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PHOSPHATE-INDUCED LEAD IMMOBILIZATION IN CONTAMINATED SOIL
A THESIS PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
MASTER OF SCIENCE
UNIVERSITY OF FLORIDA
This thesis is dedicated to my grand parents, Jinhyung Lee and Youngsook Kim.
I wish to extend my thanks to Dr. Lena Q. Ma for her assistance and guidance of
my research. I would also like to thank members of my committee, Drs. Willie G. Harris
and Jean Claude Bonzongo, for their advice, support and critical review of this thesis. I
gratefully acknowledge Dr. Xinde Cao who worked with me and helped me a lot. I wish
to thank Thomas Luongo for laboratory analysis and being a good instructor. Special
thanks go to Drs. Mrittunj ai Srivastava and Gregory E. MacDonald for their assistance
with plant identification procedure.
I would like to thank my parents Byungkyu Yoon and Sookj ae Lee for their love,
encouragement and support throughout my studies.
TABLE OF CONTENTS
ACKNOWLEDGMENT S ................. ................. iv.............
LIST OF TABLES ................ ..............vii .......... ....
LI ST OF FIGURE S ................. ................. viii............
AB STRAC T ................ .............. ix
1 INTRODUCTION ................. ...............1.......... ......
Lead in the Environment .............. ...............1.....
Health Effects of Lead ................. ...............3.......... .....
Lead Remediation Techniques .............. ... ...............6...
Phosphate Induced Lead Immobilization ....._____ ..... ...___ .. ........_......6
Phosphoric acid .............. ...............8.....
Synthetic apatite .............. ...............8.....
Bone meal apatite ................. ...............9...._._._ .....
Phosphate rock ........._.._.. ...._... ...............10....
Rhizosphere ........._.. ..... ._ ._ ...............11....
Phytoextracti on ................. ................. 13..............
Phytostabilization ................. ...............13......._ ......
Sum m ary ................. ...............14......... ......
2 EFFECTS OF APPLICATION METHODS ON P-INDUCED Pb
IMMOBILIZATION IN A SOIL COLUMN ................. ..............................16
Introducti on ................. ...............16.................
Materials and Methods .............. ...............18....
Soil Characterization ................ ...............18.................
Soil Sample Preparation .............. ...............18....
Column Leaching .............. ...............19....
Analytical Procedure .............. ...............20....
Re sults ................ ............ ...............21.......
Soil Characterization ................. ..... ...............21
Effects of P Application on Soil pH ................. ...............22........... ..
T CLP Pb ................. ...............23...
Lead Bioavailability .............. ...............24....
Distribution of Pb in Soil Column............... ...............25.
Leachate Analysis............... ...............29
3 THE EFFECTS OF PLANTS ON PHOSPHATE-INDUCED Pb
IMMOBILIZATION IN THE RHIZOSPHERE SOIL .............. .....................3
Introducti on ............. ...... ._ ...............33...
Materials and Methods .............. ...............34....
R e sults.............. ......._ ...............37...
Soil Characterization .............. ...............37....
Bioavailable Pb.................... ...............38
Formation of Lead Phosphate ................. ...............39 ........._....
Conclusion ................. ...............43......... ......
4 SCREENING OF PLANTS FOR ACCUMULATION OF Pb, Cu, Zn FROM A
CONTAMINATED SITE ................. ...............45.................
Introducti on ................. ...............45.................
Materials and Methods .............. ...............47....
Site Characterization ............... .... .... ................4
Sample Preparation and Chemical Analysis............... ...............48
R e sults................ ...... .... .. .. ......... .............4
Metal Concentrations in Soils .............. ...............48....
Metal Concentrations in Plants................. ..... ...............5
Accumulation and Translocation of Metals in Plants ................. ................ ...52
Conclusion ................ ...............60.................
5 CONCLUSION ................. ...............62.................
LIST OF REFERENCES ................. ...............64................
BIOGRAPHICAL SKETCH .............. ...............72....
LIST OF TABLES
1-1 Theoretical solubility of Pb in different ................. ...............7............ ..
2-1 Soil characterization at the contaminated site ........._.._.. ...._... ......._.._......21
2-2 pH of soil after treatment. ........... ..... .._ ...............22..
2-3 Pb concentration in column leachate (Clg L1) ..........._.._ ....... .........._.......29
3-1 Selected characteristics of the lead-contaminated soil ................ ......................37
3-2 Water soluble phosphorous in plant rhizosphere soil (ppm) ................. ................42
3-3 pH of rhizosphere soil treated with phosphate rock after 4-weeks growth..........._...43
4-1 Selected properties of soil samples collected from the contaminated site at
Jacksonville, Florida. ................. ...............49......... .....
4-2 Lead concentrations in soil and plant samples (mg kg- ) collected from the
contaminated site at Jacksonville, Florida............... ...............52
4-3 Copper concentrations in soil and plant samples (mg kg- ) collected from the
contaminated site at Jacksonville, Florida............... ...............54
4-4 Zinc concentrations in soil and plant samples (mg kg- ) collected from the
contaminated site at Jacksonville, Florida............... ...............55
4-5 Accumulation and translocation of Pb, Cu and Zn in selected plants ................... ...56
4-6 Plants with high BCF and low TF for phytostabilization............... ...........5
LIST OF FIGURES
2-1 TCLP metal contents in sectioned soil columns after leaching .............. ................24
2-2 PBET extractable Pb contents in sectioned soil columns after leaching..................26
2-3 Total Pb concentration in sectioned soil columns (bottom section) ................... ......28
2-4 P concentration in column leachate (mg L 1) .............. ...............30....
3-1 Lolium rigidum after 4 weeks of growth ....._.._................ .............. ....3
3-2 Agrostis capillaries~~~1111~~~~111 after 4 weeks of growth ................. ............... .............36
3-3 Bra~ssica napus after 4 weeks of growth .............. ...............37....
3-4 Lead concentrations using the physiologically based extraction test in the
rhizosphere soil with phosphoric acid and with phosphate rock treatment. .............38
3-5 Scanning electron microscopy elemental dot map of Agrostis capillaries~~~1111~~~~111 ..............41
3-6 Scanning electron microscopy elemental dot map of Bra~ssica napus...................42
3-7 Scanning electron microscopy elemental dot map of Lolium rigidum .....................42
4-1 Map of Jacksonville site with lead concentration contour map and sampling
Abstract of Thesis Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Master of Science
PHOSPHATE-INDUCED LEAD IMMOBILIZATION INT CONTAMINATED SOIL
Chair: Lena Ma
Major Department: Soil and Water Science
Activities such as mining and manufacturing, and the use of synthetic products
result in heavy metal contamination of soil and water resources. The obj ectives of this
research were to 1) examine the effectiveness of Hyve P application methods on Pb
immobilization in soil using a column study; 2) explore the effects of plants on the
effectiveness of P-induced Pb immobilization in soil using a pot study; and 3) evaluate
metal accumulation in plants growing on a contaminated soil, based on Hield screening.
The soil column test was conducted to evaluate the effectiveness of different application
methods using phosphate rock (PR) and phosphoric acid (PA). Mixing both PA and PR
with the soil was the most effective method of Pb immobilization, reducing TCLP-Pb by
up to 95% (Toxicity Characteristic Leaching Procedure) and bioavailable Pb by 25 to 42%
in the soil. The application of PR as a layer in the soil column was effective in reducing
Pb migration in the soil column by 73 to 79%. In a greenhouse study, all three plants
(Agrostis capillariesll1~~~111~~~11 Lolium rigidum, and Bra~ssica napus) were effective in enhancing
Pb immobilization in the rhizosphere soil when P was applied as insoluble PR (13 to 20%
reduction in bioavailable Pb) but not when applied as a mixture of PR and PA.
A fieldstudy was conducted to screen 17 native plants for their potential for accumulation
of Pb, Cu and Zn from a heavy metal contaminated site. Elevated metal concentrations were
observed in the 17 plants collected from the experimental site, with total concentrations
ranging from 20 to 1,183 mg kg-l for Pb, from 5 to 460 mg kg-l for Cu, and froml7 to
598 mg kg-l for Zn. Though no metal hyperaccumulator was identified, G. pennelliana
was the most effective in taking up all three metals, while Cyperus was the most effective
in translocating all three metals from the roots to the shoots. Our study demonstrated that
the most effective P-application method coupled with plant growth can maximize the
effectiveness of phosphate-induced lead immobilization in soils.
Lead in the Environment
Lead (Pb), from the Latin plumbum, has an atomic number of 82 and an atomic
weight of 207. 19. While lead is the 36th most abundant element, it is the most abundant
heavy metal in the earth' s crust with an average concentration of 13 mg kgl (Brown and
Elliott, 1992; Nriagu, 1978). The relatively low Pb concentration in soil solutions
confirms reports oflead's low mobility among the heavy metals (Kabata-pendias and
Pendias, 1984). Natural mobilization of lead into the environment occurs principally from
the erosion of lead-containing rocks and through gaseous emissions during volcanic
activity (Waldron, 1980).
Important physical attributes of lead (i.e. high density, softness, flexibility, and
electric potential) ensure its widespread industrial application. An Anthropogenic source
of lead includes flanking paint, mining operations, smelter and industrial emissions, and
application of insecticides, which have contributed to elevated lead levels in
environment. Lead is also used in the production of ammunition, and has been used in
other metallic products (such as solder, some brass, bronze products, and piping), and
ceramic glazes, but its use in these capacities has been outlawed in the U.S.
(Scheuhammer and Norris, 1995). Chemicals such as tetraethyl lead and tetramethyl lead
were once used as gasoline additives to increase octane rating. However, their use was
phased out in the 1980s, and lead was banned from use in gasoline for transportation
beginning January 1, 1996 (USEPA, 2001a).
Lead occurs naturally in all soils in concentration ranging from 1 to 200 mg kg- ,
but rarely at levels that are considered to be toxic. However, increases of anthropogenic
sources of Pb in the environment have caused the Pb loading rate in soil to exceed its
natural removal rate by approximately 20-fold (Nriagu, 1978). There are several
thousands of Pb contaminated sites in United States. Currently there are 440 sites listed
on the EPA National Priority List (NPL), in which the main contamination of concern is
lead. Many studies show that there are high levels of Pb in various environments. A
national survey in the U.K indicates that total Pb concentrations in garden soils ranges
from 13 to 14,100 mg kg-l (Cotter-Howells, 1996). Soil samples obtained from a former
battery-cracking site in Florida had Pb concentrations up to 13,500 mg kg-l (Cao et al.,
2001). Pichtel et al. (2000) reported Pb concentrations of a Superfund site and a Pb
battery dump that ranged from 195-140,500 mg kg- A recent survey found Pb
concentration in soils adj acent to homes and near expressways in Tampa, Florida as high
as 2,000 mg kg-l (Cao et al., 2001). Also uptake of heavy metals by humans, largely
through the respiratory and digestive pathways, has contributed to elevated blood Pb
concentration, which can cause impaired mental development, nervous system disorders,
immune-system dysfunction and DNA damage. Presence of hazardous levels of Pb in the
environment and the toxicity of Pb requires remedial action to protect the health of the
public. Considerable attention is now being paid to develop cost effective and less
disruptive remediation technologies to reduce human exposure via direct ingestion and
Ionic lead (Pb2+) is the dominant form of lead in the soil and groundwater.
However, Pb4+ Occasionally can be found in some highly oxidized soils. With the
exception of a few sporadic measurements in urban air, marine fish, and in human brains
there is relatively little information available on organometallic lead compounds (Nriagu,
Because lead enters the soil in various complex compounds, its reactions may differ
widely between geographic areas. Lead is found in soils most commonly as galena (PbS)
and in smaller quantities as cerrusite (PbCO3), angleSite (PbSO4) and crocoite (PbCrO4).
Lead added to soil may react with available soil anions such as SO42-, PO43-, Or CO32- to
form sparingly soluble salts (Waldron, 1980). Compounds such as lead carbonate
[Pb3(OH)2(CO3)2] and chloropyromorphite [Pb5(PO4)3C1] prOVide the least soluble
inorganic salts at near neutral pH's. There are several other mechanisms by which lead
may be immobilized in soils, such as complexation by soil organic matter humicc and
fulvic acids), which are then adsorbed onto soil solids, or by ion exchange at sites on the
solid material in the soil. These mechanisms may facilitate attachment of substantial
amounts of lead, but cannot account entirely for the low mobility of lead in soils
(Harrison and Laxen, 1981).
Health Effects of Lead
Since lead is a natural constituent in the environment, man has been under
inescapable exposure to Pb (Waldron, 1980), which originates from an array of sources
that include Pb contaminated soil, atmospheric Pb particles, Pb paint, Pb water pipes, and
Pb solder used to seal food cans. With such an array of potential sources, it is no wonder
that lead poisoning is ranked as the number one environmental health threat, by both the
U.S Agency for Toxic Substance and Disease Registry (ATSDR) and the Center for
Disease Control (CDC). Dust and soil have emerged as important lead ingestion
pathways in recent years. Although their amounts are small, once Pb enters the blood
only approximately 10% is excreted and the remainder stays in the body, affecting the
nervous system, blood cells, and metabolic processes and eventually residing in bone
material for decades (Scheckel and Ryan, 2003).
The maj or health effects of lead are manifest in three organ systems; the
hematological system, the central nerve system and the renal system (Needleman, 1992).
The hematological effects of lead have been recognized for many years. Such effects are
related to the derangement of hemoglobin synthesis, the shortened life span of circulating
erythrocytes and the secondary stimulation of erythropoiesis, which, may result in
anemia. Higher lead concentration may affect the degree of maturation of red blood cells,
globin synthesis, and the morphology and stability of cells (Nriagu, 1978).
Acute effects of lead on nervous system are both structural and functional,
involving the cerebellum as well as the spinal cord and the motor and sensory nerves
leading to specific areas of the body (Nriagu, 1978). Effects may also lead to the general
deterioration of intellectual functions, sensory functions, neuromuscular functions and
psychological functions. If calcium is deficient, synaptic transmission is impaired. In
addition, lead has also been reported to interfere with the normal metabolism of some
central neurotransmitters, including acetylcholine and catecholamines. Other effects,
which have been noted, include inhibition of cytoplasmic NAD(P)H oxidation by
cerebral mitochondria and the inhibition of adenylcyclase. Lead also affects the intrinsic
muscles and oculomotor nerves of the visual system. Changes in the character of the
blood cells and supporting fluids may lead to changes in the intraocular tension. Such
changes may lead to mydriasis or visual paralysis. Neuritis may occur within the visual
system itself and a scotoma, which may be limited to certain colors, is usually present in
optic neuritis. The actual effects of lead on auditory dysfunctions are not clear (Nriagu,
Nriagu (1978) summarized the functional and morphological effects of lead on the
kidney as follows:
Stage 1: swollen proximal tubular lining cells; mitochondrial changes; intranuclear
inclusion bodies (mostly lead-protein complexes); proximal tubular dysfunction,
Stage 2: fewer inclusion bodies; intense interstitial fibrosis; tubular atrophy and
Stage 3: renal tumor (observed in rats only).
The first stage is usually seen in children with short-term heavy exposure to lead.
The functional effects include the fanconi syndrome manifested by an increase in urinary
excretion of glucose, amino acids and phosphate. The second stage nephropathy can be
produced in experimental animals fed a high dose of lead and in workmen with excessive
lead exposure for several years. The duration and degree of lead exposure necessary to
produce irreversible kidney damage is not known. Dingwall-Fordyce and Lane (1963)
followed 425 retirees and others with long exposure to lead in a battery factory. They
found no increase in cancer incidence but saw excess in cerebrovascular accidents. Lead
can effect chromosomes, reproduction, immune system and endocrine system, but usually
is not considered as a human carcinogen.
Toxicity of lead to plants has not been frequently observed. Plants do not
translocate absorbed lead into the shoots while they are growing; instead, they
accumulate it in their roots. The health concern surrounding plants is that lead containing
soil particles can contaminate plant surfaces, which might be ingested by animals
Lead Remediation Techniques
In contrast to many organic contaminants, the immobile and non-degradable nature
of lead necessitates active remediation because natural attenuation will do little to reduce
its concentration (Nedwed and Clifford, 1997). There are several technologies available
for the remediation of land contaminated by heavy metals (USDA, 2000; Mulligan et al.,
2001). The selection of the most appropriate soil remediation method depends on the site
characteristics, concentration and types of pollutants of concern, and end use of the
contaminated medium (Mulligan et al., 2001). However, many of these technologies are
costly (e.g. excavation of contaminated material and chemical/physical treatment) or do
not achieve a permanent or aesthetic solution (e.g. encapsulation and vitrification).
Formation of insoluble heavy metal compounds immobilizes the metal and reduces
their bioavailability (Cotter-Howells and Caporn, 1996). Phosphate amendment has been
suggested as a cost-effective remediation option for Pb contaminated soils. Along with
that, phytoremediation, in which plants are used to do the remediation work, can be
another cost-effective and less disruptive approach.
Phosphate Induced Lead Immobilization
When considering potential Pb-immobilizing anions(Table 1) phosphates are less
soluble under earth surface equilibrium conditions than oxides, hydroxides, carbonates,
and sulfates (Ruby et al., 1994).
In its reactions with lead, phosphorus forms sparingly soluble orthophosphates
(pyromorphites) (Nriagu, 1972). Experimental evidence supports that Pb phosphates can
form rapidly when adequate Pb and phosphate are present in aqueous system (Zhang and
Ryan, 1999a; Ma et al., 1994b).
Phosphate minerals have been used to immobilize Pb in situ from aqueous
solutions and Pb contaminated soils (Cao et al., 2001, 2002; Ma et al., 1993, 1995;
Mench et al., 1994). The primary mechanism of Pb immobilization appears to be through
phosphate-mineral dissolution and subsequent precipitation of pyromorphite-like
minerals, though mechanisms such as cation substitution, adsorption and precipitation as
other Pb minerals are also possible. Hence, solubility of phosphate minerals largely
determines the effectiveness of in situ Pb immobilization in soil (Ma, 1996).
Table 1-1. Theoretical solubility of Pb in different phases
Pb Phase Stoichiometry log Ksp
Litharge PbO 12.9
Anglesite PbSO4 -7.7
Cerrusite PbCO3 -12.8
Galena PbS -27.5
Chloropyromorphite Pb5(PO4)3Cl -84.4
Hydroxypyromorphite Pb5(PO4)30H -76.8
Fluoropyromorphite Pb5(PO4)3F -71.6
BromoPyromorphite Pbs(PO4)3Br -78.1
Corkite PbFe3(PO4)(SO4)(OH)6 -112.6
Drugmanite Pb2(Fe,Al)(PO4)2(OH)2 NA
Hindsdalite PbAl3(PO4)(SO4)(OH)6 -99.1
Plumbogummite PbAl3(PO4)2(OH)sH20 -99.3
Lead-metal sulfate (Fe,Zn,Pb)SO4 NA
Iron-nanganese-lead oxide (Fe,Mn,Pb)O NA
Lead-metal oxides (Fe,Al,Zn,Pb)O NA
The rate of chloropyromorphite, one of pyromorphite minerals, precipitation is
kinetically rapid and controlled by the availability of soluble Pb and P as shown in Eq (1)
5Pb2+ + 3H2PO4- +Cl- O Pbs (PO4) 3l(s) + 6H+ (1)
Conversion of soil Pb to pyromorphite could immobilize soil Pb and reduce its
bioavailability. If sufficient soluble P is provided, the dissolution of Pb solids would then
be the limiting factor for pyromorphite formation (Yang et al., 2001). Dissolution of soil
Pb increases with decreasing pH (Zhang and Ryan, 1999b). Neutralization of protons
generated during chloropyromorphite precipitation shown in Eq (1) could favor the
reaction toward chloropyromorphite formation. Thus, initially acidifying the soil
followed by a gradual increase of soil pH should enhance the transformation of soil Pb to
pyromorphite. It is believed that adding H3PO4 to calcareous soil would lower soil pH to
facilitate dissolution of soil Pb and increase the activity of soluble P, thereby enhancing
pyromorphite formation (Yang et al., 2001).
Hydroxyapatite [HA, Calo(PO4)60H2], a member of the apatite family, is the
prototype of the inorganic constituent of bones and teeth, thus its crystal and solution
chemistry have been researched extensively. Most research has centered on synthetic HA,
because relatively pure natural HA is uncommon. Hydroxyapatite effectively and rapidly
attenuates Pb from aqueous solutions, exchange sites, and Pb contaminated soils (Ma et
Ma et al. (1993) have examined the utility of natural and synthetic apatite for in
situ stabilization of Pb in contaminated soils. The immobilization of Pb by HA was
attributed to HA dissolution followed by precipitation of hydroxypyromorphite [HP,
Pblo(PO4)6(OH)2]: aS shown in Eqs. (2) and (3):
Calo(PO4)6(OH)2(S) + 14 H (aq) O 10Ca2+(aq) + 6H2PO4-(aq) + 2H20 (2)
10Pb2+(aq) + 6H2PO4-(aq) + 2H20 OPblo(PO4)6(OH)2(S) + 14 H (aq) (3)
Laperche et al. (1996) investigated the use of HA as a soil additive with the goal
of converting soil Pb to HP. Hydroxyapatite was mixed with contaminated soil to test its
feasibility in reducing aqueous Pb and in converting the solid phase forms of Pb to HP.
The concentration of dissolved Pb in the suspension was reduced from 0.82 to 0.71 mg L~
Sat solution pH of 7.7 and to 0.22 mg L^1 at pH 5. Dissolved Pb concentrations in the
control samples increased with reaction time and decreasing pH.
Also, the addition of HA to a Pb polluted soil led to a substantial decrease of Pb
concentrations in the shoots of Gradd sudax (Boisson et al., 1999).
The mineral portions of animal bones and teeth are composed of HA. Bonemeal is
a less soluble source of phosphorous than synthetic apatite, but it is much more
economically feasible for field scale use. The ability ofbonemeal (finely ground, poor-
crystalline apatite) to immobilize pollutant metals in soils and reduce metal
bioavailability through the formation of metal phosphates has been evaluated by Hodson
et al. (2000). The researchers used a 1:50 bonemeal:soil mix. The mixture was packed
into columns and leached with synthetic rainwater. The pH increase appears to be due to
the dissolution of the bonemeal. Monitoring of leachates over a three-month period
indicated that bonemeal additions resulted in the immobilization of metals and an
increase in the pH of the column leachate, the soil pore water and the soils themselves.
Analytical scanning electron microscopy (SEM) and X-ray diffraction (XRD)
identified Pb and Ca-Zn phosphates as the reaction products. Assessment of metal
bioavailability by chemical extraction indicates that bonemeal additions to acidic soils
reduced metal bioavailability in soils via the formation of highly insoluble Pb and Zn
Ma et al. (1995) have shown that phosphate rock [PR, primarily fluorapatite (Calo
(PO4) 6F2)] effectively immobilized Pb from aqueous solutions, with Pb immobilization
ranging from 39 to 100%. The primary mechanism of Pb immobilization is via
dissolution of PR and subsequent precipitation of a fluorpyromorphite-like mineral [Pb l0
(PO4)6F2], although precipitation of Pb as hydrocerussite also occurred in some instances.
Phosphate rock immobilizes lead according to the following simplified equations (Ma
and Rao, 1999).
Ca9.5(PO4)5CO3FOH +10H+ 4 9.5Ca2+ + 5H2PO4-+ CO32- + F- + OH- (4)
9.5Pb2+ + 5H2PO4- + CO32- + F- + OH- 4 Pb9.5 (PO4)5CO3FOH + 10H+ (5)
Ma et al. (1995) demonstrated the potential of Florida PR to immobilize aqueous
Pb from Pb contaminated soils. The two most effective Florida PRtypes in immobilizing
Pb [Chemical (CF) and Occidental Chemical (OC)] were chosen for the column
experiments to evaluate the feasibility of using PR to immobilize Pb from a contaminated
soil and to evaluate the effects of different methods of mixing PR and soil and incubation
time on Pb immobilization efficiency. A lead contaminated soil was mixed with PR at a
ratio of 0: 10, 1:10, 2: 10 and 4: 10. After uniform mixing, the soil-PR mixture was placed
in columns. All PR types were effective in reducing Pb concentration from an initial
concentration of 4.82 Clmol L1 to below 72.4 nmol-l (the current EPA action level for Pb)
after 16 days of reaction and at a 4: 10 PR: soil ratio. Incubation time and mixing
methodology had little effect on the efficiency of Pb immobilization by PR. Hettiarachchi
et al. (2000) reported that PR was equally or more effective than triple super phosphate
Phosphorous is an essential plant nutrient and is of the most limiting factors in
plant growth.. Because roots can only absorb free phosphate, several mechanisms exist to
increase the soil P available to them (Cotter-Howells, 1996). Root exudates contain
phosphatase enzymes that can convert organic P to phosphate in the rhizosphere
(Haeussling and Marschner, 1989), as do rhizosphere microorganisms (bacteria and
fungi), not those infecting the roots (Cosgrove, 1967). This free phosphate would also be
available to heavy metal compounds to form metal phosphates. Cotter-Howells et al.
(1996) showed that the formation of metal phosphates could also be induced by the
biochemical action of the roots of Agrostis capillaries.ll1~~~111~~~11 However, the exact mechanism is
not clear. Agrostis capillaries~~~1111~~~~111 roots in these experiments were free from mycorrhizal
infection. Thus, root exudates containing phosphatase enzymes or non-root infecting
microorganisms could be responsible for the conversion of organic P to phosphate in the
Another study by Hinsinger and Gilkes (1997) indicated that additional dissolution
of PR occurred in the presence of plant roots. The largest root-induced dissolution was
achieved by ryegrass (Lolium rigidum) and rape (Bra~ssica napus), amounting to 19 to
32% of the PR present in the first two mm of the rhizosphere.
Phytoremediation is a general term for using plants to remove, degrade, or contain
soil pollutants such as heavy metals, pesticides, polyaromatic hydrocarbons, and landfill
leachates. These processes include (1) modifying the physical and chemical properties of
contaminated soil; (2) releasing root exudates, thereby increasing organic carbon; (3)
improving aeration by releasing oxygen directly to the root zone as well as increasing the
porosity of the upper soil zones; (4) intercepting and retarding the movement of
chemicals; (5) effecting co-metabolic microbial and plant enzymatic transformation of
recalcitrant chemicals; and (6) decreasing vertical and lateral migration of pollutants to
the ground water by extracting available water and reversing the hydraulic gradient
(Susarla et al., 2002). Recent researches have demonstrated that plants are effective in
cleaning up heavy metal contaminated soils (Wenzel and Jockwer, 1999; Pichtel et al.,
2000; Baker and Brooks, 1989). In many remediation proj ects, phytoremediation is seen
as a final polishing step following the initial treatment of the high-level contamination.
However, when contaminants are low in concentration, phytoremediation alone may be
the most economical and effective remediation strategy (Susarla et al., 2002). Plants have
been used to stabilize or remove metals from soil and water. The main methods used
include phytoextraction and phytostabilization (USDA, 2000).
Many higher plant species have adaptations that enable them to survive and to
reproduce in soils heavily contaminated with heavy metals. Such species are divided into
two main groups: the so called pseudometallophytes that grow on both contaminated and
non-contaminated soils, and the absolute metallophytes that grow only on metal
contaminated and naturally metal-rich soils (Baker and Brooks, 1989). Phytoextraction
involves growing plants in metal contaminated soil and harvesting the metal-rich plant
biomass, which can then be incinerated or composted to recycle the metals. Several crop
growth cycles are needed to decrease contaminant levels to allowable limits. If the plants
are incinerated, the ash must be disposed of in a hazardous waste landfill, but the volume
of such ash is far smaller than the volume of waste generated from direct soil
manipulation techniques (i.e. soil removal and incineration or chemical treatment)
Phytoextraction is often done with plants called hyperaccumulators, which absorb
unusually large amounts of particular metals in comparisons to other plants.
Hyperaccumulators can tolerate, uptake, and translocate high levels of certain heavy
metals that would be toxic to most other organisms. They are defined as plants whose
leaves may contain >100 mg kg-l of Cd, >1000 mg kg-l of Ni and Cu, or >10,000 mg kgl
of Zn and Mn (dry weight) when grown in a metal-rich medium. Hyperaccumulators of
Co (26 species), Cu (24), Mn (8), Ni (145), Pb (4), and Zn (14) have been reported
(Baker and Brooks, 1989).
Root induced changes in rhizosphere soil properties may have significant influence
on the mobility and bioavailability of nutrients and trace metals (Arienzo et al., 2003).
Phytostabilization is the use of perennial, non-harvested plants to stabilize or immobilize
contaminants in the soil and groundwater. This process takes advantage of the ability of
plant roots to alter soil conditions, such as pH and soil moisture content by exudation
(Susarla et al., 2002). Metals are absorbed in and accumulated by the roots, adsorbed onto
the roots, or precipitated within rhizosphere. Phytostabilization reduces the mobility of
the contaminant and prevents further movement of the contaminant into the groundwater
or the air and reduces its bioavailability for entry into the food chain (USDA, 2000). One
advantage of this strategy over phytoextraction is the disposal of metal-laden plant
material is not required. By choosing and maintaining an appropriate cover of plant
species, coupled with appropriate soil amendments, it may be possible to stabilize certain
contaminants in the soil and reduce the interaction of contaminants with associated biota
(Susarla et al., 2002).
In order to achieve a successful phytoremediation of soil polluted with metals, a
strategy of combining a rapid screening of plant species possessing the ability to tolerate
and accumulate heavy metals with agronomic practices that enhance shoot biomass
production and increase metal bioavailability in the rhizosphere must be adapted (Kamal
et al., 2004).
Lead occurs naturally in the earth's crust, but due to increased release of
anthropogenic Pb into the environment, contamination of soil by Pb constitutes a great
environmental concern. In particular, the ubiquity of Pb, its toxicity even in trace
quantities and tendency to bioaccumulate in the food chain make lead poisoning a leading
environmental health threat.
Presence of hazardous levels of Pb in the environment and the toxicity of Pb
requires remedial action to protect public health. Remediation technologies considered to
be cost effective and less disruptive, namely phosphorous induced Pb immobilization and
phytoremedation have been discussed in this review. This study was conducted for the
better understanding of phosphorous induced Pb immobilization mechanisms and the
feasibility of phytoremediation in Pb contaminated site.
EFFECTS OF APPLICATION METHODS ON P-INTDUCED Pb IMMOBILIZATION
IN A SOIL COLUMN
Activities such as mining and manufacturing, and the use of synthetic products (e.g.
pesticides, leaded paints, batteries, and industrial wastes have resulted in many Pb-
contaminated sites. Over time, the Pb loading rate in soil has exceeded its natural removal
rate by more than 20-fold (Nriagu, 1978). Contamination of soil by lead is of maj or
concern due to not only its high toxicity to humans and animals, but also to its ease of
exposure through ingestion or inhalation. A soil is generally considered contaminated
with Pb when its total Pb concentration exceeds 400 mg kgl (USEPA, 1996), and
remediation is required at this level (Ma and Gao, 1999).
Various remediation technologies have been developed to clean up metal
contaminated soils. Among those, in situ stabilization of heavy metals using binding
agents is a promising approach due to its sustainability and cost-effectiveness. Other
remediation technologies, including excavation, solidification and chelation/extraction,
are either very costly or only partially effective. The estimated cost of solidification is
about $750/m3 and that of stabilization $250/m3. Stabilization seems to be the more
feasible option due to its ease of operation and relatively low cost (Wang et al., 2001).
In situ immobilization of soil lead using phosphate has been considered a cost-
effective and environmentally benign remediation technology. When phosphorous reacts
with lead in soil, it transforms the reactive and bioavailable Pb species into a more stable
form. Pyromorphite is one of the least soluble forms of lead found in soils under a wide
range of environment conditions. The bioavailability and mobility of soil Pb can be
drastically reduced when unstable Pb forms, such as cerrusite (PbCO3), are COnverted into
pyromorphite (Zhang and Ryan, 1999a). During the last decade, many researchers have
successfully demonstrated the effectiveness of phosphate-induced Pb immobilization by
mixing phosphate minerals with Pb contaminated soils (Ma et al., 1993, 1995; Cotter-
Howells, 1996). The immobilization mechanism is considered to be the dissolution of the
lead compounds followed by the precipitation of lead phosphate. Thus, successful
immobilization of lead in soil requires enhanced solubility of soil Pb and P by decreasing
soil pH and applying sufficient phosphorous.
Among the various phosphorous sources (apatite, phosphate rock, phosphatic clay,
and soluble P), the feasibility of using phosphate rock (PR) to immobilize Pb in soils has
been rarely researched. Phosphate rock, a complex assemblage of phosphate minerals, is
mainly composed of microcrystalline carbonate fluorapatite (Ma et al., 1995). In addition
to reducing metal solubility, phosphate amendments are also effective at reducing metal
bioavailability associated with incidental ingestion of soil by humans (Basta and
The effectiveness of PR in immobilizing Pb primarily depends on its ability to
provide soluble P (Ma et al., 1993, 1995). To compensate for the low solubility of PR in
soil soluble phosphoric acid (PA) has been used in combination with PR, which
effectively immobilized Pb from a Pb-contaminated soil and facilitated precipitation of
Pb-phosphate compounds (Cao et al., 2001). The role of PA in the mixture was to
solubilize lead minerals in soils and PR, thereby increasing the readily available Pb and P
in the soil. This, in turn, facilitates more precipitation of Pb-phosphate compounds.
The obj ective of this study was to evaluate the effectiveness of different application
methods on lead immobilization in soil using PR and PA as P sources. This was
accomplished by (1) determining Pb leaching characteristics and bioavailability; and (2)
to evaluate Pb distribution in soil after P application.
Materials and Methods
The soil used for this study was collected from a lead-contaminated site in an urban
area of northwest Jacksonville, Florida. The site is located in a vacant, fenced rectangular
area (4, 100 m2), and covered by vegetation, mainly grasses. Past industrial activities,
which included a gasoline station, salvage yard, auto body shop, and the recycling of lead
batteries, have all contributed to the contamination of this site. Total lead concentrations
in the soil ranged from 36 to 21,074 mg kg l. Lead concentration decreased with soil
depth, with the maj ority of the Pb present near the soil surface (0-20 cm). Mineralogical
characterization of the site by x-ray diffraction (XRD) reveals that PbCO3 (COTUSsite) is
the predominant Pb mineral on the site (Cao et al., 2001).
Soil Sample Preparation
Soil samples were collected from the top 20 cm at the Jacksonville site. They were
air-dried, sieved through a 2-mm stainless steel screen and stored at room temperature.
The soil sample was thoroughly mixed to ensure uniformity. They were then digested
using the hot-block digestion procedure (USEPA Method 3050A) for total Pb
Contaminated soil samples were collected from a location where high
concentrations of lead are present (Cao et al., 2001). The clean soil sample was collected
from outside the fence directly adj acent to the site. The Pb content in the contaminated
soil was greater than 5,000 mg kg- and the soil collected from outside contained 77 mg
kg-l of Pb.
The column test was conducted to simulate field conditions. 400 g of soil was
packed into a PVC column (40 cm height by 3.5cm diameter), with the top half (0-20 cm)
being filled with the contaminated soil and the lower half (20-40 cm) the clean soil to
simulate Pb distribution on the site. The bulk density of packed soil was 1.2 g cm-3
A total of 18 columns were mounted with each treatment replicated three times.
The phosphate application rate was based on a P/Pb molar ratio =4, which was the
optimum rate for Pb-immobilization in the contaminated soil (Cao et al., 2001).
Phosphorous was applied half as PR and half as PA. Both PR and PA were applied only
to the first 20-cm of the soil column. The PR was either mixed (M) with the
contaminated soil or added as a layer (L) at the bottom of the contaminated soil portion of
the column, while the PA solution was added to the top of the column at 75% field
Five different P application methods were used in this experiment and they were as
RMAs1: P was added as '/ PR and '/ PA where PR was mixed with the
contaminated soil and PA was added from the top of the column at the same time;
R\I.An I: P was added as '/ PR and '/ PA where PR was mixed with the
contaminated soil and DI was added to the top of the column at 75% field capacity. PA
was added from the top of the column one week later;
RLAs1: P was added as '/ PR and '/ PA where PR was added as a l er at the
bottom of the contaminated soil and PA was added from the top of the column at the
R\I.111 2: P was added as '/ PR, '/ PA and '/ PA where PR was mixed with the
contaminated soil and PA was added from the top of the column two times with one week
RLAW2: P was added as '/ PR, '/ PA and '/ PA where PR was added as a la er at
the bottom of the contaminated soil and PA was added from the top of the column two
times with one week aart;
A fine textured glass wool was applied at the end of the column, acting as a filter
for the leachate. The soil columns were leached with deionized distilled water (DDW)
twice up to 2 pore-volumes 1 and 4 weeks after the application of PR.
Soil pH was measured with deionized water at a 1:1 soil:solution ratio after 24 h of
equilibrium. Total P was measured colorimetrically with a Shimadzu 160U spectrometer
using the molybdate ascorbic acid method.
At the end of the experiment, the soil columns were divided into 4 sections (0-10,
10-20, 20-30, and 30-40 cm). The modified physiologically based extraction test (PBET)
and Toxicity Characteristic Leaching Procedure (TCLP) were conducted on each section
to determine the effectiveness of P-induced Pb-immobilization in the soil. Also total Pb
in each soil layer was determined to compare the downward movement of Pb among
Contaminated soil 6.2 5,017 990 2,200
*Detection limit in soil concentration = 20 mg kg- (Pb); 5 mg kg- (Cu); 1 mg kg- (Zn)
The soil is very sandy, with a pH of 6.2-6.7, which is within the range typical of
Florida soils (Chen et al., 1999). Lead was the main contaminant with a concentration of
5,017 mg kg- which exceeds the critical level for industrial soils (1,400 mg kg- ). Also,
elevated concentrations of Cu and Zn were observed (990 and 2,200 mg kg- ,
respectively). The complete details of the metal contamination at this site have been
described by Cao et al. (2001). Generally, Pb, Cu and Zn were concentrated on the
surface soil (0-20 cm) and their concentrations decreased with soil depth. However, a
substantial amount of Pb (>2,000 mg kg- ) was found at depths below 30 cm. In the long
Total Pb Total Cu Total Zn
(mg kg-') (mg kg-') (mg kg- )
Background soil 6.7 77 20 195
different treatments. After leaching columns with up to 2 pore volumes of DDW twice,
column leachates were analyzed for Pb and P concentrations. Pb analysis was performed
using a flame atomic absorption spectrophotometer or graphite furnace atomic absorption
spectrophotometer, depending on the analytic needs of the samples. Soil clay fractions
were separated by centrifugation and analyzed using XRD to detect the formation of
pyromorphite in the soil.
Selected chemical properties of the collected surface soil (0-20 cm) are listed in
Table 2-1 Soil characterization at the contaminated site
run, it is possible that the metals may leach downward to the subsurface soil (Cao et al.,
Effects of P Application on Soil pH
The pHs of the P treated soils were measured to evaluate the degree of soil
acidification. As seen from Table 2-2, reduced soil pH in the surface soil (0-20 cm) was
observed in all the P treated soils due primarily to the addition of phosphoric acid
(H3PO4). Significant reduction in pH, approximately 1 pH unit, was limited to only the
top 10 cm.
Table 2-2 pH of soil after treatment
Soil layer(cm) Control R~MAsi RLAsi ,.1 RLAw2
0-10 6.2110.12 5.210.16 5.25+0.07 5.1210.17 5.2110.08 5.2610.06
10-20 6. 1610.08 5.7510.12 5.9910.07 5.8910.05 5.8710.02 5.8910.04
20-30 6.7810.02 5.9310.02 6.0810.04 6. 1210.06 6.3 510.01 6.4610.02
30-40 6.7410. 16 6.0710.06 6.1110. 01 6. 1810.05 6.4110.04 6.4610.06
Of all of the treatments, RLAs1 induced the greatest decrease in soil pH. The pH of
the surface soil was reduced from 6.21 in the control to 5.12 in treatment RLAs1, which is
typical of Florida soils (Chen et al., 1999). Yang et al. (2001) added 5,000 mg kg-l of P as
phosphoric acid to immobilize Pb and reported a reduction of soil pH by 3 units from
7.22 to 4.34. Reduction in soil pH was mostly limited to the top 20 cm of the soil column
(Table 2-2), indicating limited movement ofP in the soil profile (data not shown).
Without the application of phosphate rock, soil pH may have further decreased to levels
where increased metal solubility may become an issue.
Reduction of soil pH to near 5.5 is necessary for efficient Pb immobilization in soil.
Since a large fraction of the Pb in the soil is associated with carbonate (Cao et al., 2001),
addition of PA to the soil is essential for the dissolution of PR and Pb carbonate. Soluble
P and Pb may enhance precipitation of insoluble pyromorphite-like minerals in the soil,
which at given lead and phosphate concentrations is formed a greater rate at pH 5 than at
pH 6 or 7 (Laperche et al., 1996). Other researchers have used liming materials along
with phosphate treatments to buffer against a decrease in pH (Basta and McGowen, 2004;
Brown and Elliot, 1992). Also the application of phosphoric acid to acidify soil may
increase the risk of groundwater contamination with P and other heavy metals and,
therefore, caution should be exercised.
The toxicity characteristic leaching procedure (TCLP) is designed to evaluate
whether hazardous constituents may migrate through the vadose zone soils to the water
table by simulating landfill conditions. Higher Pb concentration in TCLP extracts means
higher Pb mobility in the soil. TCLP-Pb concentrations in the control soil without P
treatment were as high as 127 mg L^1, significantly exceeding the regulatory level of 5
mg L^1 (Fig. 2-1). This is possibly because most of the Pb is in the carbonate fraction,
which would readily dissolve in the acidic TCLP solution (Melamed et al., 2003).
Phosphate treated soils, on the other hand, reduced TCLP-Pb by >95% compared to the
control (Fig. 2-1). As a result of P application, TCLP-Pb in the surface soil (0-20 cm) was
reduced to below 5 mg L^1 for all five treatments (Fig. 2-1). .
Among the five P treatments, the highest reduction in TCLP-Pb was observed for
treatment RhlAs1 where PR and PA were applied at the same time, which had less than 2
mg L^1 in the 0-10 cm fraction of the soil column (Fig.2-1). On the other hand, treatment
RMAW2 where PA was added in two aliquots one week apart was the least effective in
reducing TCLP-Pb in the soil. This result may indicate that spontaneous reaction of PA
with PR to solubilize both Pb and P is essential in reducing TCLP-Pb in the soil.
Fig.2-1 TCLP metal contents in sectioned soil columns after leaching
*Detection limit = 0.2 mg L^1
* Data for lower half section (20-40 cm) are not shown due to no significant difference
Hettiarachchi et al. (2000) and Chen et al. (2003) reported the addition of P
significantly reduced TCLP- Pb in contaminated soil to below the critical level of 5 mg
Incidental ingestion of Pb-contaminated soil has been proposed as a primary
exposure pathway to humans for elevated blood Pb levels (Chen et al., 2003).
Physiologically based extraction test (PBET) has been used to estimate Pb bioavailability
(in vivo), which simulates Pb dissolution under gastrointestinal conditions using a
chemical extraction (Yang et al., 2001). Lead bioavailability in contaminated soils has
been shown to vary with its mineralogical forms (Davis et al., 1993). In vivo and in vitro
assays have indicated that the mammalian gastrointestinal availability of Pb is controlled
by the form and relative solubility of Pb solids (Ruby et al., 1996). Reduction in Pb
bioavailability was measured by PBET after various amounts and sources of P were
added to a Pb contaminated soil (Hettiarachchi et al., 2000).
The bioavailability of soil Pb is associated with its solubility and dissolution rate in
the gastrointestinal tract. The data in Fig. 2-2 indicate that bioavailable Pb in the
contaminated soil based on PBET measurements was reduced after P application in all
five treatments. The control soil showed 24-25 mg kg-l of bioavailable Pb while P-treated
soils showed reduction of PBET-Pb by 25-42%, which was similar to the 25-3 5%
reduction reported by Hettiarachchi et al. (2000) and 39% by Yang et al. (2001). The
greatest reduction was obtained from treatment RMAs 1 where PR and PA were added to
the soil at the same time, this result is similar to the TCLP results (Fig.2-1).
Distribution of Pb in Soil Column
In most contaminated soils, metals do not appear to leach downward in significant
quantities in the short run primarily due to their strong interactions with the soil.
However, in the long run, metals can leach downward in a soil due to their complexation
with solubilized organic mater especially in an alkaline environment where organic
matter is more soluble (Marschner and Wilczynski, 1991). This may be true at the
Jacksonville site where soil organic matter (3.91%) and pH (6.95) are much higher than
those of typical Florida soil (Chen et al., 1999).
S0-10 cani 10-20 cani
Co ntrol1 RMAS1 RMAW1 RLAS1 RMAW2 RLAW2
Fig. 2-2 PBET extractable Pb contents in sectioned soil columns after leaching
*Detection limit = 0.2 mg L^1
* Data for lower half section (20-40 cm) are not shown due to no significant difference
To determine the effectiveness of P treatment to prevent Pb downward migration,
total Pb concentrations in bottom half of the column (20-40 cm) were determined (Fig. 2-
3). As expected, significant downward movement of Pb was observed in the control soil,
especially in the 20-30 cm fraction, which was located directly under the Pb-
contaminated (0-20 cm). The initial Pb concentration in the lower half of the column (20-
40 cm), the control soil, was 77 mg kg-l (Table 2-1). After being in contact with the Pb-
contaminated soil for six weeks under 75% field capacity, the Pb concentration in the 20-
30cm-soil column fraction increased to 365 mg kg- However, Pb migration was mostly
limited to that section of the column as Pb concentration in the 30-40 cm fraction
remained at 77 mg kg-l and little difference was observed among the five treatments (Fig.
All treatments were effective in reducing Pb migration except RMAW2 where PR
was mixed with the contaminated soil in the top 20-cm, and PA was applied two times
with one week between applications. Since PA was added in two aliquots, the amount of
acidity was probably insufficient to dissolve enough Pb from the soil to induce the
precipitation of Pb-phosphate minerals. This was confirmed by the P data where P
concentration in the leachate of RMAW2 WAS the lowest (Fig. 2-4). This was also
consistent with the highest TCLP-Pb (Fig. 2-1) and PBET-Pb (Fig. 2-2) observed in
Treatments with PR as a layer (e.g. RLAs1, RLAW2) Were mOre effective in reducing
Pb migration than those where PR was mixed with the soil (RhlAs1, RhlAw1, and RMAW2,
Fig.2-3). This may be due to the formation and/or containment of stable Pb minerals
within the PR layer. Among the treatments where PR was mixed with the soil, treatment
RhlAs1, where PA was applied at the same time as when PR was mixed with the
contaminated soil, was most effective.
To determine the mechanism of reduction in Pb downward movement, soil around
the PR layer was separated and analyzed with XRD. Previous analysis of mineralogical
changes over time resulting from P addition using XRD showed formation of stable
pyromorphite-like minerals in P-treated soil (Cao et al., 2002). However, no Pb-
phosphate was identified in the clay fractions in this research (data not shown). With the
Pb concentration the studied soil being near the MDL for XRD the lack of lead-phosphate
5 20-30 can 5 30--10 can
confirmation cannot be ruled out. Further investigation is needed to verify the mechanism
within the PR layer.
Control RMAS1 RMAW1 RLAS1
Tre atme nts
Fig. 2-3 Total Pb concentration in sectioned soil columns (bottom section)
*Detection limit=20mg kg-l
* Data for upper half section (0-20 cm) are not shown due to no significant difference
All P-treated soil columns showed that leachable Pb had been reduced to below the
EPA drinking water regulatory level of 15 ClgL^1 or non-detectable (Table 2-3). The
second highest Pb concentration was observed in treatment RhlAW2, and was comparable
to the control soil. Lead concentrations in the second leachate were lower than those in
the first leachate. Significant differences were seen in the release of Pb between
treatments with different application methods of PA, where treatment RMAW2 WAS
observed to be the least effective and yet still below the EPA regulatory level (15 ClgL^)~.
All Hyve treatments proved to be effective in reducing Pb contamination from soil.
Table 2-3 Pb concentration (ppb) in column leachate (Clg L 1)
Treatment 1st leaching 2nd leaching
Control 33.611.5 14.511.74
RMAc1 nd nd
RMAs1 nd nd
RLAc1 nd nd
RMAs2 11.011.9 5.710.91
RLAS2 2.310.45 nd
*Detection Limit = 1 CgL^1
Column studies more closely simulate field conditions than the TCLP extraction
does. To simulate field conditions, soil columns were leached twice with DDW and the
leachate was collected after one and five weeks of incubation. The leachate was analyzed
for total Pb and P concentrations (Table 2-3, Figure 2-4).
The addition of large amounts of P to a contaminated soil may increase the risk of
eutriphication to water bodies (Basta and McGowen, 2004). Taking into account P
leaching in the soil after P application is important with respect to secondary
contamination. Although it is well recognized that phosphate amendment is an effective
method for immobilizing metals in contaminated soils, some elevated soluble
phosphorous may enhance eutrophication risk (Cao et al., 2001).
1. 5Ist letiching 2ndl letiching
Control RMAS1 RMAW1~ RLAS1 RMAW2 RLAW2
Tre atme nts
Fig. 2-4. P concentration in column leachate (mg L^1)
*Detection limit = 1 CgL^
This was not the case in this study as phosphorous leached from the P-treated soil
columns was less than 1 mg L^1 for all treatments (Fig. 2-4). Treatments where PA was
applied in two aliquots (RMAW2, RLAW2) had lower P concentrations in the leachate than
in the other treatments. This may indicate that the application of P in small quantities can
limit the downward movement of PA, but with correspondingly lower efficiency in
immobilizing Pb in the soil (Figures 2-1 and 2-2).
Though phosphorous release from P-treated soil was minimal, some of the P was
released as it was not converted to lead phosphate. Therefore, caution is needed to assess
the P movement within P-treated fields to prevent eutrophication risk, and consideration
given to soil type and the amount of P added.
The efficiency of in situ P-induced Pb immobilization depends on the type and rate
of P amendment, along with appropriate application methods. Preliminary laboratory
studies to assess its mechanism and efficiency are important.
This study has shown that applying a mixture of PA and PR effectively reduced
TCLP-extractable Pb, bioavailable Pb and vertical Pb migration in the Pb-contaminated
soil studied. All P-treated soil resulted in significant reductions in bioavailable and
leachable Pb as compared to the control soil. The effective immobilization of Pb was
attributed to the formation of stable Pb minerals after P application. Of all the treatments,
the mixture of PR and soil coupled with the simultaneous application of PA (RhlAs1) was
the most effective in decreasing leachable and bioavailable Pb with the least impact on
soil pH and lowest eutrophication risk. Treatments with PA being applied in two aliquots
and PR being applied as a layer was the least effective overall. Possible explanations of
these results are: 1) the reaction between soil and P amendments before PA application
(after one week) inhibited the formation of lead phosphate; 2) less PR was mixed with
soil (RLAs1 and RLAW2), which caused lower efficiency in reducing leachable and
bioavailable Pb in soil. These results also indicate that a minimal amount of
simultaneously applied PA with the PR is necessary for effective (maximal)
immobilization of Pb in soil to take place. The fact that layered PR showed lower
efficiency in Pb immobilization as compared to the other treatments may be due to an
insufficient amount ofP supplied to the soil. However, the PR layer showed improved
reduction of Pb migration than other mixture treatments. Furthermore, these results
suggest layered PR below contaminated soil may serve as a reactive barrier to prevent Pb
from migrating into the groundwater. Although the effects of PA on pH and leachibility P
were acceptable in this experiment, competitive heavy metal leaching and eutrophication
may be potential drawbacks of its indiscriminate utilization.
Further studies are needed to determine the mechanism of Pb-migration reduction
by the PR layer. Also, the combination of the soil-PR mixture and PR layer can be
considered as well for remediation of Pb contaminated soils, and with reduced leaching,
bioavailability and mobility of Pb.
THE EFFECTS OF PLANTS ON PHOSPHATE-INDUCED Pb IMMOBILIZATION INT
THE RHIZOSPHERE SOIL
Of the various technologies for the remediation of lead contaminated sites,
phosphate-induced Pb immobilization has received much attention in recent years. Lead
phosphates have low solubility, and are generally several orders of magnitude less
soluble than analogous carbonates and sulfates (Cotter-Howells, 1996).
Various forms of phosphate have been used to immobilize Pb in situ from
aqueous solutions and Pb contaminated soils (Ma et al., 1995; Hodson et al., 2000; Yang
et al., 2001). The primary mechanism of Pb immobilization appears to be through
phosphate mineral dissolution and subsequent precipitation of pyromorphite-like
minerals. Hence, the solubility of phosphate-mineral applied largely determines the
effectiveness of in situ Pb immobilization in soils (Ma, 1996). For soil P to be available
to react with Pb to form lead phosphate, it must be present as free phosphate (HnPO4n-3,
where n=1 to 3) in the soil solution. However, phosphate minerals in soils also have
relatively low solubilities. In a study on Pb immobilization using phosphorous, a mixture
of phosphoric acid (PA) and phosphate rock (PR) as the sources of P were applied to
enhance solubilization of P and Pb-carbonate in soil (Cao et al., 2001).
Phosphorous is one of the essential macronutrients, and among the most limiting
nutrients for plant growth in soils. The average P-content in soil is about 0.05 % (w/w)
but only 0. 1% of that P is available to plants (Illmer and Scinner, 1995). It has been
shown that organic acids can greatly increase the concentration of plant-available P in the
soil solution through acidification, chelation and anion exchange (Reyes et al., 2001).
Root exudates containing organic acid or phosphatase enzymes play an important role in
the liberation of phosphate ions from organic and inorganic compounds in soils (Cotte-
Howells, 1996). Through roots-induced chemical modifications in the rhizosphere, higher
plants may be directly responsible for the dissolution of PR. In particular, rhizosphere pH
has been shown to differ by 2 units as compared to that of the bulk soil. Several
researchers have concluded that excretion of protons or organic acids by the roots is a
maj or process by which plants can acquire P from PR present in soil (Hinsinger and
Gilkes, 1997). Free phosphate in the soil solution and available to plant roots may be
available to react with lead and form lead phosphate.
The overall obj ective of this experiment was to determine the effectiveness of
phosphate-induced Pb immobilization in the rhizosphere soil of different plant species.
Specific obj ectives were, 1) to evaluate the effect of plant roots on P-induced reduction of
Pb bioavailability; and 2) to determine the effect of plant roots on formation of stable Pb
Materials and Methods
Surface (0-20 cm) soil samples were collected from a heavy-metal contaminated
site in Jacksonville, Florida. Soils were air-dried for 2 weeks, sieved through a 2-mm
stainless steel screen and stored at room temperature prior to the experiment. The soil
sample was mixed thoroughly to ensure uniformity. Soil pH was measured using a 1:1
soil:water ratio. Soil samples were digested using the hot-block digestion procedure
(USEPA Method 3050a) for total lead concentration. Ground rock phosphate [PR,
primarily Calo (PO4)6F2] and PA were used as the sources of phosphorous.
Three plant species were selected to determine the effects of plant root exudates
on phosphate-induced Pb immobilization in the rhizosphere soil. The first, Agrostis
capillariesll1~~~111~~~11 has proven to be effective in inducing the formation of metal-phosphates
(Cotter-Howells, 1996). The other two, Lolium rigidum and Bra~ssica napus, may also
enhance root-induced PR dissolution in the rhizosphere (Hinsinger and Gilkes, 1997).
The following treatments were considered:
Control 1: Soil + Phosphate rock (PR)
Control 2: Soil + PR + Phosphoric acid (PA)
Treatment 1: Soil + PR + Plant
Treatment 2: Soil + PR + PA + Plant
Pots of 4 inches in diameter were used and filled with 520 g of the Pb-
contaminated soil collected from Jacksonville. 10-20 plants seeds were sown into each
pot. The plant seeds were sown directly onto the surface soil, with the rhizosphere soil
separated from the bulk soil. This was accomplished by using a circle of plastic netting
9.5cm in height and 5.5cm in diameter, which was covered with a nylon mesh cloth
(mesh size 45 Clm) placed in the center of the pot, and filled with what would become the
rhizosphere soil. This limited root growth to this area of the pot. The plants (10 plants
per pot) were grown in a greenhouse for 4 weeks (plant heights were approximately 10
cm) and watered as necessary. At the end of four weeks, the rhizosphere soil was
separated from the bulk soil. The roots were washed gently to remove the adhered soil
particles and placed in a glass vial containing approximately 10 ml of deionized distilled
water (DDW). This container was then ultrasonicated for 30 seconds to remove any
remaining soil particles. The resulting solution in the glass vial was filtered using a 45
Clm filter paper and collected. The residues on the filter containing both soil and plant
root samples were mounted on a carbon stub for further analysis.
Fig. 3-1 Lolium rigidum after 4 weeks of growth
Fig. 3-2 Agrostis capillaries~~~1111~~~~111 after 4 weeks of growth
Examination of the rhizosphere soil was conducted with a scanning electron
microscope (SEM) to investigate formation of Pb-P minerals (i.e. pyromorphite). The
presence of pyromorphite like minerals was further analyzed by X-ray diffraction (XRD).
For XRD analysis, the clay fractions of the soil samples were separated and
Fig. 3-3 Bra~ssica napus after 4 weeks of growth
Also PBET (physiologically-based extraction test) was conducted on the
rhizosphere soil to see the effects of plant roots on bioavailable Pb in the soil.
The lead contaminated soil had a pH of 6.2 (Table 3-1), which was high compared
to the average value for Florida soils (pH 5.04) (Chen et al., 1999). Lead was the main
contaminant at the site, with a concentration of 5,017 mg kg- In addition, elevated
concentrations of Cu and Zn were also observed (990 and 2,200 mg kg-l each).
Table 3-1. Selected characteristics of the lead-contaminated soil
Soi pH Total Pb (mg Total Cu (mg Total Zn (mg
Soi pHkg-') kg-') kg- )
6.2 5,017 990 2,200
*Detection limit in soil concentration = 20 mg kg-l (Pb); 5mg kg- (Cu); Img kg-l (Zn)
In vitro metal bioavailability is an estimate of metal bioavailability to humans and
animals by simulating metal dissolution under gastrointestinal condition using a chemical
extraction (Yang et al., 2001).
30 5 AgolYS tiS Cap illar Iie s
O Bnissica l naipus
25 O Loliumll rigidumll
PR only PR+PA
Tre atme nt
*Detection limit = 0.2 mg L^1
Fig. 3-4. Lead concentrations using the physiologically based extraction test in the
rhizosphere soil with phosphoric acid and with phosphate rock treatment.
Bioavailable Pb in the rhizosphere soil as measured by PBET was reduced in the
presence of all three plants when growing in soils treated with PR (Fig. 3-4). The
rhizosphere soil in the control had 26 mg kg-l of bioavailable Pb while those with plants
had 21-23 mg kg- an 11-20% reduction. Such a reduction in the bioavailable Pb in the
rhizosphere soil may result from the formation of Pb phosphate, which was supported by
the SEM observation of the plant roots (Figs. 3-5, 6, 7). In the treatment with PR+PA, no
significant reduction of bioavailable Pb was found.
Formation of Lead Phosphate
The main mechanism of Pb immobilization in soil is via dissolution of P and/or
meta-stable Pb compounds and the subsequent precipitation of pyromorphite-like
minerals (Cao et al., 2001). The analysis of mineralogical changes in the rhizosphere soil
that were affected by root exudates using SEM helps to determine whether root-induced
formation of Pb phosphate occurred. The SEM element dot maps of root samples after
growing in a soil for 4-weeks show (Figs. 3-5, 6, 7) close association of Pb with P and
Ca, which may suggest the formation of lead phosphate in the rhizosphere soil
For A. capillariesll1~~~111~~~11 no significant associations between Pb with P and Ca were
observed in the SEM elemental dot map (Fig. 3 -5), but broad association of the three
elements may indicate Pb phosphate in the phosphate rock treated rhizosphere soil (Fig.
3-5b,c,d). For B. napus, the SEM data suggests a close association between Pb, P and Ca.
(Fig 3-6c, d, and e). For L. rigidum, association of P and Pb was observed but not a
strong association with Ca.
The association of Ca, P and Pb may suggest that Pb precipitation occurred on the
surface of the phosphate rock particle, mainly as Calo (PO4)6F2 (CRO et al., 2001).
Formation of stable lead-phosphate minerals may be enhanced due to the biochemical
actions of plant roots, i.e. root exudates containing phosphatase or organic acid. In XRD
analysis, no pyromorphite-like mineral was identified in the clay fractions of the
rhizosphere soil (data not shown). However, this is because the Pb concentration in the
soil was barely detectable by XRD, and the lack of confirmation of this technique does
not prove that no pyromorphite was formed. However, root induced phosphorous
dissolution is proved by analyzing water-soluble phosphorous in the rhizosphere soil. As
seen in Table 3-2, higher water soluble phosphorous was observed in the soil with plants
growing as compared to the control soil. This increase ranged from 37-46% in the PR-
only treated soils to 8-9% in the PR+PA treatment. These results indicate that more
phosphorous solubilization occurred in the presence of plant roots and that higher
available phosphorous may enhance the formation of stable lead-phosphates in the
Root exudates containing phosphatase enzymes and/or root-infecting
microorganisms may be responsible for this dissolution of P and precipitation of lead
phosphates. Another possible explanation for lead phosphate formation in the
rhizosphere soil may be that the plant roots created localized acidity, which may have
enhanced the dissolution of the phosphorous minerals and caused precipitation on the
root surface (Cao et al., 2002). Soil pH was analyzed to evaluate the effects its reduction
by root exudates on phosphate-induced Pb immobilization. Numerous studies have
shown that soil pH may also be critical in the dissolution of PR in soil (Ma et al., 1993,
1995; Singh et al., 2000). The dissolution of PR may be enhanced by the supply of
protons and removal of dissolution products in particular Ca and P. Acidic soils are likely
to yield an extensive PR dissolution because they often combine a low pH, with a high P
sorption capacity due to the abundance of Al and Fe oxyhydroxides (Hinsinger and
Gilkes, 1997). As seen in Table 3-3, slightly reduced soil pH was observed in the
rhizosphere soils treated with PR but no significant differences were found in those
treated with the PR+PA mixture. Solubilization of phosphorous is thought to be caused
by the release of organic acids and/or phosphatase, which has been reported to be
exclusively acid (George et al., 2002). The reduction of pH induced by root exudatation
(i.e. organic acid, phophatase) may enhance the dissolution of PR and Pb-carbonate,
promoting the formation of lead phosphate. Several authors have concluded that proton
excretion by the roots was the maj or process by which plants such as buckwheat, rape or
various legumes can acquire P from PR (Bekele et al., 1983; Hinsinger and Gilkes, 1997).
Youssef and Chino (1989) have shown that a single species of rape can strongly acidify
its rhizosphere when grown in an alkaline soil.
IMArGE, 255 Si~jKa, 28 IPKa, II
PbMal, 12 CllaKa, 11
Fig. 3-5. Scanning electron microscopy elemental dot map of Agrostis capillaries~~~1111~~~~111
Hoffland et al. (1989) considered the dissolution of PR by rape was due mainly to
the excretion of organic acid by the roots. Significant reduction of bioavailable Pb and
pH, and higher PR solubilization in soil was observed only in PR-treatments whereas no
significant differences in bioavailable Pb was found for soils treated with PR+PA. It is
possible that the addition of PA provided both the necessary acidity and soluble P to the
plants so no effect from plant roots was observed.
Control 0. 1710.01 5.7911.16
Agrostis capillaries 0.3210.05 6.3110.51
Lolium rigidum 0.2810.04 6.3610.57
Brassica napus 0.2710.04 6.3210.55
Fig.3-6. Scanning electron microscopy elemental dot map of Bra~ssica napus
II~wlAR 755 IlblKa 11 ||SiKa 24
Fig. 3-7. Scanning electron microscopy elemental dot map of Lolium rigidum
Table 3-2 Water soluble phosphorous in plant rhizosphere soil (ppm)
Table 3-3 pH of rhizosphere soil treated with phosphate rock after 4-weeks growth
PR Only PR+PA
Control 5.7210.21 5.0110.23
Agrostis capillaries 5.3110.41 4.9710.34
Lolium rigidum 5.3410.07 4.8910.43
Brassica napus 5.2310.13 4.9010.23
This study has shown that all three plant species had the ability to induce the
formation of lead phosphate in the rhizosphere soil as supported by the SEM analysis of
plant root samples (Figs. 3-5, 6, 7). Association between Pb, P and Ca in the SEM dot
maps suggests formation of lead phosphates induced by the biochemical action of the
plant roots. This experiment established that the biochemical effects of plant roots can
promote additional dissolution of phosphorous minerals and formation of lead
Reduction of the rhizosphere soil pH may indicate that plant roots enhance the
dissolution of PR and Pb-carbonate by excreting H thereby facilitating the formation of
lead phosphate. Reduction in bioavailable Pb in the rhizosphere soil may due to enhanced
dissolution of PR and formation of lead phosphate by biochemical reaction in the
rhizosphere soil. Using various leachates to examine the bioavailability of Pb in a
contaminated soil, Rabinowitz (1993) found a 43% decrease in Pb concentration
extracted with 10% citric acid in phosphate-amended soils. Our results indicate that all
three plants may have the potential to be applied for the reclamation of Pb-contaminated
sites in combination with the application of PR. Also soil microorganisms are involved in
this process and their role in the solubilization of phosphate-bearing materials has been
the subj ect of m any studi es. There i s a con si derable numb er of pho sphate- solub ili zi ng
bacteria present in soil and in the rhizospheres of plants, and their concentration much
higher in the rhizosphere as compared to non-rhizosphere soil (Rodriguez and Fraga,
1999). The combination of plant root exudates and microorganisms may increase the
amount of free phosphate, thus enhancing the formation of lead phosphate in the
rhizosphere soil. However, the extent to which either of these factors could re-solublize
heavy metal phosphate for subsequent phosphate extraction is unknown. Further research
should investigate the stability of the phosphate dissolution mechanisms of both plants
and soil microorganisms.
Vegetative techniques have been widely used in the reclamation of mine waste,
mainly to stabilize the soil and promote nutrient cycling. This work suggests that the
growth of plants, which has the ability to induce formation of insoluble lead phosphates,
can enhance the reclamation of a contaminated site. The successive growth and decay of
species over several seasons could eventually produce enough lead phosphates (Cotter-
Howells, 1996) to achieve this end. Substantial amounts of pyromorphite present in soils
can reduce bioavailable Pb. The increased effectiveness in phosphate-induced Pb
immobilization as a result of plant root activity would be more ecologically acceptable
than the addition of a large amount of soluble phosphate alone because of less concern
with soil acidification and eutrophication. However, additional research is needed to
further assess the effects of plant-root enhanced phosphate-induced Pb immobilization in
soil on a field scale.
SCREENING OF PLANTS FOR ACCUMULATION OF Pb, Cu, Zn FROM A
Heavy metals are one of the most serious environmental concerns today. They are
harmful to humans and animals and tend to bioaccumulate in the food chain. Past usage
of fossil fuels, mining and smelting of metal ores, industrial emissions and the application
of insecticides and fertilizers all have contributed to elevated heavy metal levels in the
environment. The threat that heavy metals pose to human and animal health is aggravated
by their elemental nature and long-term persistence.
There are many technologies available for the remediation of land contaminated by
heavy metals (USDA, 2000). However, many of these technologies are costly (e.g.
excavation of contaminated material and chemical/physical treatment) or do not achieve
either a long term or aesthetic solution (Cotter-Howells, 1996). A new approach, termed
phytoremediation, offers an alternative technology. Phytoremediation can provide a cost-
effective, long-lasting and aesthetic solution for the remediation of contaminated sites.
One of the strategies of phytoremediation with respect to metal contamination is
phytoextraction, i.e. the uptake and accumulation of metals into plant shoots, and their
Metal accumulating plants are not only of scientific interest, but they also serve to
protect several aspects of the environment, such as phytoremediation of lands polluted
with heavy metals, long-term stabilization of wastelands and the reduction of potentially
toxic elements (Raskin and Enseley, 2000). There are over four hundreds plants known to
be hyperaccumulators of metals, which accumulate high concentrations of metals into
their aboveground biomass. These plants include trees, vegetable crops, grasses and
weeds. Based on Baker and Brooks (1989), hyperaccumulators are defined as plants that
accumulate >1,000 mg kg-l of Cu, Co, Cr, Ni or Pb, or >10,000 mg kg-l of Mn or Zn.
Hyperaccumulators of Co (26 species), Cu (24), Mn (8), Ni (145), Pb (5), and Zn (4)
have been reported. The five hyperaccumulators of lead include Armeria martimza
Thlaspi rotundifolium, Thlaspi alpestre, Alyssum wulfenianum, and Polycalpaea
It is important to use indigenous native plants species for the phytoremediation
process as these plants are better adapted for survival, growth and reproduction under the
particular environmental stresses encountered than plants introduced from another
environment. However, few studies have evaluated the phytoremediation potential of
natural hyperaccumulators under field conditions (McGrath and Zhao, 2003).
This study was conducted to screen plants that were growing in a contaminated site
for their potential in accumulating Pb, Cu and Zn. The overall objectives were: 1) to
determine the concentrations of Pb, Cu and Zn in the plant biomass; 2) to compare metal
concentrations in the aboveground biomass to those in the roots and in the soils; and 3) to
assess the feasibility to use these plants for the purpose of phytoremediation The
information obtained from this study should provide insight for the use of native plants to
remediate metal contaminated sites.
Materials and Methods
The plant and soil samples for this study were collected from a known lead contaminated
site located at an urban area of northwest Jacksonville, Florida. The fenced vacant site is
rectangular in shape, occupying approximately one acre. The site is covered by
vegetation, mainly grasses. Past industrial activities include being home to a gasoline
station, salvage yard, auto body shop, and recycler of lead batteries, which all
presumably, contributed to its contamination. Total lead concentration in the soils ranged
from 90 to 4,100 mg kgl (Fig. 4-1 and Table 4-1).
0 10 20 30 40 5
In~r addtio toP otmntotest locotie lvtdlvl fC n
Zn ie.2090 nd19-,20 g g-(abe -1.Th cntmiaio b hay etlsi
conentate intetp2 mo h st dt o hw)
Sample Preparation and Chemical Analysis
Plant samples, together with associated soil samples were collected from the site.
The selection of plant samples was based on their coverage at the site. In total, 36 plant
samples of 17 different species were collected from 10 locations at the site in December
of 2002 (Fig. 4-1). Soil samples from the rooting zone (0-20 cm) were taken from each
Soil samples were air-dried at room temperature for two weeks, and then sieved
through a 2-mm stainless steel sieve. They were then digested using the hot-block
digestion procedure (USEPA Method 3050a) for total metal concentration. For water-
soluble Pb concentration, 25 ml of deionized distilled water (DDW) was mixed
thoroughly with 2 g of soil for 12 h. The mixture was then centrifuged for 15 min. at
3,500 rpm. The supernatant was analyzed for Pb by flame atomic absorption
Plants samples were divided into roots and shoot, washed gently with DDW for
approximately 1 minute to remove the soil. After washing, plant samples were air-dried
at room temperature for two weeks. They were then ground to a powder before digestion
via USEPA Method 3050a.
Metal Concentrations in Soils
Characteristics of the soil samples collected from this study are shown in Table 4-1.
Soil pH ranged from 6.62 to 7.20, relatively high compared to typical Florida soils (Chen
et al., 1999). The high soil pH has been attributed to the presence of cerrusite (PbCO3),
the predominant form of the Pb minerals at the site (Cao et al., 2001).
Table 4-1. Selected properties of soil samples collected from the contaminated
site at Jacksonville, Florida
Water soluble Total Pb Total Cu Total Zn
Site # Soil pH Pb (mg L ) (mg kg-') (mg kg-') (mg kg- )
1 7.03 1 90 20 200
2 7.00 3 143 21 195
3 6.62 32 4,100 990 2,200
4 7.06 82 1,375 980 900
5 7.2 65 1,886 860 683
6 7.08 20 767 314 551
7 6.95 234 2,405 746 1,000
8 6.97 96 1,451 300 572
9 7.01 2 333 29 532
10 6.63 1 145 26 720
*Detection limit = 0.1Img kg- (Pb); 5mg kg- (Cu); and Img kg- (Zn)
Total lead concentrations in the soil samples collected from 10 different locations
were variable, ranging from 90 mg kg-l at site 1 to 4,100 mg kg-l at site 3 (Fig. 4-1). The
mean Pb concentration in Florida soils is 77 mg kg- The global baseline level of Pb in
natural surface soils is reported to be 20 mg kg-l (Chen et al., 1999). In addition to total
Pb, water-soluble Pb was also determined (Table 4-1). As expected, they were much
lower than total Pb concentrations. However, total Pb correlated with water-soluble Pb (r
= 0.48). There were also elevated concentrations of Cu and Zn, ranging from 20 to 990
mg kg-l for Cu, and 195 to 2,200 mg kg-l for Zn. Metal concentrations in the soil
samples collected from different locations were highly correlated with r2= 0.72-90, i.e. a
site that had a high Pb concentration also tended to have high Zn and Cu concentrations.
Among the 10 locations sampled, sites 3, 4, 5, and 7 were the most contaminated with all
three metals (Table 4-1).
Metal Concentrations in Plants
A total of 17 plant species were collected from 10 locations at the site. They were
then identified and analyzed for heavy metal (Pb, Cu, and Zn) concentrations in their
biomass. Concentrations of Pb, Cu and Zn in the soil and in the plant biomasses are listed
in Tables 4-2, 3, and 4. It is generally agreed that metal concentrations in plants vary with
plant species (Alloway, 1994; Alloway et al., 1990). Plant uptake of heavy metals from
soil occurs either passively with the mass flow of water into the roots or through active
transport, crossing the plasma membrane of roots epidermal cells. Under normal growing
conditions, plants can potentially accumulate certain metal ions an order of magnitude
greater than the surrounding medium (Kim et al., 2003).
Lead concentrations in the plants ranged from non-detectable to as high as 1,183
mg kg- with the maximum value found in the roots of Phyla nodiflora collected from
site 4 (Table 4-2). In addition, other species such as Bidens alba, Rubus frutocosus and
Gentianapennelliana also contained significantly higher Pb concentrations than the rest
plant species. None of the plants accumulated Pb at 1,000 mg kg-l in the aboveground
biomass, which is the criterion for a Pb hyperaccumulator (Baker and Brooks, 1989). In
95% of the plant samples, the roots Pb concentration was much greater than that of the
shoots, indicating a low mobility of Pb from the roots to the shoots and immobilization of
heavy metals in roots. No relationship between water soluble Pb and plant-boimass Pb
was found. Pichtel et al. (2000) also analyzed Pb in native plants, which was collected
from a dumpsite and got similar results (non-detectible to 1,800 mg kg- ). Stoltz and
Greger (2002) reported a range of 3.4-920 mg kg-l in different plant species collected
from mine trailing.
Copper concentrations in the plants varied from non-detectable to 460 mg kg-l
(Table 4-3), with the maximum value found in the roots ofP. nodi~ora. Similar to Pb, no
plant species accumulated Cu above 1,000 mg kg- the criterion for a Cu
hyperaccumulator (Baker and Brooks, 1989). As expected, the Cu concentrations in the
roots were greater than those in the shoots. In addition to P. nodi~ora, G. pennelliana,,
Bidens alba also contained significantly higher Cu concentrations than other plant
species. Copper concentrations of 6.4-160 mg kg-l in the plant biomass were reported by
Stoltz and Greger (2002), which were lower than those in our research. Shu et al. (2002)
reported Cu concentrations of 8-45 mg kg-l in the biomass ofPaspahtna distichunt and
The Zn content in the plants ranged from 17 to 598 mg kg-l (Table 4-4), with the
maximum value found in the roots of P. nodi~ora collected from site 4, which also
contained the highest Pb and Cu concentrations in the roots. Similar to Pb and Cu, no
plant species accumulated Zn at 10,000 mg kg- the criterion for a Zn hyperaccumulator
(Baker and Brooks, 1989). Generally, Zn concentrations were greater in the roots than the
shoots. Al so Paspahtna notatuns, G. pennelliana, Bidens alba and Stenotaphrunt
secundatunt showed significantly higher Zn concentrations than the rest of plant species
collected. Research conducted by Stoltz and Gerger (2002) found Zn concentrations of
68-1,630 mg kg-l in the plants they collected while those by Shu et al. (2002) contained
66-1,015 mg kg-l
Phyla. nodi~ora collected from site 4 accumulated the highest Pb, Cu and Zn in the
roots (1183, 460, and 598 mg kg- respectively). Other than P. nodi~ora, G. pennellian2a
and Bidens alba had higher concentrations of heavy metals (Pb, Cu and Zn) than other
species collected from the contaminated site. No species satisfied the criterion for a heavy
metal hyperaccumulator, but ability of the above species to accumulate heavy metals in
their biomass needs further investigations. Although total metal concentrations in soils
play an important role in the uptake of metals by plants, this can be influenced by several
factors. In general, a negative correlation was found between soil pH and metal
concentrations in the plants. Other soil factors such as cation exchange capacity also
influence metal uptake by plants in some cases (Jung and Thornton, 1996).
Accumulation and Translocation of Metals in Plants
In this study, none of the plant species showed metal concentrations >1,000 mg kgl
in the shoots (Tables 4-2, 3, 4), i.e. none of them are hyperaccumulators based on Baker
and Brooks (1989).
Bioconcentration factor (BCF) and translocation factor (TF) can be used to estimate
a plant's potential for phytoremediation effectiveness. A plant' s ability to accumulate
metals from soils can be estimated using BCF, which is defined as the ratio of metal
concentration in the roots to that in the soil. A plant' s ability to translocate metals from
the roots to the shoots is measured using TF, which is defined as the ratio of metal
concentration in the shoots to that of the roots. Phytoextraction is the removal of a
contaminant from soil, groundwater or surface water by live plants. Enrichment occurs
when a contaminant taken up by a plant, is not degraded rapidly or completely
accumulated in the plant. The process of phytoextraction generally requires the
translocation of heavy metals to the easily harvestable plant parts, i.e. the shoots.
Table 4-2. Lead concentrations in soil and plant samples (mg kg-1) collected from the
contaminated site at Jacksonville, Florida
Common name Scientific name Site # Roots Shoots Soil
Bahia grass Paspalum notatum Flugge 4 575 428 1,375
5 397 92 1,886
9 nd* nd 333
Wire grass Gentianapennellia~na Fern 1 968 453 90
8 881 491 1,451
Romerillo Bidens alba (L.) DC 2 947 91 143
3 149 23 4,100
5 660 77 1,886
Bermudagrass Cynodon dactylon (L.) DC 5 293 88 1,886
6 75 52 767
Flatsed ge Cyperus esculentus L. 1 28 18 90
2 16 26 143
8 417 26 1,451
Ticktrefoil Desmodium pan2iculatum (L.) 2 130 20 143
Horsetail Equisetum arvense L. 3 284 38 4,100
Hydrocotyle Hydrocotyle amnericana L. 8 99 8 1,451
10 nd nd 145
Turkey tangle Phyla nodiflora (L.) Greene 1 44 24 90
4 1,183 73 1,375
7 117 83 2,405
7 451 55 2,405
Plantain Plantago major L. 1 9 52 90
5 294 67 1,886
Blackberr Rubus frutocosus L. 3 825 22 4,100
6 127 12 767
9 51 nd 333
Goldenrod Solidago altissima L. 6 59 49 767
10 nd nd 145
S owthi stl e Sonchus asper (L.) Hill 7 146 39 2,405
St. Augustine Stenotaphrum secundatum 1 3 1 14 90
3 68 32 4100
Bluej acket Tradescantia ohiensis Raf: 5 206 140 1,886
Tuberous Verbena rigid Spreng. 1 23 11 90
8 35 11 1,451
Bigpod Sesbania exaltata (Raf.) Cory 2 150 nd 143
*Detection limit for Pb 0 .1mg kgl
Table 4-3. Copper concentrations in soil and plant samples (mg kg- ) collected from the
contaminated site at Jacksonville, Florida
Common name Scientific name Site # Roots Shoots Soil
Bahia grass Paspalum notatum Flugge 4 250 352 980
5 360 60 860
9 10 11 30
Wire grass Gentianapennellia~na Fern 1 432 200 20
8 375 210 300
Romerillo Bidens alba (L.) DC 2 10 8 21
3 44 17 990
5 400 32 860
Bermudagrass Cynodon dactylon (L.) DC 5 310 52 860
6 36 21 314
Flatsed ge Cyperus esculentus L. 1 16 10 20
2 10 28 21
8 150 20 300
Ticktrefoil Desmodium paniculatum (L.) 2 6 6 21
Horsetail Equisetum arvense L. 3 110 23 990
Hydrocotyle L. Hydrocotyle amnericana L. 8 32 13 300
10 21 16 26
Turkey tangle Phyla nodiflora (L.) Greene 1 3 1 14 20
4 460 20 516
7 nd* 47 746
7 180 23 746
Plantain Plantago major L. 1 23 10 20
5 150 27 860
9 24 12 29
Blackberr Rubus frutocosus L. 3 30 46 990
6 65 13 314
9 47 265 29
Goldenrod Solidago altissima L. 6 277 241 314
S owthi stl e Sonchus asper (L.) Hill 7 46 34 746
St. Augsie ras Stenotaphrum secundatum 1 22 17 20
3 42 15 990
Bluej acket Tradescantia ohiensis Raf: 5 194 117 860
Tuberous vervain Verbena rigid Spreng. 1 14 10 20
8 25 18 300
Bipo sesbania Sesbania exaltata (Raf.) Cory 2 12 48 21
*Detection limit for Cu=5mg kg- .
Table 4-4. Zinc concentrations in soil and plant samples (mg kg l) collected from the
contaminated site at Jacksonville, Florida
Common name Scientific name Site # Roots Shoots Soil
Bahia grass Paspalum notatum Flugge 4 450 316 900
5 260 166 683
9 250 200 532
Wire grass Gentianapennellia~na Fern 1 524 250 200
8 310 359 572
Romerillo Bidens alba (L.) DC 2 25 20 195
3 143 230 2,200
5 462 17 683
Bermudagrass Cynodon dactylon (L.) DC 5 244 171 683
6 231 162 551
Flatsedge Cyperus esculentus L. 1 172 80 200
2 162 165 195
8 260 290 572
Ticktrefoil Desmodium paniculatum (L.) 2 44 63 195
Horsetail Equisetum arvense L. 3 246 160 2,200
Hydrocotyle L. Hydrocotyle amnericana L. 8 267 50 572
10 62 36 720
Turkey tangle Phyla nodiflora (L.) Greene 1 191 86 200
4 598 110 906
7 32 200 1,000
7 400 453 1,000
Plantain Plantago major L. 1 137 70 200
5 256 161 683
9 213 169 532
Blaker Rubus frutocosus L. 3 340 400 2,200
6 173 93 551
Goldenrod Solidago altissima L. 6 111 86 551
10 90 200 720
S owthi stl e Sonchus asper (L.) Hill 7 134 250 1,000
St. Augsie ras Stenotaphrum secundatum 1 164 100 200
3 516 320 2,200
Blue acket Tradescantia ohiensis Raf: 5 220 211 683
Tuberous vervain Verbena rigid Spreng. 1 176 130 200
8 155 198 572
Bipo sesbania Sesbania exaltata (Raf.) Cory 2 283 239 195
9 213 169 532
*Detection limit for Zn=1mg kg-
By comparing BCF and TF, we can assess the ability of different plants in taking
up metals from soils and translocating them to the shoots. Tolerant plants tend to restrict
soil-root and root-shoot transfers, and therefore has much lower accumulation in their
biomasses, while hyperaccumulators actively take up and translocate metals into their
aboveground biomass. Plants exhibiting TF and particularly BCF values less than 1 are
ineffective for use in phytoextraction (Fitz and Wenzel, 2002). A few plants growing at
the site were capable of accumulating heavy metals in their roots, but most of them had
Table 4-5 Accumulation and translocation of Pb, Cu and Zn in selected plants
factor Transl location factor
Site # Pb Cu Zn Pb Cu Zn
G~entiana ennelliana Fern 1 10.76 21.6 2.62 0.47 0.46 0.48
8 0.61 1.25 0.54 0.56 0.56 1.16
Cyperus esculentus L. 2 0.11 0.48 0.83 1.6 2.8 1.02
8 0.29 0.5 0.45 0.06 0.13 1.12
Phyla nolra L. Greene 1 0.5 1.55 0.96 0.55 0.45 0.45
7 0.05 0.01 0.03 0.71 11.75 6.25
7 0.19 0.24 0.4 0.12 0.13 1.13
Rubus frticosus L. 3 0.2 0.03 0.15 0.03 1.53 1.18
9 0.15 1.62 0.4 NA 5.64 0.79
Sesbania exaltata (Raf.)Cr 2 1.05 0.57 1.45 NA 4 0.84
Plan2tago major L. 1 0.1 1.15 0.69 5.96 0.43 0.51
Bidens alba (L.) DC 2 6.62 0.48 0.13 0.1 0.8 0.8
3 0.04 0.04 0.07 0.15 0.39 1.61
Pas lum notatm Flugge 4 0.42 0.26 0.5 0.47 1.41 0.7
9 NA 0.33 0.47 NA 1.1 0.8
Stenotahrum secundatum 1 0.34 1.1 0.82 0.45 0.77 0.61
Verbena rigid Spreng 8 0.02 0.08 0.27 0.31 0.72 1.28
Sonchus asper (L.) Hill 7 0.06 0.06 0.13 0.27 0.74 1.87
Solidago altissima L. 10 NA 0.77 0.13 NA 0.5 2.22
Desmodium panicula~tum (L.) DC 2 0.91 0.29 0.23 0.15 1 1.43
low TF's and BCF's, which means they have limited ability to accumulate and
translocate the metals studied (Tables 4-5).
Among the 17 plants screened, G. pennelliana, growing at site 1, had the highest
BCF of Pb (BCF=10.8), though its Pb concentrations in the plant body were <1,000 mg
kg- On the other hand, P. nodi~ora had the highest Pb concentration in the roots (1,183
mg kg- ), but its BCF was less than 1 (Table 4-5). Cyperus esculentus L.and
Plan2tagomajor L. had TF's greater than 1 (TF=1.6, and 6.0, respectively). The BCF
value of Pb found in this research was lower than that of Kim et al. (2003) in P.
thunbergii (5 58), and higher than from those (0.3-0.45) reported by Stoltz and Greger
(2002). Shallari et al. (1998) reported a high BCF of Pb in Koeleria eriostachya (31i),
Setaria viridis (18) and Verba~scum blattaria (3).
Similar to Pb, no plant species accumulated Cu above 1,000 mg kg-l (Table 4-3).
The maximum value for Cu was 460 mg kg-l in the roots ofP. nodi~ora. Though several
plant species showed BCFs or TFs greater than 1 for Cu, only Rubus fruticosus L.
growing at site 9 had both a BCF (1.6) and a TF (5.6) greater than 1 (Table 4-5).
However, soil Cu concentration at this site was relatively low, at 29 mg kg- The BCF
values for these species were lower than those found in P. thunbergii (41-168) by Stoltz
and Greger(2002). The highest BCF for Cu was 6.3 in Silene gallica by Shallari et al.
The highest Zn concentration was 598 mg kg- which was in the roots of P.
nodi~ora (Table 4-4). Both G. pennelliana. and S. exaltata had a BCF greater than 1, and
several species had a TF greater than 1, with P. nodi~ora having the highest TF of 6.3.
The highest BCF for Zn was 2.6 in G. pennelliana, lower than those reported by Kim et
al. (2002) in P. thunbergii (22 136) but higher than those (0.004-0. 11) reported by Stoltz
and Greger (2002). Research conducted by Shallari et al. (1998) reported the highest Zn
BCF for Alyssim markgurafii at 5.9.
Though none of the plants sampled were metal hyperaccumulators, some
interesting observations were noted. Based on the average BCFs of all plant samples,
plant roots were most efficient in taking up Cu (BCF=1.1), followed by Pb (0.79) and Zn
(0.50) (data not shown). Based on the average TFs of all the plant samples, the plants
were most efficient in translocating Cu (TF=1.2), followed by Zn (0.98) and Pb (0.54).
From a general point of view, among the three metals tested, the plants growing on the
Jacksonville site were most efficient in taking up and translocating Cu. Low translocation
of Pb indicates that plants are unwilling to transfer Pb from their roots to shoots due to
toxicity of Pb. Lead can be toxic to photosynthetic activity, chlorophyll synthesis and
antioxidant enzyme production (Kim et al., 2003). Baker and Brooks (1989) also
discussed the restriction of metal uptake by plants from contaminated soils and the
presence of exclusion mechanisms in such species. Since Zn and Cu are essential
nutrients for plant systems, their higher translocation factors can be supported. Thomas
and Eong (1984) treated established Rhizophora mucronata L.am. and Avicennia alba B.1
seedlings in sediment with Pb and Zn. For these species, root accumulation and reduced
translocation to the leaf tissues were observed for Zn and Pb.
In general, all three heavy metals occurred at elevated levels in the plant biomass as
a result of previous industrial activity at the site. Normal and phytotoxic concentrations of
Pb, Zn and Cu were reported by Levy et al. (1999), which were 0.5-10, and 30-300 mg
kg-l for Pb, 3-30 and 20-100 mg kg-l for Cu, and 10-150 and >100 mg kg-l for Zn.
Almost all collected species had heavy metal concentrations higher than the normal or at
phytotoxic levels. These results may indicate that plant species growing in heavy metal
contaminated site were tolerant of these metals. Restriction of upward movement into the
shoots can be considered a tolerance mechanism, as can the increase in metal-binding
capacity of their cell walls (Verkleij and schat, 1990).
In addition, the relationships of BCFs and TFs among the three metals were
determined through simple correlation. A few studies have been published to show that
high Pb levels reduce the uptake of other elements such as Fe, Mn and Zn. The
correlation of BCFs of all plant samples between two of the metals ranged from 0.63 to
0.84 (data not shown), i.e. a plant, which was effective in taking up Pb, was very likely to
be effective in taking up Cu and Zn. However, the relationship between TFs was
different. Only the TFs of Zn and Cu were correlated (r2=0.79), whereas no correlation of
TFs was found between Pb and Cu, or Pb and Zn. In other words, a plant, which was
effective in translocating Zn, was also effective in translocating Cu and vice versa.
However, Cu and Zn translocation in these plants were not related to Pb translocation.
These relationships may indicate that elevated Pb concentration can inhibit the transfer of
essential micronutrients to the plant biomass.
Table 4-6. Plants with high BCF and low TF for phytostabilization
Bioconcentration Translocation Applicable
factor factor Phystabilazation
Site # Pb Cu Zn Pb Cu Zn
Gentiana pennelliana Flugge 1 10.8 21.6 2.62 0.47 0.46 0.48 Pb, Cu Zn
8 0.61 1.25 0.54 0.56 0.56 1.16 Cu
Bidens alba (L.) DC 2 6.62 0.48 0.13 0.10 0.80 0.80 Pb
Sesbania exaltata (Raf.) Coryl 2 1.05 0.57 1.45 0.01 4.00 0.84 Pb,Zn
Plantago major L. 1 0.10 1.15 0.69 5.96 0.43 0.51 Cu
Stenotaphrum secundatum 1 0.34 1.10 0.82 0.45 0.77 0.61 Cu
Phytostabilization is another technology that can be used to minimize migration of
contaminants in soils (Susarla et al., 2002). This process uses the ability of plant roots to
change soil conditions via root exudation. Plants can immobilize heavy metals through
absorption and accumulation by the roots, adsorption onto roots, or precipitation within
the rhizosphere. This process reduces metal mobility and leaching into the ground water,
and also reduces metal bioavailability for entry into the food chain. One advantage of this
strategy over phytoextraction is the disposal of the metal-laden plant material is not
required (Susarla et al., 2002). By using metal-tolerant plant species for stabilizing
contaminants, particularly metals, in soil, could also provide improved conditions for
natural attenuation or stabilization of contaminant in the soil. Metals accumulated in the
roots are not considered to be a threat for release into environment. Studies are needed
regarding the turnover of nutritive roots and the potential release of metals from
decomposing roots (Weis and Weis, 2004). Also effects of plant-bacteria and/or plant-
mycorrhizae interactions, which might affect the metal uptake and translocation need
further investigation. Although no heavy metal hyperaccumulators were found in our
study, heavy metal tolerant plant with high BCF and low TF can be used for
phytostabilization of contaminated site (Table 4-6).
This study was conducted to screen plants growing on a contaminated site to
determine their potential for metal accumulation. A total of 17 plant species were
collected and analyzed for heavy metal concentrations. Only plants with both a BCF and
a TF greater than 1 have potential to be used for the purpose of phytoextraction. In this
study, no plant species were identified as metal hyperaccumulators based on Baker and
Brooks (1989). However, several plants had BCFs or TF >1. G. pennellian2a. was the
most effective in taking up all three metals, with BCFs at 1.1-22 with one exception
(Table 4-5). Cyperus esculentus was the most effective in translocating all three metals
(Table 4-5). Among those plant species collected from the contaminated site, G.
pennelha~na was considered as the most promising species for phytoextraction. Further
clarification of the phytoremediation potential of these plant species needs to be
investigated. Also the plant sampling was conducted randomly, and a more systematic
sample collection could yield more definitive results.
This study has shown that applying a mixture of phosphoric acid and phosphate
rock effectively reduced TCLP-extractable Pb (Toxicity Characteristic Leaching
Procedure), bioavailable Pb and Pb leachability in the Pb-contaminated soil studied.
Collected soil form the Hield had TCLP-extractable Pb < 127mg L- bioavailable Pb <26
mg L^1. Application of phosphoric acid and phosphate rock reduced TCLP-Pb by >95%
and bioavailable Pb by 25-42% and reduced Pb leachability from the contaminated soil.
A layer of phosphate rock resulted in improved reduction of Pb migration. Result
suggests layered phosphate rock below a contaminated soil may serve as a reactive
barrier to prevent Pb from migrating into the groundwater. No significant soil
acidification or phosphorous leaching was observed by the application of phosphoric acid.
Of all the treatments, mixture of phosphate rock and soil with co-application of
phosphoric acid (RhlAs1) was the most effective in decreasing leachable and bioavailable
Pb with acceptable impact on soil pH and eutrophication risk.
A greenhouse study showed that the three plants (Agrostis capillariesll1~~~111~~~11 Lolium
rigidum, and Bra~ssica napus) were effective in enhancing Pb immobilization in the
rhizosphere soil. Plant rhizosphere soil showed 13-20% reduction in bioavailable-Pb as
compared to the control soil. Reduction of the bioavailable Pb in the rhizosphere soil may
result from the formation of Pb phosphates, which was supported by the SEM
observation of the plant roots. Formation of stable lead-phosphate minerals may be
enhanced via biochemical actions of roots (i.e. organic acid, phosphatase exudation). This
study demonstrated that using the most effective P-application method coupled with plant
growth can maximize the effectiveness of phosphate-induced lead immobilization in
Elevated metal concentrations were observed in the 17 plants collected from the
experimental site located in Jacksonville, Florida, with total Pb concentrations ranging
from non-detectable to 1,183, Cu 5 to 460, and Zn 17 to 598 mg kg l.Among thel17
collected species, G. pennelliana. was the most effective in taking up all three metals,
with BCFsbeing 1.1-22 and Cyperus was most effective in translocating all three metals
with TFs being 1.0-2.8. Though no metal hyperaccumulator was identified, heavy metal
tolerant plants with a high BCF and a low TF can be used for phytostabilization to
minimize the migration of a particular contaminant in a soil.
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Joonki Yoon was born on May 27, 1976, in Seoul, Korea, to ByungKyou Yoon and
SookJae Lee. He attended Yonsei University where he received Bachelor of Science
degree in geology with minors in biology in February of 2002. Upon completion of his
B.S., he j oined the University of Florida in the fall of 2002 to pursue a Master of Science
degree in soil and water Science specializing in remediation of Pb-contaminated soils
with Dr. Lena Ma.