization of 
arsenite by nanoscale iron based particles should also be investigated. 
 
70 
 
 
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