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ARSENIC HYPERACCUMULATION BY Pteris vittat L. AND ITS POTENTIAL
FOR PHYTOREMEDIATION OF ARSENIC-CONTAMINATED SOILS
GINTA MARIE KERTULIS-TARTAR
A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY
UNIVERSITY OF FLORIDA
Gina Marie Kertulis-Tartar
This work is dedicated to my wonderful husband, Kenneth Tartar, for his unending love,
patience and encouragement.
I wish to thank my advisor and mentor, Dr. Lena Q. Ma, for her invaluable advice,
guidance, critiques and devotion. I am grateful that she always expressed genuine
interest in my future, as well as in me as a student, a scientist and a person. I am also
grateful to my committee members, Drs. Nicholas Comerford, Charles Guy, Gregory
MacDonald and Joseph Vu, who provided valuable assistance and advice to ensure the
quality of my research. I also wish to sincerely thank Dr. Bala Rathinasabapathi, who
graciously spent countless hours advising me in plant physiology and biochemistry.
Much of the data collected and presented would not have been possible without the
assistance of Mr. Thomas Luongo. I am grateful not only for his analytical assistance but
also for his invaluable friendship and advice. I wish to thank Dr. Tait Chirenj e, who
provided experimental and statistical advice as well as friendship. I am grateful to Ms.
Heather Williams for her much needed assistance in harvesting ferns and soil sampling. I
also wish to thank the past and present members of the Biogeochemistry of Trace Metals
Laboratory, Maria Silva, Donald Hardison, Joonki Yoon, Abioye Fayiga, Drs. Jorge
Santos, Mrittunjai Srivastava, Nandita Singh, Rocky Cao, Chip Appel, Bhaskar Bondada,
Mike Tu, Carmen Rivero and Ju-Sik Cho, for all that they have taught me.
I am eternally grateful to my parents, Anthony and Barbara Kertulis, for their
unending love and support and for their constant encouragement of every one of my
endeavors. All that I am and all that I have accomplished is truly a result of their
dedication and commitment. I also wish to thank my mother-in-law, Margaret Tartar, for
her continuous words of encouragement and praise.
I would not have completed this work without the support, love and patience of my
husband, Kenneth Tartar. I am thankful for his unrelenting encouragement and
dedication, despite the countless sacrifices he made in order for me to complete my Ph.D.
I am truly thankful that God has blessed me by putting him in my life. I lovingly
dedicate this study to him.
TABLE OF CONTENTS
ACKNOWLEDGMENT S ................. ................. iv.............
LIST OF TABLES ................ ...............x............ ....
LI ST OF FIGURE S .............. .................... xii
AB STRAC T ................ .............. xiv
1 INTRODUCTION ................. ...............1.......... ......
2 LITERATURE REVIEW .............. ...............4.....
A rsenic ................... ......... ...............4.......
Chemistry of Arsenic ................. ...............4............ ....
Toxicity of Arsenic ................. ...............5............ ....
Arsenic in the Atmosphere ................ ...............6................
Arsenic in M inerals .............. ...............8.....
Arsenic in Water............... ...............8..
Arsenic in Soils............... ...............10.
Behavior of Arsenic ................. ...............10........... ....
Arsenic Availability............... ..............1
Arsenic Speciation............... ...............1
Arsenic Contamination ................. ...............14......... .....
Pesticides ................. ...............14......... .....
Mining and Smelting ..........._.._._ ...............16..............
Combustion of Fossil Fuels............... ...............17.
Biosolids ............._.. ... .. ....... ...... ..__. .............1
Remediation of Arsenic Contaminated Soils.. ....._.__.. ... ...__.. .........._.........18
Physical Remediation ........._.__........_. ...............18....
Chemical Remediation .............. ...............20....
Bioremediation .............. ...............20....
Phytoextraction ................. ...............22......._.. .....
Hyperaccumulators ................. ...............23.................
Pteris vittat L ................. .... ........ ...............24......
Other Arsenic Hyperaccumulating Ferns ....._____ ........___ ...............26
Arsenic in Plants ................... ...............27................
Arsenic Uptake by Plants .............. ...............28....
Antioxidants and Antioxidant Enzymes ......___ ..... ... ._._ ......_.... ......28
3 ARSENIC SPECIATION AND TRANSPORT IN Pteris vittat L. .........................34
Introducti on ........._..._. ...._ ... ...............34......
M materials and M ethods .............. ...............36....
Experimental Setup .............. ...............36....
Xylem Sap Extraction............... .. .................3
Chemical Analysis of Arsenic and Phosphorus. ................ ........................37
Arsenic Speciation in Plant and Xylem Sap Samples .............. ....................3
Experimental Design and Statistical Analysis............... ...............40
Re sults................. ...... ....... .... ....... ... . ........4
Arsenic Concentration and Speciation in Roots and Fronds ............... ... ............40
Arsenic Concentration and Speciation in Xylem Sap .............. ....................4
Phosphorus Concentration in Xylem Sap ................. ............... ......... ...44
Discussion ................. ...............47.................
4 EFFECTS OF ARSENIC ON GLUTATHIONE REDUCTASE AND
CATALASE IN THE FRONDS OF Pteris vittat L. ................... ...............5
Introducti on ........._._........ ..... ...............53.....
M materials and M ethods ............... ...............55....
Plant and Chemical Materials............... ...............5
Enzyme Extraction .............. .... ....... ...... ........5
Protein and Enzymatic Activity Determinations ................. .......................56
Enzyme Induction Study .............. ...............57....
Determination of Apparent Kinetics .............. ...... ....__ ......__............58
Determination of Arsenic Effects on Enzyme Activities .............. ................59
R e sults................ .......... .. ........_ ........... ............5
Glutathione Reductase and Catalase Induction Study ................. ........._.._.. ...59
Glutathione Reductase and Catalase Apparent Kinetics .............. ...................61
Effect of Arsenic on Enzyme Activities ................. ..........__ ....._.. .....68
Discussion ........._..... ...._... ...............72.....
5 PHYTOREMEDIATION OF AN ARSENIC-CONTAMINATED SITE USING
Pteris vittat L. ............. ...............76.....
Introducti on ........._._........ ..... ...............76.....
M materials and M ethods .............. ...............78....
Experimental Site .............. ...............78....
Planting and Plot Maintenance ...._.. ...._._._._ .........__. ............7
Plot 1 .............. ...............79....
Plot 2 .............. ...............79....
Plant Harvests ........._... ...... ..... ...............80....
Plot 1 .............. ...............80....
Plot 2 .............. .. .......... .. .... ............8
Determination of Frond Biomass and Arsenic Concentrations ................... ........83
Soil Sam pling .............. ...............83....
Plot 1 .............. ...............83....
Plot 2 ................... .. ..... .... ..........8
Determination of Total Soil Arsenic .............. ...............84....
Sequential Soil Arsenic Fractionation ................ ...............84......._.._....
Bioconcentration Factor .............. ...............86....
Statistical Analysis .............. ...............87....
Re sults ................ .......... .. .................87.....
Arsenic Removal by Ferns .............. ...............87....
Plot 1 .............. ...............87....
Plot 2 .............. ....... ...............90.
Soil Arsenic Concentrations ................ ...............91........... ....
Plot 1 .............. ...............91....
Plot 2 .............. ........ .. ............9
Sequential Soil Arsenic Fractionation ................ .....................__........95
Mass balance of Arsenic ........._..._.._ ...._._. ...............97....
Plot 1 .............. ...............97....
Plot 2 .............. ...............97...
Bioconcentration Factor .............. ...............98....
Discussion ................. .......... ...............98.......
Plant Arsenic Removal .............. ...............99....
Soil Arsenic Concentrations ................ ...............103................
Sequential Arsenic Fractionation .............. ...............105....
M ass Balance ................. ........._.. ...............107......
Estimated Time of Remediation ......__....._.__._ ......._._. ...........13
Estimated Remediation Cost ............. ......___ .....___............1
Suggested Phytoextraction Setup ................. ...............117................
6 EFFECT OF Pteris vittat L. ON ARSENIC LEACHING AND ITS
POTENTIAL FOR THE DEVELOPMENT OF A NOVEL
PHYTOREMEDIATION METHOD ................. ...............119................
Introducti on ................. ...............119................
M materials and M ethods ................. .... .......... ..........2
Overview of Proposed Phytoleaching System ............. ...... __ ..............121
Soil ................. ...............122................
Treatm ents ................... ... ......... ...............122......
Fern, Soil and Leachate Analyses .............. ...............124...
Experimental Design and Statistical Analysis............... ...............12
Re sults ................ ...............125................
Leachate ................. ...............125................
Ferns ................. ................. 127........ ....
Soil ................. ...............129................
Discussion ................. ................. 13......... 0....
Future Directions ................. ................. 13......... 5....
7 CONCLUSIONS .............. ...............138....
LIST OF REFERENCES ................. ...............142................
BIOGRAPHICAL SKETCH ................. ...............156......... ......
LIST OF TABLES
2-1 Summary of select current remediation technologies for arsenic-contaminated soil..19
3-1 Total arsenic concentrations in xylem sap of P. vittata exposed to 0, 10 or 50 mg 1-
1 ars eni c................. ...............43.._._._ .....
4-1 Summary of apparent kinetic parameters for GR and CAT .............. ...................68
5-1 A comparison of the total biomass removed, average frond arsenic concentration
and amount of arsenic remediated from the senescing frond harvests in 2001 and
(DD I) in 2002 ........._.... ....._.. ...............88....
5-2 Comparison of average frond arsenic concentrations, total amount of biomass
removed and amount of arsenic removed between the senescing fern fronds
harvested in 2001 and 2002 (DDI), and all fronds harvested in December 2001
and August 2002 (A2x) ..............._ ...............89......... .....
5-3 Comparison of the total amount of biomass removed between the fronds harvested
in 2003 and 2004. ............. ...............91.....
5-4 Comparison of the average frond arsenic concentrations and amount of arsenic
removed between the fronds harvested in 2003 and 2004 .............. ....................91
5-5 Average soil arsenic concentrations and arsenic depletion of soil samples taken in
plot 1 in 2000, 2001 and 2002 ..........._ .....___ ...............92
5-6 Average soil arsenic concentrations and net arsenic depletion of soil samples taken
inside plot 2 in 2002, 2003 and 2004 .............. ...............93....
5-7 Average soil arsenic concentrations and net arsenic depletion of soil samples taken
outside plot 2 in 2002, 2003 and 2004 .............. ...............93....
5-8 Calculated mass balance of arsenic in the soil-plant system of plot 1 from 2000 to
2002 ........... __..... ._ ...............97....
5-9 Calculated mass balance of arsenic in the soil-plant system of plot 2 from 2002 to
2004 ........... __..... ._ ...............98....
5-10 Estimated time for phytoextraction of plot 2 with P. vittata.........._.._... ................115
6-1 The effects of chemical treatment and leaching frequency on frond biomass, frond
arsenic concentration and the amount of arsenic removed from the arsenic-
contaminated soil............... ...............128.
LIST OF FIGURES
2-1 Chemical structures of arsenate, arsenite, monomethylarsonic acid (MMA) and
dimethylarsinic acid (DMA) .............. ...............5.....
2-2 Global arsenic cycle............... ...............7..
2-3 Arsenic concentrations in groundwater sampled in the United States ..........................9
2-4 Pteris vittat L. growing at an arsenic-contaminated site .............. .....................2
3-1 Total arsenic concentrations in the fronds and roots of P. vittat exposed to 0, 10
or 50 mg 1-1 arsenic as As(III), As(V), MMA or DMA ................. ............... .....41
3-2 Percentages of As(III) and As(V) in the fronds and roots of P. vittat exposed to
As(III) or As(V) .............. ...............42....
3-3 Concentrations of As(III), As(V), DMA and MMA in the xylem sap of P. vittat....45
3-4 Comparison of total arsenic and Pi (inorganic phosphorus) concentrations in the
xylem sap of P. vittata................. .............4
4-1 Glutathione reductase activity in P. vittata plants exposed to 0 and 10 mg 1-1
4-2 Immunoblot of GR activity in (A) crude extract of arsenic treated P. vittata (B)
crude extract of control P. vittat and (C) crude extract of Zea mays ...................60
4-3 Catalase activity in P. vittat plants exposed to 0 and 10 mg 1-1 arsenic. ........._.........61
4-4.Apparent kinetic analysis s of sub state, GS SG, for GR activity in P. vittata..............62
4-5 Apparent kinetic analysis of substrate, GSSG, for GR activity in P. ensiformis ........63
4-6 Apparent kinetic analysis of substrate, NADPH, for GR activity in P. vittat...........64
4-7 Apparent kinetic analysis of substrate, NADPH, for GR activity in P. ensiformis.....65
4-8 Apparent kinetic analysis of H202 for CAT activity in P. vittat .........._.... ..............66
4-9 Apparent kinetic activity of H202 for CAT activity in P. ensiformis ................... .......67
4-10 Effect of arsenite on GR activity in P. vittata and P. ensiformis .............. ................69
4-11 Effect of sodium ar senate on CAT activity in P. vittat .............. .....................7
4-12 Effect of sodium arsenate on CAT activity in P. ensiformis ..........._................... 70
4-13 Effect of sodium arsenate on CAT activity in bovine liver ...........__.... ................ 71
4-14 Comparison of the percent change in CAT activity in P. vittata, P. ensiformis and
bovine liver (CAT positive control) upon exposure to arsenate. ........... ................71
5- 1 Photographs of P. vittat growing in the first experimental plot (200 1 to 2002)........8 1
5-2 Photographs of P. vittat growing in the second experimental plot (2003 to 2004)...82
5-3 Soil sampling plan for experimental plot 2 .............. ...............86....
5-4 Area graphs of plot 1 showing the total soil arsenic concentrations in the top 15
cm of soil ............ ...............94.....
5-5 Sequential arsenic fractionation concentrations for soil sampled within plot 1..........96
6-1 Schematic diagram of the phytoleaching system .............. ..... ............... 12
6-2 Total amount of arsenic removed from the soil through leaching for each chemical
and frequency treatment ...........__..... .___ ...............126...
6-3 Leachate arsenic concentrations for every frequency, chemical and fern treatment
of each leaching event ...........__..... .___ ...............127...
6-4 Total arsenic removed from the arsenic-contaminated soil via the phytoleaching
(leaching and fern) system .............. ...............129....
6-5 Total soil arsenic concentrations before and after the leaching treatments .............130
Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Doctor of Philosophy
ARSENIC HYPERACCUMULATION BY Pteris vittat L. AND
ITS POTENTIAL FOR PHYTOREMEDIATION OF ARSENIC-
Gina Marie Kertulis-Tartar
Chair: Lena Q. Ma
Major Department: Soil and Water Science
Pteris vittat L, an arseni c-hyperaccumulating fern, was examined to understand its
hyperaccumulating ability and for its use in remediating arsenic-contaminated soils.
Transport of arsenic in xylem sap of P. vittat was investigated. Ferns were subj ected to
arsenate, arsenite, dimethylarsinic acid (DMA) or monomethylarsonic acid (MMA).
Xylem sap was collected and analyzed for arsenic concentration, speciation and
phosphorus concentration. When inorganic arsenic was supplied, arsenate appeared to be
the preferred species transported in the xylem sap. When arsenic was supplied in
methylated form, it was transported mainly in that form. Results from glutathione
reductase (GR) and catalase (CAT) enzymatic studies in P. vittat revealed that, upon
arsenic exposure, CAT activity was induced but GR activity was not. Further, GR was
not inhibited or activated by arsenic. However, CAT activity appeared to be activated by
arsenate. This activation may allow P. vittat to more efficiently mediate stress caused
by arsenic. A field study was conducted to determine the efficiency ofP. vittata in
phytoextraction of arsenic contaminated soil. The study suggested P. vittata is capable of
accumulating arsenic from contaminated sites, and a single harvest per year yields the
greatest arsenic removal. Further, results from sequential arsenic fractionation analyses
suggested that P. vittata is able to access arsenic from more unavailable soil fractions.
Phytoextraction of arsenic-contaminated soils using P. vittat may be competitive with
conventional remediation systems, but its application may be more practical for low-level
contamination. The phytoextraction study revealed a discrepancy in mass balance. One
hypothesis was that combination of over watering and solubilization of arsenic by root
exudates caused leaching. Therefore, it was important to identify if leaching was
occurring. It was also hypothesized that leaching may be harnessed for development of
an innovative ex-situ soil remediation method, phytoleaching. Water and chemical
solutions were added to promote arsenic leaching, while ferns removed arsenic via
uptake. More arsenic was leached from soil when ammonium phosphate solution was
applied. When ferns were present in contaminated soil, less arsenic was leached,
indicating that P. vittata does not promote arsenic leaching. Phytoleaching may be a
feasible remediation option with additional studies and refinement.
Arsenic (As) contamination of soil is a growing concern worldwide because it is
toxic and is a suspected a carcinogen. When arsenic is in soil and water, it can be taken
up by plants and indirectly ingested by animals and humans. Arsenic occurs naturally in
the environment, but significant arsenic levels result from anthropogenic sources, such as
mining and smelting operations, fuel combustion, biosolids, tanning, wood preservatives
and pesticides (O'Neill 1990).
Recent attention has been focused on chromated copper arsenic (CCA) treated
lumber, which has been widely used as a preservative. The treated lumber can serve
many purposes: telephone poles, decks, pilings, home construction and playground
equipment. However, there is concern regarding the leaching of arsenic from CCA
treated lumber, prompting numerous studies that have addressed this issue (Cooper,
1991; Stilwell and Gorny 1997; Lebow et al., 2003). Arsenic contamination occurs
through other sources, such as those previously mentioned. It is important to address
contamination of soil by arsenic and target it for appropriate remediation to prevent
possible impacts on the ecosystems.
Hyperaccumulators are plants that can take up and concentrate greater than 0.1% of
a given element in their tissue. Recently, an arsenic hyperaccumulator, Pteris vittat L.
(Chinese brake fern), was discovered (Ma et al., 2001). This arsenic hyperaccumulator
may offer an alternative to more traditional remediation technologies for arsenic
Phytoremediation is the use of plants to remove or render contaminants harmless in
the ecosystem. Phytoremediation actually includes several methods, such as
phytovolatilization, phytostabilization and phytoextraction. Phytovolatilization refers to
the uptake, translocation and volatilization of contaminants from plants. The
contaminants may or may not be transformed during this process. Phytostabilization
employs plants in order to contain contaminants in the soil, preventing migration of the
contaminant off site. Phytoextraction is the use of plants, preferably hyperaccumulators,
to take up contaminants. Subsequently, the plants are harvested, transported and
disposed off site (Schnoor, 2002).
Phytoextraction has become increasingly popular because of its low cost compared
to more traditional remediation technologies. The costs involved in phytoremediation
may include planting, maintenance, harvesting and disposal of plant biomass. The
volume and mass of the plant disposal are significantly less than the disposal of soil when
excavation is required. However, because phytoextraction is dependent on the plant,
conditions at the site must be able to maintain plant production, and the contaminant must
be accessible to the roots for uptake. In addition, soils with very high contaminant
concentrations may inhibit plant growth and/or significantly prolong the amount of time
required for remediation (Schnoor, 2002). Much research is still required to ensure
proper employment and utilization of phytoextraction.
In general, arsenic is toxic to plants, especially in high concentrations. Arsenate
can disrupt oxidative phosphorylation, and the production of ATP (Meharg and MacNair
1994; Oremland and Stolz 2003), while arsenite affects the function of enzymes and
proteins by binding to sulfhydryl groups (Leonard and Lauwerys 1980; Oremland and
Stolz 2003). Also, the conversion of arsenate to arsenite, the more toxic form of arsenic,
in the plant may create reactive oxygen species (ROS) that can damage plant cells. The
toxicity of arsenite may be ameliorated through the production and use of glutathione
(GSH) and/or phytochelatins (PC). Antioxidants and antioxidant enzymes may also
assist in stress management by responding to the increased levels of ROS in the plant.
It is important to understand the ability of P. vittata to hyperaccumulate arsenic and
its usefulness in the phytoremediation of arsenic-contaminated soils. The experiments
included in this proj ect were conducted to better understand P. vittat. The two main
objectives were: 1). to increase understanding of the ability of P. vittat to
hyperaccumulate arsenic; and 2). to determine the practicality, efficiency and ability of P.
vittat to phytoremediate arsenic-contaminated soil.
The first obj ective was addressed by examining the speciation and transport of
arsenic in P. vittat and the effects of arsenic on the antioxidant enzymes glutathione
reductase (GR) and catalase (CAT) in P. vittata. The second objective was achieved
through phytoextraction Hield studies at a CCA-contaminated site, which specifically
investigated into the effects ofP. vittat on the leaching of arsenic and its as a new
phytoremediation technique, and examined the availability of arsenic in CCA-
Arsenic has a long history being employed for medicinal uses, pigments and
poisons. Around 1775, Carl Scheele developed the compound Paris Green, which was
used as a pigment in wallpaper, paints and fabrics. However, persons living in the homes
containing Paris Green often became ill from direct contact with arsenic or from arsenic
volatilization from the pigment. There are also numerous documented accounts of
individuals using arsenic to intentionally poison others (Buck, 1978). There is even a
theory suggesting that the distressing fate of the 90% of the Jamestown colonists who
perished during the winter of 1609-1610 may have been a result of arsenic poisoning and
not of starvation (Marengo, 2001; Gundersen, 2002). It is apparent that much of the
history of arsenic is blemished with its poisonous properties.
Chemistry of Arsenic
Arsenic, element number 33 in the periodic table; atomic weight 74.9216, is a
crystalline metalloid or transition element (Group Sa). Its outer electronic configuration
is 4s24p3. Arsenic can exist in an allotropic form of alpha (yellow), beta (black) or
gamma (gray). It can be present in several oxidations states such as -3, 0, +3 and +5.
However, the most common forms of arsenic found in the environment include are
arsenate [As(V)] and arsenite [As(III)] (Fig. 2-1) (Adriano, 1986; Matera and Le Hecho,
O O O
HO As-- OH HO As--- OH HO As- OH CH,--- As-- OH
OH OH CH, CHa
Arsenate Arsenite Monornethylarsonic acid Dimethylarsinic acid
Figure 2-1. Chemical structures of arsenate, arsenite, monomethylarsonic acid (MMA)
and dimethylarsinic acid (DMA).
Toxicity of Arsenic
A human may ingest as much as 900 Clg of arsenic per day, depending on the
environment (Fowler, 1977). Generally speaking, inorganic forms of arsenic, arsenite
and arsenate, are more toxic than the organic forms of arsenic. This is unlike most other
metals (O'Neill, 1990). Overall, the arsenic toxicity pattern is as follows: AsH3 > As3+>
As5+ > organic arsenic.
Arsine gas (AsH3) is considered to be an extremely toxic form of arsenic. As little
as 4 Clg I~ inhaled into a human body can interfere with many metabolic processes.
Arsine gas inhalation can result in decreased erythrocyte osmotic resistance, reduced
hemoglobin and erythrocytes and increased reticulocytes. Ultimately, the number of red
blood cells may decrease by 50% in as little as one hour after exposure to arsine gas
Of the inorganic forms of arsenic, the trivalent form, or arsenite, is considered more
toxic than the pentavalent form, or arsenate. This is because arsenite can readily combine
with thiol (SH) groups. In addition, many enzymes and enzymatic processes may be
inhibited by arsenite (Fowler, 1977; Oremland and Stolz 2003). In general, arsenite
compounds are considered to be carcinogenic to humans (Hathaway et al., 1991;
Gochfeld, 1995). However, it is interesting to note that arsenite is a component of the
drug, Trisenox", which is used in the United States to treat people with afflicted acute
myeloid leukemia (Hall, 2002).
Because of the chemical similarities between arsenate and phosphate, arsenate has
the ability to replace phosphate in many biochemical processes. For example, arsenate
can disrupt mitochondrial oxidative phosphorylation and thus the production of the
nucleotide, adenosine triphosphate (ATP), which is a main energy source for cells. This
process is known as arsenolysis, or the hydrolytic process whose first step is the
replacement of arsenate for phosphate (Meharg and MacNair, 1994; Hall, 2002;
Oremland and Stolz, 2003). Arsenate also has the ability to replace phosphate in DNA,
ultimately compromising DNA processes (Fowler, 1977).
Ironically, arsenate is often converted to the more toxic form, arsenite, via
enzymatic or non-enzymatic processes in the environment. The arsenate reductase
enzyme has been identified in animals, bacteria and yeast. However, this enzyme has not
yet been identified in plants (Mukhopadhyay et al., 2002; Rosen, 2002). Arsenite may be
detoxified through methylation to monomethylarsenate (MMA) or dimethylarsenate
(DMA) (Fig. 2-1). The methylated forms of arsenic, which are generally excreted from
the body, are considered to be less toxic compared to inorganic forms of arsenic (Johnson
and Farmer, 1991).
Arsenic in the Atmosphere
Arsenic is cycled between the lithosphere, pedosphere, biosphere, hydrosphere and
atmosphere (Fig. 2-2). The atmosphere contains 0.8 x 106 kg (Walsh et al., 1979) to 1.74
x 106 kg of arsenic (Chilvers and Peterson, 1987). Approximately 85% of this arsenic is
located in the northern hemisphere, due to a higher number of industrialized countries
and a larger land mass (Matschullat, 2000).
Arsenic may be emitted into the atmosphere through natural sources (i.e.,
volcanoes) or anthropogenic sources. Approximately 60% of the anthropogenic arsenic
emissions results from coal combustion and copper smelting. Wood preservation,
herbicides, steel production, lead and zinc smelting, and incineration account for the
remaining 40%. Most of metallic arsenic emitted into the atmosphere is present as
particulate matter, and it may be retained in the atmosphere for seven to 10 days
Plants Anthropogenic I Rivers ) Oceans
I ooduction I subduction
Figure 2-2. Global arsenic cycle (adapted from Matschullat, 2000).
Arsenic in Minerals
Arsenic, which is the 52nd most abundant element in the earth's crust, has an
average crustal concentration of 1.5 to 2.0 mg kg-l (Adriano, 1986). Approximately 4.01
x 106 kg of arsenic is present in the earth's crust (Matschullat, 2000). On average, shales,
granites and sandstones have 13.0, 3.0 and 1.0 mg As kg- respectively (Onishi, 1969).
In general, arsenic concentrations in igneous rocks range from <1 to 15 mg As kg- ;
argillaceous sedimentary rocks (such as shale, sandstone and slate) from <1 to 900 mg As
kg- ; limestones from <1 to 20 mg As kg- ; and phosphate rocks from <1 to 200 mg As
kg-l (O'Neill, 1990). However, rocks associated with uranium may contain much higher
concentrations of arsenic [Committee on Medical and Biological Effects on
Environmental Pollution (CMBEEP), 1977].
More than 200 arsenic-containing minerals exist. Of these minerals, most are
arsenates (60%), with the rest being sulphides and sulphosalts (20%) and arsenides and
arsenites oxides (20%). Arsenopyrite (FeAsS2) is the most common arsenic mineral
(O'Neill, 1990). Arsenic-containing sulfides, such as arsenopyrite, tend to be important
arsenic-containing minerals. Examples of these are realgar (AsS), niccolite (NiAsS) and
cobaltite (CoAsS) (Allard, 1995; Reimann and deCaritat, 1998).
Arsenic in Water
Arsenic is relatively soluble in salt and fresh waters, and it can be present as
arsenite, arsenate or methylated arsenic. Mobilization of arsenic from soils or inputs
from anthropogenic sources can cause increases in stream concentrations and eventually
ocean concentrations (Matschullat, 2000). Groundwater contamination can be a result of
the dissolution of minerals from rocks and soil or anthropogenic sources.
Contamination of drinking water by arsenic is a serious threat to millions of people
worldwide. Of most prominence are the severe health problems of thousands of people
in Bangladesh and west Bengal, India. Health concerns arise due to arsenic-contaminated
groundwater (Chatterjee et al., 1995; Das et al., 1995; Abernathy et al., 1997). The
regulated upper limit of arsenic in drinking water in the United States is 10 Clg I 1. All
public drinking water systems must meet this standard by 2006 [United States
Environmental Protection Agency (USEPA), 2001]. Figure 2-3 shows estimated
concentrations of arsenic in groundwater in the United States. Concentrations tend to be
highest in parts of the West, Midwest and upper Northeast.
Arsenic on entraton s in atleast
25%of-- s aampe exed
50ug/L lt O lnufa n
Fiur -3rsnc- cocnrtosingonwtrsmpe nteUiedSae Rkr
Arsenic in Soils
As previously mentioned, arsenic is found in many different minerals and rocks.
As a result, arsenic is also found naturally in soils due to the weathering of these rocks
and minerals (Adriano, 1986). Arsenic concentration generally ranges from 0.2 to 41 mg
kg-l in soils worldwide (Kabata-Pendias and Pendias, 2001). Arsenic concentration
averaged 7.2 mg kg-l for surface soils in the United States (Shacklette and Boerngen,
1984). However, agricultural surface soil exposed to repeated arsenic pesticides can have
an arsenic concentration as high as 600 mg kg-l (Adriano, 1986), and soil arsenic may
range from 400 to 900 mg kg-l in areas of arsenic mineral deposits (National Research
Council Canada [NRCC], 1978). Also, soils in areas near coal mining and those
overlying sulfide ore deposits may have even higher arsenic concentrations.
Behavior of Arsenic
Arsenic and phosphorus (P) have similar chemical properties; therefore, they act
similarly in the soil. Phosphorus and arsenic may compete with each other for soil
fixation sites and for plant uptake (Adriano, 1986). The phytotoxicity of arsenic may
increase with decreasing soil phosphorus levels (Rumburg et al., 1960; Juska and
Hanson, 1967). Still, other experiments have indicated that additional phosphorus may
increase arsenic phytotoxicity by releasing more arsenic into solution (Schweizer, 1967;
Jacobs and Keeney, 1970).
The total arsenic concentration in soils does not necessarily determine the arsenic
phytoavaliability (Adriano, 1986). Although a finite amount of the total arsenic in the
soil is readily mobile, the rest is not available to plants because it is associated mostly
with iron (Fe) and aluminum (Al). The arsenite form (reduced form) is generally more
soluble in soil than the arsenate form (oxidized form). The concentration of soluble
arsenic is directly proportional to plant arsenic toxicity, although soil properties are also
important in the determination of arsenic availability (Kabata-Pendias and Pendias,
The availability of arsenic in soils may be affected by many soil factors, such as
soil pH (Adriano, 2001). In general, soil pH is important because it affects arsenic
speciation and leachability. The adsorption optimum for arsenite is approximately at pH
7.0; however, arsenate adsorbs optimally at pH 4.0 (Pierce and Moore, 1982). Overall, at
a low soil pH the hydroxyl groups on the outside of clays, amorphous silicates and metal
oxides become protonated. These sites are then able to adsorb arsenic anions present in
the soil. Therefore, arsenic is less mobile at lower pH because most of the arsenic is
present as arsenate in (aerobic) soils, and there are high concentrations of arsenic-binding
species, such as iron and aluminum at low pH (Sposito, 1989). As the pH increases there
are fewer protonated sites, allowing the arsenic to become more mobile.
However, arsenic does have the ability to form a strong association with calcium
(i.e., calcite) allowing it to possibly be retained at a higher pH. This association may be
found under high arsenic concentrations, where arsenic has a secondary preference to
calcium over aluminum (Woolson, 1983). At lower pHs, calcite is dissolved by the
acidic conditions, and the arsenic is released.
Soil texture is another important factor affecting arsenic availability (Adriano,
2001). For example, soil texture affects the soil surface area. Finer textured soils (silts
and clays) have much more surface area than coarse (sandy) soils; therefore, they are
more reactive. Finer-textured soils are more likely to retain higher amounts of trace
elements compared to sandy soils (Chen et al., 1999; Berti and Jacobs, 1996). Apart from
increased surface area, fine textured soils also have higher cation exchange capacity
(CEC). A greater CEC leads to higher retention for cationic species like copper (Chen et
It is also possible to find more organic matter (OM) in finer textured soils with a
high CEC, compared to sandy soils with low CEC. Often, high OM leads to high CEC,
mostly from the pH-dependent charge. Conditions in fine textured soils are also more
conducive to OM accumulation and retention. Organic matter increases retention of both
cationic and anionic species. This is achieved through cationic bridging by iron and
aluminum, resulting in anion retention, and the dissociation of edges of organic
complexes in response to changes in pH. This allows for the retention of both cations
and anions, depending on pH.
Soils with sandy textures may increase the toxicity of arsenic to plants and arsenic
mobility; compared to soils with clayey textures (Jacobs and Keeney, 1970; Woolson,
1973; Akins and Lewis, 1976; Adriano, 1986). The presence of iron and aluminum
oxides also plays an important role in the ability of a soil to retain arsenic (Adriano,
2001; Jacobs et al., 1970; Lumsdon et al., 1984). Further, soil iron and phosphorus
concentrations are important factors influencing arsenic concentrations in Florida soils
(Chen et al., 2002).
Research on phosphorus indicated that sand grains with clay coatings have a higher
ability to retain elements compared to bare quartz grains (Harris et al., 1987a, b). The
common coating components, for example, metal oxides and aluminosilicates, have a
high affinity for trace elements, such as arsenic. Some soil horizons (i.e., albic horizons
in Spodosols) have been exposed to extreme weathering and leaching. This weathering
results in the sand grains being stripped of their clay coatings (Harris et al., 1987a, b).
However, Rhue et al. (1994) found that some of these horizons are able to retain their
clay coatings. As such they exhibit greater retention ability compared to those that did
not retain their coatings.
As previously mentioned, arsenic can be present in four oxidation states: (-3), (0),
(+3) and (+5). However, arsenate (+5) and arsenite (+3) are the more prevalent forms in
the soil environment. The form of arsenic in soil is very important, as it can dictate the
behavior. Arsenite is considered to be more water soluble, or mobile, in soils compared
to arsenate (Pierce and Moore, 1982). This is important because arsenite is considered to
be more toxic form of arsenic. The occurrence of arsenite or arsenate in soil is a function
of both the pH and redox potential (Eh) (Masscheleyn et al., 1991)
In aerobic soils, arsenate constitutes up to 90% of the total arsenic. However,
under anaerobic conditions, only 15 to 40% of the arsenic is present as arsenate (O'Neill,
1990). Arsenate can form insoluble compounds with aluminum, iron and calcium in the
soil. As previously mentioned, arsenate and phosphate are chemical analogues.
Therefore, arsenate behaves similarly to phosphate, and it often competes with phosphate
in the soil.
When soil conditions are moderately reducing (Eh 0 to -100 mV), the solubility of
arsenic is dictated by iron oxyhydroxides, as the arsenate precipitates with iron
compounds. In soils that are flooded, conditions are extremely reducing (Eh -200 mV),
and arsenic is much more mobile (Matera and Le Hecho, 2001). However, the
transformation of arsenate to arsenite is very slow. Therefore, arsenate can often be
detected in highly reduced soils (Onken and Hossner, 1996).
Microorganisms also play a role in the speciation of arsenic in soils. There are
arsenite-oxidizing bacteria that can transform arsenite to arsenate. Similarly, arsenate-
reducing bacteria can convert arsenate to arsenite (Cullen and Reimer, 1989). Also
present in soils are microorganisms that can convert arsenite to methylated forms of
arsenic (Pongratz, 1998).
Arsenic contamination of soil and water can result from several anthropogenic
activities, such as: pesticide use/production, mining, smelting, combustion and
sewage/solid waste (O'Neill 1990; Davis et al., 2001; Oremland and Stoltz, 2003).
Arsenical compounds have been used in pesticides for over one hundred years.
However, since the 1970's their total use has declined (O'Neill, 1990). Arsenic is
effective as an herbicide, in wood treatment and as a desiccant of cotton. Worldwide
average uses have been estimated at 8,000 t yr- for herbicides, 12,000 t As yr- for cotton
desiccants and 16,000 t yr- for wood preservatives (Chilvers and Peterson, 1987).
Of recent concern has been the use of chromated copper arsenate (CCA).
Chromated copper arsenate is a pesticide that helps to reduce microbial and fungal decay
of wood products. Arsenic and copper (Cu) act as the insecticide and fungicide,
respectively. Chromium (Cr) fixes the arsenic and copper to the wood's cellulose and
other components (Dawson et al., 1991). In January 2004, domestic use of CCA-treated
wood was voluntarily discontinued (USEPA, 2002b). Prior to that, it constituted
approximately 75% of the treated wood market by volume (Solo-Gabriele et al., 1999),
which is a strong statement to its effectiveness (Warner and Solomon, 1990).
In the Southeastern United States CCA-treated wood use was particularly high.
This high use is a result of the hot, humid summers and mild winters of the region. Such
conditions increase both the rate of weathering and biological (i.e., microbial and fungal)
activity and subsequent decay (Chirenj e et al., 2003). However, the massive manufacture,
use and disposal of CCA-treated wood has led to increased loadings of these elements
into the environment (Carey et al., 1996; Lebow, 1996; Cooper and Ung, 1997; Stilwell
and Gorny, 1997; Solo-Gabriele et al., 2000; Townsend et al., 2000; Rahman et al.,
2004). For example, when treating wood with CCA, up to 250 1 of CCA solution are
applied under pressure for every 1 m3 Of wood. This results in treatment solutions
containing arsenic, chromium and copper concentrations in the range of 1000-5000 mg
kgl (Aceto and Fedele, 1994). A single 12 ft x 2 inch x 6 inch piece of lumber treated by
type C CCA contains approximately 27 grams of arsenic. This is enough arsenic to
poison more than 200 adults. On average, one tablespoon, or approximately 20 grams, of
CCA wood ash can contain enough arsenic to kill an adult human.
In CCA solution, arsenic is generally used in its anionic form, arsenate. Solo-
Gabriele et al. (2000) have shown that although new CCA wood contains predominantly
arsenate, which is moderately toxic and carcinogenic, the concentrations of arsenite,
which is highly toxic and carcinogenic, increased as the wood aged. These results show
that the form of arsenic changes in both soil and in CCA treated wood. Also, the increase
in arsenite concentrations is of concern for both human and ecosystem health.
The concentrations of the three elements in CCA-treated wood, copper, chromium
and arsenic, are not very different from each other. However, compared to copper and
chromium, arsenic can leach out as much as an order of magnitude more from the treated
wood products. The type of CCA, wood type, orientation of the wood and surface area
may affect the degree of arsenic leaching from the treated wood (Hingston et al., 2001).
Climate and moisture conditions also affect arsenic leaching from treated lumber (Kaldas
and Cooper, 1996; Lebow et al., 2004). Arsenic concentrations of approximately 550 mg
kg-l have been reported in the vicinity of CCA-treated utility poles (Cooper and Ung,
Mining and Smelting
Arsenic is often a by-product of smelting lead, zinc, copper, iron, gold and
manganese (Benson et al., 1981). A CMBEEP report (1977) indicated that copper, zinc
and lead smelting and refining releases 955, 591 and 364 metric tons of arsenic for every
million metric tons produced, respectively. During mining and smelting processes,
arsenic may be released as a gas or as fly ash. Soils in the vicinity of smelters can
become contaminated through deposition by rain or the settling of fly ash. An
examination of a smelter in Tacoma, Washington found 7 to 152 t As yr- were deposited,
while a smelter in Canada deposited 19 to 2600 t As yr- (Woolson, 1983).
Mine spoils/dumps can also cause arsenic contamination. Arsenic may leach from
these spoils and/or finer material may be dispersed by wind. Arsenic concentrations were
found to be over 4 g kg-l in the vicinity of old mine spoils in Virginia. The condition
may be exasperated by the difficulty or inability of plants to grow and thrive on these
soils/spoils. A lack of productive vegetation may decrease the stability of the soil, thus
increasing water and wind erosion (O'Neill, 1990).
Combustion of Fossil Fuels
Fossil fuels naturally contain arsenic. On average fuel oils contain 0.015 mg As
kg- (O'Neill, 1990). However, arsenic concentration in coal can range from 15 to 150
mg kg-l (Cullen and Reimer, 1989). The increase in the burning of fossil fuels has also
increased the opportunity for arsenic contamination in soil.
During the combustion of fossil fuels, such as coal and oil, arsenic may be
volatilized. The amount of arsenic volatilized is dependent on the form of the arsenic in
the coal. For example, arsenical sulfides are more volatile than organically-complexed
arsenic. Approximately 600 million tons of coal were burned in 1983 in the United
States; this resulted in the emission of an estimated 800 tons of arsenic (CMBEEP, 1977;
Woolson, 1983). Coal combustion also produces ash, which contains approximately 7 to
60 mg As kg- (O'Neill, 1990).
An increase in industrialization has also lead to an increase in the amount of arsenic
present in biosolids. Deposits from the atmosphere, runoff and from effluents of
industries often increases the concentration of arsenic in biosolids. Woolson (1983)
reported a range of 0 to 188 mg As kg-l dry weight of biosolids. Biosolids are often
disposed on land and may subsequently increase arsenic concentrations in the top 20 cm
of soil by up to 0.15% (O'Neill, 1990).
Remediation of Arsenic Contaminated Soils
Currently there are several options that exist for the remediation of arsenic
contaminated soils. Remediation methods vary greatly in cost, intensity and necessary
treatment length. No single soil remediation technique is suitable for all situations.
Therefore, careful investigation of the contaminated site characteristics, contaminant
problem, treatment options and treatment timeframe must be considered in order to
achieve a successful clean up of a site. Several of the current arsenic remediation
methods are summarized in Table 2-1, and several will be discussed in the following
Excavation, capping and solidiaication are three examples of physical remediation
methods. Excavation is a commonly used remediation method. It is simply the physical
removal and disposal of contaminated soil. This method produces rapid remediation
results. However, it is often expensive because of the operation, transport and special
landfill requirements (Sparks, 1995; USEPA, 2002a).
Capping is also a rather simple method. It requires covering contaminated soil with
a hard cover (i.e., concrete or asphalt) to reduce exposure. However, this method does
not remove contaminants from the soil, as the contaminants are still present in the soil
(Sparks, 1995; USEPA, 2002a).
Stabilization and solidiaication are in situ physical treatments where soil is mixed
with cement or stabilizers to create a hardened mixture. Solidiaication reduces the
mobility of arsenic in the soil. Vitrifieation is a type of solidiaication. Soil is chemically
bonded inside a glass matrix, where the arsenates become silicoarsenates. The drawbacks
to solidification and stabilization remediation techniques are that they can be relatively
costly. Also, soil conditions often dictate the feasibility of implementing these methods
(Tadesse et al., 1994; USEPA, 2002a).
Table 2-1. Summary of select current remediation technologies for arsenic-contaminated
soil (adapted from USEPA, 2002a).
Arsenic Remediation Technology Description
*Ex-situ method that removes soil from site
Excavation Contaminated soil stored in designated
Hard cover placed on soil
Reduces the mobility of arsenic in soil
Solidification and stabilization Contaminated soil is mixed with stabilizers
*Arsenic is chemically bonded inside a
Arsenates become silicoarsenates
In situ treatment
Arsenic is suspended or dissolved in a
Soil washing/Acid extraction
Water-based and ex-situ treatment
In situ method uses water, chemicals or
organic to flush soil
Soil flushing. .
Arsenic is mobilized and is collected for
removal or treatment
Uses heat to concentrate arsenic
Arsenic is volatilized and collected
Arsenic is mobilized as charged particles
by using a low-density current
Electrokinetic treatment Arsenic removed through several means,
such as electroplating and precipitation
In situ treatment
In situ method using plants to take up
Phyto rem ed iatio n/p hyto extractio n arsenic from soil
*Biomass is harvested and disposed
Chemical remediation can include soil washing/acid extraction and soil flushing.
Each of these methods utilizes chemicals to aid in the removal of arsenic from the soil.
Soil washing/acid extraction is an ex situ remediation method used to dissolve and
concentrate arsenic. The concentrated arsenic can then be disposed. Similarly, soil
flushing uses chemicals and/or water to remove arsenic from the soil. However, this
method is performed in situ, which can create concern regarding groundwater
contamination (USEPA, 2002a).
Bioremediation includes any method that uses microbes or plants for remediation,
such as: bioleaching, bioaccumulation and phytoremediation. Methods of bioremediation
often require inputs into the soil or system to enable the microorganisms and/or plants to
produce or grow properly.
Bioleaching involves the use of microbes to alter soil factors, such as pH and redox
potential, to increase the solubility of arsenic. This can be accomplished through organic
acid production. Once the arsenic becomes more mobile, it can be leached and collected
from the soil. On the other hand, bioaccumulation utilizes microbes to absorb
contaminants from the soil (Zwieten and Grieve, 1995; USEPA, 2002a).
Phytoremediation is an all-encompassing term to include any remediation method
that utilizes plants. Phytoremediation involves plants to either remove pollutants or
render them harmless in soil and water systems. This practice has been growing in
popularity because of its overall cost-effectiveness (Salt et al., 1995; Watanabe, 1997;
Kabata-Pendias and Pendias, 2001). The term phytoremediation includes several
methods, and a few will be discussed in greater details in the following section.
Plants can phytoremediate soil and/or water by degrading, removing or containing
contaminantss. The degradation of chemicals can take place in the rhizosphere or
possibly the bulk soil through plant root exudation of compounds to convert the
contaminant into non-harmful chemical forms. This is known as phytodegredation.
Sometimes the plants can take up the contaminant, at which point several things can
happen. The chemical can be transported through the plant and to the leaves, and then
volatilized via the plant' s transpiration. This is termed phytovolatilization. Another fate
of a plant-absorbed chemical is storage and sequestration somewhere in the plant (i.e.,
roots or leaves). These are termed phytoextraction (leaves) or rhizofiltration (roots)
(Raskin and Ensley, 2000; Lasat, 2002; McGrath et al., 2002). Ideally in phytoextraction,
the contaminant will be translocated to the aboveground biomass where it can be
harvested and transported off-site. Lastly, phytostabilization is the use of plants in order
to contain the contaminant by reducing its leaching potential.
There are several things that must be considered prior to the initiation of any
phytoremediation proj ect. The most important is the ability of a given plant to actually
remediate the contaminant in question. Some plants will simply tolerate a contaminant,
some will die and some will thrive. The site factors, such as soil properties, source of
contamination, extent of contamination, etc., must also be considered.
Phytoremediation is generally thought to be an inexpensive alternative to
traditional remediation technologies. This is due to the fact that often less labor and
heavy equipment is needed. For example, in phytoextraction the harvested plants are
much lighter to transport than soil (i.e., excavation). Phytoextraction is also considered
to be more aesthetically appealing than some traditional remediation technologies. The
plants may be relatively easier for the public to accept, as they are more attractive
compared to bare soils or caps. Also, in some cases, the plants can actually breakdown
the contaminantss. This is unlike many other remediation technologies where the
contaminants are simply contained and/or transported off-site (Schoor, 2002; USEPA,
2002a; Wolfe and Bjornstad, 2002).
However, phytoremediation is a fairly new technology and is very dependent on the
plant in the system. This impacts the efficiency and dependability of phytoremediation;
therefore, there are many questions or areas of concern that need to be addressed. First, if
the plant has a shallow root system it may not be able to fully remediate the soil or water
because contaminants may be out of uptake range of the roots. Second, the plants may be
limited to low or moderately contaminated sites. If the contaminant levels are too high
there is the risk of killing the plant or compromising its growth, which would impede the
remediation. Third, clean-up rate is generally much slower than traditional remediation
methods. This may pose a problem when the requirements for remediation of the soil are
more immediate. Fourth, there is not always a full understanding of the physiology,
biochemistry, uptake, etc. of the plants employed (Schoor, 2002). Therefore, it is not
always clear what is occurring between the plant and soil (i.e. volatilization or leaching).
Phytoextraction is an in situ remediation method that employs the use of plants to
remove contaminants from soil or water. The plants are able to take up the contaminant
and store it in its roots or shoots. Some plants can efficiently translocate the contaminant
to its aboveground biomass (Cunningham et al, 1997; USEPA, 2002a). In this case, the
aboveground biomass can be removed and disposed of in a properly constructed landfill
or incinerated. If the contaminant is of value, such as nickel or copper, it can be
removed, or phytomined, from the plant. Ideally, the plants used for phytoextraction are
hyperaccumulators of the contaminant in question.
Plants often contain trace concentrations of many contaminants of concern. At low
levels, plants can usually metabolize or dispose of these compounds without any
significant injury. Generally, at high contaminant concentrations in soil or water, plants
often suffer and/or die because of their inability to metabolize these harmful elements.
However, some plants can survive and/or thrive when they accumulate high
concentrations of toxic elements.
Hyperaccumulators are plants that contain more than 1000 mg kg l, or 0. 1%, of an
element or compound. Ideally, hyperaccumulators should have a high rate of
accumulation, be fast growing, and have a high production of biomass (Wantanbe, 1997;
Brooks, 1998). The concentration of the contaminant is generally very high in these
plants when grown in contaminated media. They must have both a bioconcentration
factor (BF) and transfer factor (TF) greater than one. The BF is the plant to soil ratio for a
particular contaminant, while the TF is the ratio of contaminant concentration in the plant
to the contaminant concentration in the growing media.
The fern, P. vittat L., (Fig. 2-4) is an example of a plant that removes arsenic from
soil and/or water, and it can be defined as an arsenic hyperaccumulator (Ma et al., 2001).
Ferns are lower plants, unlike many of the other identified hyperaccumulating plants,
which are dicots or monocots. For example, several of these other hyperaccumulating
plants are in the mustard family (dicots), such as Thalaspi spp. and Bra~ssica spp. Also,
many of these plants are able to hyperaccumulate a given metal, but are not very efficient
at transporting that metal from the roots to shoots, unlike the identified arsenic
hyperaccumulators (Schoor, 2002).
When a plant can hyperaccumulate a metal, such as Thlaspi spp. with zinc or
Alyssum spp. with nickel, these metals can be mined from the plant, purified and reused,
thus, increasing the value of these plants. However, arsenic is generally not an element
of great value for mining.
Figure 2-4. Pteris vittat L. growing at an arsenic-contaminated site.
Pteris vittata L.
Pteris vittat was recently discovered as the first known arsenic hyperaccumulator.
Pteris vittat is very efficient at removing arsenic from soil. It cannot only take up high
amounts of arsenic from soil and water, but it can transport arsenic very efficiently from
its roots to its fronds (Ma et al., 2001).
This fern also produces a relatively high biomass, it and is a fast-growing plant.
Pteris vittat is a perennial, and it survives winter fairly well in Florida and warmer
climates, thus increasing its value as a hyperaccumulator. It is also tolerant of full sun,
unlike many other ferns, but it also grows well under shady conditions.
Pteris vittat prefers an alkaline soil environment. This can contribute to its ability
of arsenic-hyperaccumulation because, in general, arsenic is more available at a higher
pH. Pteris vittat is also able to take up many forms of arsenic (Tu and Ma, 2002).
Because of its fast-growth and arsenic hyperaccumulation, this fern exhibits potential for
use in the phytoremediation of arsenic-contaminated soils.
After 20 weeks of growth, P. vittat accumulated arsenic of 1 1.8 to 64.0 mg kg-l
when grown in uncontaminated soil; however, it accumulated 1,442 to 7,526 mg kg-l
arsenic when grown at an arsenic-contaminated site. The arsenic concentration was
much higher in the fronds than roots. Therefore, it has both a high TF and a high BF.
Although P. vittat is capable of taking up many different arsenic species but did not
readily take up FeAsO4 and A1AsO4. These arsenic species are generally insoluble in the
soil (Tu and Ma, 2002).
After 12 weeks of growth P. vittat produced more above ground biomass in soils
containing 50 and 100 mg kg-l arsenic compared to ferns grown in the soil not
contaminated with arsenic. These results indicate that low levels of soil arsenic may
actually be beneficial to the growth of this fern. However, when the soil arsenic
concentration was 200 mg kg- there was a slight decrease in fern biomass. No
significant differences in root biomass were found at any arsenic concentration. Overall,
the mature fronds and the old fronds had the highest arsenic concentrations, while roots
had the lowest after 23 weeks of growth (Tu and Ma, 2002). Phosphorus levels in fronds
were higher in ferns grown in soil containing 50 and 100 mg kg-l arsenic versus 0 or 200
mg kg-l arsenic. Pteris vittat roots, however, had higher phosphorus concentrations in
the control soil (Tu and Ma, 2003).
Arsenic has been shown to leach from P. vittat fronds as they senesce. This may
pose a potential drawback to the use ofP. vittata in phytoremediation of arsenic
contaminated soils, as the arsenic may be returned to the soil (Tu et al., 2003).
Other Arsenic Hyperaccumulating Ferns
Since the initial i dentifi cation of P. vittata as an arseni c-hyperaccumulator, other
ferns have been identified to hyperaccumulate arsenic. However, not all ferns are able to
hyperaccumulate arsenic (Kuehnelt et al., 2000; Meharg, 2002; Visoottiviseth et al.,
2002; Zhao et al., 2002; Meharg, 2003). To date, the maj ority of ferns that do
hyperaccumulate arsenic belong to the Pteris genus. Pteris cretica, P. longifolia and P.
umbrosa have been shown to hyperaccumulate arsenic to the same extent as P. vittat
(Zhao et al., 2002). However, not all members of the Pteris genus are able to
hyperaccumulate arsenic. For example, Meharg (2003) found that Pteris tremula and
Pteris stramnina do not hyperaccumulate arsenic. In this same study, the author found that
ferns that are able to hyperaccumulate arsenic developed comparatively late,
evolutionarily speaking, for ferns.
In a study performed by Zhao et al. (2002), four other non-Pteris ferns were
examined. However, they did not exhibit any ability to hyperaccumulate arsenic.
Numerous fern species were also examined by Meharg (2003), and most of these fern
species also did not hyperaccumulate arsenic. To date the only non-Pteris fern to exhibit
this ability is Pityrogramma calomelan2os (Francesconi et al., 2002). Its fronds were able
to accumulate 2760 to 8350 mg As kg-l when grown in soil containing 135 to 510 mg As
kg- However, the authors were not able to establish a direct correlation between the
arsenic concentrations in the fronds to those in the soil. Such a correlation was seen with
P. vittat (Ma et al., 2001). Interestingly, the fronds with the greatest arsenic
concentration were collected from ferns growing in the lowest soil arsenic concentration
(135 mg As kg- ). The authors stated that Pityrogramnma calomelan2os may be readily
able to remove arsenic from soils that are less contaminated. Therefore, these ferns have
the ability to effectively reduce soil arsenic concentrations. However, the
hyperaccumulating ability under higher arsenic concentrations was not addressed. It was
also suggested that P. calomelan2os is a better phytoextraction candidate than P. vittat
because it appeared able to grow better in the arsenic-contaminated soils from which both
species were collected. However, there was no formal experimental comparison
performed to evaluate this theory.
Arsenic in Plants
Arsenic is not an essential element for plants, and it is generally considered
poisonous. However, plants have varying sensitivities to arsenic. Legumes are known to
be highly sensitive to arsenic (Adriano, 1986), while P. vittata may grow better in the
presence of arsenic (Ma et al., 2001).
Arsenic distribution within plants also varies. At high soil arsenic levels, old leaves
and roots tend to have higher arsenic concentrations. At lower soil arsenic
concentrations, plant arsenic levels are greater in leaves than in roots (Kabata-Pendias
and Pendias, 2001). However, this is not the case with P. vittat, where arsenic
concentrations are generally greater in the aboveground biomass than the roots.
Arsenic toxicity may be evident in plants in several ways. Characteristic symptoms
of arsenic toxicity in plants are: wilting of leaves, slow root growth and shoot growth,
leaf necrosis, violet leaf color and ultimately plant death (Liebig, 1965; Woolson et al.,
1971; Adriano, 1986). In general, arsenic inhibits metabolism in most plants (Kabata-
Pendias and Pendias, 2001). More specifically, arsenate can disrupt oxidative
phosphorylation and the production of ATP (Meharg and McNair, 1994, Oremland and
Stolz, 2003). Arsenite affects the function of enzymes and proteins by binding to
sulfhydryl groups (Leonard and Lauwerys ,1980, Oremland, and Stoltz 2003).
Arsenic Uptake by Plants
Plant arsenic uptake is influenced by arsenic source and solubility (Marcus-Wyner
and Raines, 1982). It is hypothesized that arsenite is taken up passively via
aquaglyceroporins, or channels allowing movement of water and neutral solutes, in the
roots. Arsenate is a chemical analogue of phosphate, and is taken up via the phosphate
transport system (Asher and Reay, 1979). However, in Holcus lan2atus L., Deschampsia
cespitose L. and Agrostis capillaries~~~1111~~~~111 L. an altered phosphorus transport system has been
found. This transport system enables these plants to be arsenic tolerant by lowering the
Vmax and affinity for arsenate uptake (Meharg and MacNair, 1990; 1991a; 1991b; 1992).
Antioxidants and Antioxidant Enzymes
Exposure of plants to arsenic and heavy metals may result in the production of
reactive oxygen species (ROS), such as superoxide anions, hydrogen peroxide (H202)
and hydroxyl radicals, resulting in damage to cell components (Weckx and Clij sters,
1996, Conklin, 2001; Hartley-Whitaker et al., 2001a). It is thought that the production of
ROS when plants are exposed to arsenic is the result of the conversion of arsenate to
arsenite (Hartley-Whitaker et al., 2001a; Meharg and Hartley-Whitaker, 2002). The
metabolism of arsenite within the plant, for example through methylation, can result in
the production of additional ROS (Zaman and Pardini, 1996).
In response to the creation of ROS, plants synthesize enzymatic and non-
enzymatic antioxidants (Meharg and Hartley-Whitaker, 2002). Through the use of
antioxidant molecules, such as L-ascorbic acid, reduced glutathione (GSH), a-tocopherols
and carotenoids, plants can manage the detrimental effects of ROS. Specifically,
ascorbic acid, which makes up approximately 10% (wt/wt) of the plant soluble
carbohydrate, is an important and major plant antioxidant due to its high abundance
(Noctor and Foyer, 1998).
As an antioxidant, ascorbic acid can manage ROS through the direct elimination of
superoxide anions, hydrogen peroxide and hydroxyl radicals (Padh, 1990). It can also act
as a secondary antioxidant through the maintenance of reduced a-tocopherol, another
plant antioxidant (Liebler et al., 1986; Noctor and Foyer, 1998; Conklin, 2001), or
obliquely through the action of ascorbate peroxidase (Foyer and Halliwell, 1976; Asada,
Under no arsenic stress, P. vittat was found to have intrinsically higher
concentrations of non-enzymatic antioxidants, ascorbate (Asc) and glutathione (GSH), in
its fronds compared to Pteris ensiformis (a non arsenic hyperaccumulator). This suggests
that the ascorbate-glutathione pool may play a significant role in the ability ofP. vittat
to tolerate and hyperaccumulate arsenic (Singh et al., unpublished).
It is also possible for plants to bind oxygen free radicals and to detoxify organic
contaminants using GSH. Through a reaction catalyzed by glutathione S-tranferases
(GSTs), GSH can respond to oxidative stress by binding the organic contaminants or their
metabolites, storage or incorporation into plant cellular components (Lamoureux et al.,
1994; Xiang and Oliver, 1998). Glutathione is composed of the amino acids, glutamate,
cysteine and glycine. Its significance lies mostly in its role as a reductant, as well as in its
ability to detoxify harmful components within a cell. Glutathione is also a precursor for
phytochelatins (PC) (Kneer and Zenk, 1992; Zenk, 1996; Pawlik-Showronska, 2001).
Therefore, GSH has been implicated in aiding plants to cope with various environmental
stresses, either directly by binding and detoxifying, or indirectly through conversion into
Glutathione forms glutathione disulfide (GSSG) after oxidation as a result of its
antioxidant activity. The GSSG can be reduced or recycled back to GSH, the reduced
form, by the antioxidant enzyme glutathione reductase (GR). There are also several
antioxidant enzymes in plants, including catalase (CAT) and superoxide dismutase
(SOD) (Xiang and Oliver, 1998).
Glutathione reductase is the enzyme that, in conjunction with NADPH, catalyzes
the reduction of GSSG to GSH (Eq. 2-1) (Carlberg and Mannervik, 1985).
GSSG + NADPH + GSH +NADP' (Eq. 2-1)
Glutathione reductase has been detected in bacteria, yeast, plants and animals. It is
essentially responsible for maintaining the GSH levels in the cell. Glutathione reductase
activity has been shown to increase in plants exposed to environmental stress. For
example, GR in Triticum durum increased due to temperature stresses (Keles and Oncel,
2002). Exposure to copper also induced GR in the roots of Phaseolus vulgaris (Gupta et
Superoxide dismutase eliminates superoxide anions (02 -), yielding oxygen and
hydrogen peroxide (H202) (Eq. 2-2).
202- + 2 H + 02 + H202 (Eq. 2-2)
Superoxide dismutase is associated with various metal cofactors: CuSOD and
ZnSOD are located in cytosol, peroxisome, plastid and root nodules; MnSOD is located
in the mitochondria; and, FeSOD is located in the plastids. Because SOD can degrade
superoxide anions, it can play a very important role in the defense of cells upon stress
(Fridovich, 1978; Fridovich, 1986; Fridovich, 1995).
Catalase, which can be found in primarily in the peroxisomes and root nodules,
converts hydrogen peroxide (H202) to water and oxygen (Eq. 2-3).
02 + H202 & H20 + V/2 02 (Eq. 2-3)
There are several forms, or isozymes, of CAT. These isozymes may respond
differently under the same conditions. For example, in Zea mays L., the activities of two
CAT isozymes (CAT-1 and CAT-2) were examined for their responses to salicylic acid
inhibition. The CAT-1 isozyme was inhibited upon exposure to salicylic acid; however,
the CAT-2 isozyme was not (Horvath et al., 2002).
A study on enzymatic antioxidants found that SOD and CAT are induced in the
fronds of P. vittata upon arsenic exposure arsenic. However, under the same conditions
they are not induced in the fronds of P. ensiformis (Srivastava et al., 2005). This further
suggests a role for antioxidant enzymes in arsenic tolerance and/or hyperaccumulation by
P. vittat. Similarly, SOD and CAT activities were found to increase in Zea mays L.
embryos upon arsenic exposure (Mylona et al., 1998).
Plants that take up heavy metals from soil or water often use phytochelatins to help
limit the toxic effects of the metals. Phytochelatins, which contain thiol (SH) groups, are
peptides in the plant with the ability to chelate heavy metals. Their general make-up is
two or more y-glutamylcysteine units that repeat and have glycine as the terminal residue.
Glutathione is a source of non-protein thiols, and is the precursor for
phytochelatins. Using GSH, the phytochelatins are synthesized by the transpeptidase
phytochelatin synthase enzyme (Kneer and Zenk, 1992; Zenk, 1996; Chen et al., 1997;
Xiang and Oliver, 1998; Rhodes et al., 1999; Pawlik-Showronska, 2001). The
synthesized phytochelatins are able to bind some metals in the cytosol, and the
phytochelatin-metal complex is transported to the plant vacuole (Rauser, 1990).
Plants have several metal-sensitive enzymes, such as alcohol dehydrogenase,
glyceraldehyde-3 -phosphate dehydrogenase and ribulose-1 ,5-diphosphate carboxylase.
Kneer and Zenk (1992) found that these enzymes were able to tolerate cadmium (Cd) at
10 to 1000 times greater concentration when it was completed with phytochelatins. They
also concluded that when heavy metals are at concentrations below the lethal level
phytochelatins completely complex them.
However, Leopold et al. (1999) found that when Silene vulgaris, a heavy metal
tolerant plant, was exposed to copper and cadmium there was no detectable heavy metal-
phytochelatin complexation. They concluded that not all plants are able to form stable
heavy-metal-PC complexes. Another study involving Silene vulgaris exposed to arsenic-
contaminated soil showed that there was a continuous accumulation of phytochelatins as
exposure time increased (Sneller et al., 1999). Higher phytochelatin levels were also
found in freshwater algae (Stigeoclonium tenue) when the metal solution pH was 8.2
versus 6.8 (Pawlik-Skowronska, 2001).
Pteris vittat was shown to have only 4.5% of its arsenic completed with
phytochelatins, as a glutathione-arsenite- phytochelatin complex (Zhao et al., 2003). In a
study by Raab et al. (2004), it was determined that the arsenic hyperaccumulator, P.
cretica, had only 1% of its arsenic completed with phytochelatins. The conclusion
reached in both studies was that the phytochelatins may act as shuttles for the arsenic for
transport in a non-toxic form through the cytoplasm and into the vacuoles. They
theorized that the vacuolar membrane may contain an arsenic-phytochelatin shuttle to aid
in this process. Therefore, phytochelatins may not be the main source of detoxification of
arsenite in arsenic hyperaccumulators.
ARSENIC SPECIATION AND TRANSPORT INT Pteris vittat L.
Plant arsenic uptake generally depends on arsenic source and solubility (Marcus-
Wyner and Raines, 1982). It has been suggested that arsenic uptake by plants is passive
and directly related to water flow (Kabata-Pendias and Pendias, 2001). Arsenic and
phosphorus are chemical analogues, thus they often compete with each other for soil
fixation sites (Adriano, 1986). It has also been hypothesized that arsenic may be taken up
as arsenate and transported by the plant via the phosphate transport system (Asher and
Reay, 1979). However, in grasses Holcus lan2atus L., Deschampsia cespitosa L. and
Agrostis capillaris~~~~1111~~~~111 L., an altered phosphorus transport system has been identified, aiding
these plants in arsenic tolerance (Meharg and MacNair, 1990; 1991a; 1991b; 1992).
Pteris vittat is able to remove large amounts of arsenic from soil (Komar et al.,
1998; Komar, 1999; Ma et al., 2001). Typical of hyperaccumulators, arsenic
concentrations in P. vittat are mostly concentrated in the fronds (Ma et al., 2001; Tu et
al., 2002; Zhang et al., 2002). This suggests efficient transport of arsenic from roots to
fronds in P. vittata.
Arsenic speciation analysis of P. vittata grown in an arsenic-contaminated soil
showed that >67% of the total arsenic in the aboveground biomass is present as the
reduced form of arsenic, arsenite, which is considered to be the more toxic form.
However, in roots only 8.3% of the arsenic is present as arsenite. The remaining arsenic
was present in the oxidized form, arsenate (Zhang et al., 2002). Tu et al. (2003) found
similar results when arsenic was supplied to the ferns in several different forms.
Regardless of the arsenic species supplied to the fern, >90% of the total arsenic in the
roots is present as arsenate, versus approximately 94% arsenite in the fronds. In both
studies, very low concentrations of organic arsenic were found in the fern, indicating that
the arsenic is being reduced in the fern. A study conducted by Wang et al. (2002)
examined the uptake kinetics of arsenate and arsenite, and the effects of phosphate on
arsenic uptake by P. vittat. However, the study did not address methylated forms of
arsenic or the form of arsenic that was transported within the plant. Therefore, no data
exist regarding the forms of arsenic that are transported in P. vittata. Water and solutes
are mostly transported via xylem in plants (Marschner, 1995), making xylem sap an
important constituent for understanding arsenic transport in P. vittat.
In addition, transport of other constituents, such as phosphorus in the xylem sap
may be impacted by, or may impact, arsenic transport. The chemical similarity between
phosphate and arsenate raises the possibility for competition. Studies have shown that
the presence of phosphate in the growing media affects the uptake and concentration of
arsenic in the fern roots and fronds (Wang et al, 2002; Tu and Ma, 2003). Therefore, the
presence of arsenic in the xylem sap may cause phosphorus deficiency in the fern.
The main objective of this study was to determine the forms of arsenic that are
transported in the fern. More specifically, this study examined how the forms of arsenic
supplied to the roots of P. vittat affect the forms of arsenic being transported to its
fronds. In addition, the effects of arsenic concentrations and species on concentrations of
inorganic phosphorus (Pi) in xylem sap were examined. The information obtained from
this study should be useful for a better understanding of the mechanisms of arsenic
translocation in P. vittat.
Materials and Methods
Pteris vittat used for the experiments were of similar age and size. Plants were
germinated from spores and grown in a mixture of commercial potting soil, sand and peat
moss until they were approximately 8 months old. Roots of each fern were washed free
of soil using deionized water before being transferred to a hydroponics system. Each fern
was placed into individual 500 ml brown plastic bottle, which contained 0.2-strength
Hoagland-Arnon nutrient solution (Hoagland and Amon, 1938). The volume was
maintained, and the plants were allowed 7 d to acclimate to the hydroponics conditions
prior to treatment. All ferns were kept in a controlled environment with 65% humidity
and day and night temperatures of 25oC and 20oC, respectively. The femns were exposed
to an 8 h light period with a light intensity of 350 Clmoles m-2 S-1
Arsenic was added at concentrations of 0, 10 or 50 mg 1- This study was divided
into two parts. In experiment A, P. vittat were treated with arsenic in the form of either
As (III), as sodium arsenite (NaAsO2), Or As (V), as sodium arsenate (Na2HAsO4 7H20).
In experiment B, organic arsenic as monomethlyarsonic acid (MMA) or dimethylarsinic
acid (DMA), was used. Arsenic treatments were added to each bottle using stock
solutions diluted with 0.2-strength Hoagland-Amon nutrient solution. Plants were
harvested three days after treatment and separated into fronds and roots. Roots were
rinsed with deionized water before analysis.
Xylem Sap Extraction
Xylem sap samples were extracted from 1 to 2 fronds of similar age and appearance
from each fern. The xylem sap was collected using a Scholander pressure chamber (Soil
Moisture Equipment Corp., Santa Barbara, CA) (Schurr, 1998). A constant and high
pressure, up to 40 bars, was applied to all fronds, and a micropipette was used to collect
the extruded xylem sap. Xylem sap samples were preserved at -80o C immediately
Chemical Analysis of Arsenic and Phosphorus
Fronds and roots ofP. vittata were dried for 24 h at 650 C, and were ground in a
Wiley Mill to pass through a 1 mm-mesh screen. The ground tissue samples (0.25 g)
were subj ected to hot block (Environmental Express, Ventura, CA) digestion using
USEPA Method 3051 for arsenic analysis. The digested plant samples were analyzed for
total arsenic using graphite furnace atomic absorption spectroscopy (GFAAS) (Perkin
Elmer SIMMA 6000, Perkin-Elmer Corp., Norwalk, CT).
Due to arsenate interference with inorganic phosphate (Pi), the determination, Pi
concentration in the xylem sap was performed using a modified molybdenum blue
method (Carvalho et al., 1998). This method employs L-cysteine to prevent arsenate
interference. Samples were analyzed at 880 nm using VIS-spectrophotometer detection
(Shimadzu UV160U, Shimadzu Corp., Columbia, MD).
Arsenic Speciation in Plant and Xylem Sap Samples
Arsenic speciation was performed on P. vittat samples from experiment A using
frond and root samples that were stored at -80oC. Arsenic was extracted ultrasonically
using a 1:1 methanol:water solution and was repeated two times for 4 h at 60oC. This
extraction method results in 85 to 100 % recovery of arsenic from the fronds. However,
arsenic extraction efficiency in the roots is approximately 60% (Zhang et al., 2002).
The combined extracts were diluted in 100 ml with deionized water; the pH of
extract was ensured to range from 5 to 9. A solid phase extraction using an arsenic
speciation cartridge (Metal Soft Center, Highland Park, NJ) was performed. The
arsenate, which was retained in the disposable cartridge, and arsenite, which is passed
through the cartridge, were separated (Meng et al., 2001). The total arsenic and arsenite
fractions were determined using GFAAS. The arsenate fraction was calculated by the
difference between the total arsenic and the arsenite fractions.
Arsenic speciation of the xylem sap for samples from experiment A was
determined by high-performance liquid chromatography coupled with hydride generation
atomic fluorescence spectrometry (HPLC-HG-AFS). The HPLC system consisted of a
P4000 pump and an AS3000 autosampler with a 100 Cll inj section loop (Spectra-Physics
Analytical, Inc. Fremont, CA). Arsenic species were separated using a Hamilton PRP-
X100 anion exchange HPLC column (250 x 4.6 mm, 10 Clm particle size) with a 0.015
mol 1-1 potassium phosphate mobile phase (pH 5.8) at a flow rate of 1 ml min- A
hyphenated HG-AFS was a P. S. Analytical Millenium Excalibur system (PS analytical,
Kent, UK) with hydride generation sample introduction. The outlet of the column was
connected to a Teflon reactor and mixed with 12.5% HC1, then with the reductant
solution containing 14 g NaBH4 and 4 g NaOH in 1000 ml DDI water. The arsine gas
produced was separated through a gas-liquid separator and sent to an integrated atomic
fluorescence system for arsenic concentration detection. Data were acquired by a real-
time chromatographic control and data acquisition system. Arsenic was quantified
through external calibration with standard solutions containing arsenite, arsenate, MMA
and DMA. The lower detection limits for the HPLC-HG-AFS were approximately 1.0 Clg
1- for arsenite, 3.0 Clg 1-1 for MMA and DMA and 2.5 Clg 1-1 for arsenate. Quality
assurance was obtained through the use of blanks, standard curves, standard check
solutions and spiked samples, which were run during sample analysis.
Arsenic speciation of xylem sap for samples from experiment B was determined by
coupling HPLC to inductively coupled plasma mass spectrometry (ICP-MS) (Chen et al.,
2004). A VG Plasma Quadrupole II (VG Elemental, Winsford, Cheshire, UK) ICP-MS
was used. The sample was injected via a peristaltic pump (Rainin, Woburn, MA) to a
Meinhard TR-30-A concentric nebulizer (Precision Glassblowing, Englewood, CO). The
HPLC system was composed of a SpectraSYSTEM P2000 Binary gradient pump
(Thermo Separation Production Inc., Fremont, CA), an Auzx 210 inj ector valve with a 20
Cll loop and a Haisil 100 (Higgins Analytical Inc., Mountain View, CA) C18 column
(150x4.6 mm, 5 Clm particle size). The mobile phase contained 10 mM
hexadecyltrimethyl ammonium bromide (CTAB) as the ion-pairing reagent, 20 mM
ammonium phosphate buffer, and 2% methanol at pH 6.0. Arsenic was quantified
through external calibration with standard solutions containing arsenite, arsenate, MMA
and DMA, which were prepared daily. Lower detection limits for the HPLC-ICP-MS
were 0.5, 0.4, 0.3 and 1.8 Clg 1-1 for arsenite, DMA, MMA and arsenate, respectively.
Quality assurance measures were the same as those used for HPLC-HG-AFS detection
Experimental Design and Statistical Analysis
Experiments A and B employed a randomized complete block design with 4
replications. All data were analyzed using the General Linear Model (GLM) with the
Statistical Analysis System (SAS Institute, 2001).
This experiment was designed to determine the effects of arsenic concentrations
and species on the arsenic concentrations and species, and concentration of inorganic
phosphate in the xylem sap of P. vittata. Three arsenic concentrations, 0, 10 and 50 mg
1- and four arsenic species, arsenite, arsenate, MMA and DMA, were used during the 3-
d hydroponics experiments.
Arsenic Concentration and Speciation in Roots and Fronds
In this experiment, the arsenic concentration in the fronds and roots was directly
proportional to the arsenic concentration supplied. Plants treated with 50 mg As 1-1 had
the highest frond and root arsenic concentrations compared to the control and 10 mg 1-1
treatment (Fig. 3-1 A and B). No significant differences were found in plant arsenic
concentrations (fronds or roots) between the arsenite and arsenate treatments. However,
ferns treated with 50 mg As 1-1 as MMA had the highest frond arsenic concentrations
compared to the DMA treatments.
Compared to the roots, there was a higher, but not significant, level of arsenic
concentrated in the fronds (Fig. 3-1 A and B). However, in ferns treated with 10 mg As
1- as DMA or MMA arsenic was distributed evenly between the fronds and the roots.
Compared to ferns treated with 10 mg As 1-1 as As(III) or As(V), arsenic concentrations
in the fronds treated with 10 mg 1-1 DMA or MMA were approximately the same (92.5 to
0As 10DMA 10 MMA
50 DMA 50 MMA
Figure 3-1. Total arsenic concentrations in the fronds and roots of P. vittat exposed to 0,
10 or 50 mg 1-1 arsenic as As(III), As(V), MMA or DMA. (A) arsenic supplied
as As(III) or As(V). (B) arsenic supplied as DMA or MMA. Values represent
means + std. dev.
0 As 10 As(lll) 10 As(V) 50 As(lll) 50 As(V)
109 mg kg- ). However, arsenic concentrations in the roots treated with 10 mg As 1-1 as
DMA or MMA were higher than those treated with 10 mg As 1-1 as As(III) or As(V). In
other words, more arsenic remained in the roots when supplied with organic arsenic than
inorganic arsenic. Such a trend was not observed when the arsenic treatment was
increased to 50 mg 1- i.e., more arsenic was concentrated in the fronds.
Figure 3-2. Percentages of As(III) and As(V) in the fronds and roots of P. vittata exposed
to As(III) or As(V). No arsenic was detected in the roots of the control plants.
Values represent means.
Almost all of the arsenic in the roots was present as arsenate, except in the 50
As(III) treatment (Fig. 3-2). Even when supplied with 50 mg 11 As(III), 84% of the
arsenic in the roots was present as arsenate. Root speciation data for the 0 As treatment
are not presented in Figure 3-2 because the arsenic concentration was below the detection
Arsenic Concentration and Speciation in Xylem Sap
As with fronds and roots, arsenic concentrations in hydroponics solution
significantly (P < 0.05) affected the total arsenic concentrations in the xylem sap, with
greater treatment arsenic concentrations resulting in greater arsenic concentrations in the
xylem sap. However, there were no significant differences in the total arsenic
concentrations in the xylem sap of ferns treated with different arsenic species (Table 3-1).
Although the total sap arsenic concentration was not affected by arsenic species, arsenic
concentration in the xylem sap was greatest when the fern was supplied with a
concentration of 50 mg As 1-1 in either experiment.
In experiment A, the total concentration of arsenic in the xylem sap for the arsenite
treatments were 2.5 and 29 mg As 1- for the 10 and 50 mg 1-1 treatments, respectively.
The total arsenic xylem sap concentrations for the 10 and 50 mg 1-1 As(V) treatments
were approximately two times greater compared to the arsenite treatments.
For experiment B, the 10 mg 1-1 As treatment concentrations yielded the same
xylem sap total arsenic concentrations for both MMA and DMA. However, the 50 mg 1-1
DMA treatment resulted in a 3.5 times greater xylem sap concentration compared to the
50 mg 1-1 MMA treatment.
In the 50 mg 1-1 As(V) treatment, the total arsenic concentration of the xylem sap
was approximately 1.25-fold greater than that of the arsenic concentration of the
treatment solution. However, the 50 mg 1-1 MMA treatment xylem sap was only 0.26 that
of the treatment solution.
Table 3-1. Total arsenic concentrations in xylem sap of P. vittata exposed to 0, 10 or 50
mg 1-1 arsenic as As(III), As(V), MMA or DMA. Values represent means f
Solution arsenic concentrations (mg I 1)
Solution arsenic species 10 50
Control 0.6 f 0.4 0.6 f 0.4
As(III) 2.5 f 0.8 29.4 f 10.1
As(V) 4.1 f 1.9 60.7 f 25.4
MMA 5.4 f 2.8 13.0 f 6.5
DMA 5.6 f 2. 1 44.7 f 18.3
In experiment A, no methylated forms of arsenic were found in the xylem sap of
ferns exposed to arsenate or arsenite. Although more arsenic in these treatments was
transported as arsenate, it was only significant in the 50 mg 1-1 As(V) treatment, where
the xylem sap consisted of 57 mg 1-1 As(V) versus 3 mg 1-1 As(III) (Fig. 3-3 A).
Pteris vittat exposed to MMA and DMA in experiment B transported arsenic
primarily in the form it was supplied (Fig. 3-3 B). However, small concentrations of
arsenate and arsenite were detected in the xylem sap of these ferns.
Phosphorus Concentration in Xylem Sap
Inorganic phosphorus concentrations in the xylem sap ranged from 5.2 to 13.4
mg 1- However, the phosphorus concentrations in the xylem sap were not significantly
affected by arsenic concentration or arsenic species supplied in the nutrient solutions
(Fig. 3-4), nor were the phosphorus concentrations significantly different between
experiments A and B.
0 As 10 As III 10 As V 50AsIII 50As V
0 As 10 MMA 10 DMA 50 MMA
Figure 3-3. Concentrations of As(III), As(V), DMA and MMA in the xylem sap of P.
vittat exposed to 0, 10 or 50 mg I~ of (A) As(III) and As(V), and (B) DMA
and MMA. No DMA or MMA were detected in the xylem sap of ferns
exposed to As(III) or As(V). Values represent means + std. dev.
0 As 10 As III 10 As V 50 AsIII 50 As V
0 As 10 MMA 10 DMA 50 MMA 50 DMA
0 Total As
0 Total As
Figure 3-4. Comparison of total arsenic and Pi (inorganic phosphorus) concentrations in
the xylem sap of P. vittata. (A) arsenic supplied as As(III) or As(V) and (B)
arsenic supplied as DMA or MMA. The Pi concentration in the xylem sap
was not significantly affected by arsenic regardless of the form or
concentration of arsenic supplied to the ferns. Values represent means + std.
Previous studies (Ma et al., 2001; Tu et al., 2002; Wang et al., 2002; Zhang et al.,
2002) have shown that arsenic concentrations in P. vittat increase with external arsenic
concentrations, and the maj ority of the arsenic is concentrated in the fronds. This also
was confirmed by these experiments.
Research by Tu et al. (2003) showed arsenite was the predominant species present
in the fronds, and arsenate was the predominant species in the roots, with little organic
arsenic being detected in the fern. Similar results were obtained in this experiment.
Inorganic arsenic species found in the fronds and roots of P. vittat from experiment A
were not significantly affected by the arsenic species supplied to the fern. With the
exception of the 50 As(III) treatment, root arsenic was present entirely as arsenate despite
that different forms of arsenic were supplied to ferns (Fig. 3-2). However, fronds
contained 50-80% arsenite. Again, these results confirm those of previous studies (Zhang
et al., 2002; Tu et al., 2003; Webb et al., 2003), where the predominant forms of arsenic
in P. vittat fronds and roots are arsenite and arsenate, respectively. Similarly, in a study
involving Pityrogramnma calomelan2os, another arsenic-hyperaccumulating fern, most of
the arsenic found in its fronds was arsenite. Only trace amounts of MMA and DMA were
found in a few samples (Francesconi et al., 2002). Pteris vittata is apparently reducing
arsenic at some point between its presence as arsenate in the roots until it is stored as
arsenite in the fronds.
Arsenic speciation analysis showed that 56-60% of the arsenic was present as
arsenate in the hydroponics solution treated with 10 or 50 mg 11 As(III) after the 3-d
experiment (data not shown). Although this indicates that significant arsenic oxidation
occurred in the hydroponics solution, a substantial level (40-44%) of arsenic was present
as arsenite at the end of the experiment. Assuming arsenite and arsenate were taken up
by the plant at the same rate, as suggested in Figure 3-1, then 56-60% of the arsenic
should be present as arsenate in the roots. The fact that 84-100% of the arsenic was
present as arsenate in the roots treated with 10 or 50 mg 1-1 As(III) suggests that either
arsenic oxidation occurred inside the roots and/or arsenite was preferentially transported
from the roots to the fronds.
A study by Wang et al. (2002) determined that arsenite was translocated more
efficiently than was arsenate from P. vittata roots to its fronds. Similar results were found
in Arabidopsis thaliana using phosphate mutants, phol and pho2. A study of arsenic
uptake and translocation in these mutants suggests that arsenite is the form preferentially
loaded into the xylem (Quaghebeur and Rengel, 2004). This may not be the case in P.
vittata, considering these findings of a slightly greater concentration of arsenate in the
However, it may not be feasible to assume that uptake of arsenite and arsenate by
P. vittata roots is equal because the species may be taken up through different systems in
the roots. Wang et al. (2002) found that arsenate was taken up more quickly by P. vittat
than was arsenite, especially in the absence of phosphate. The authors suggest that this is
due to arsenate being taken up via phosphorus-suppressible uptake in the roots. Meharg
and Jardine (2003) suggest that aquaglyceroporins are the main inlet for arsenite into rice.
Therefore, the root uptake rates of arsenite and arsenate into the plant roots are likely
In experiment A the arsenic concentration in the xylem sap was greatest when the
fern was supplied with 50 mg 1-1 As(V); therefore, it is possible that arsenic is more
readily concentrated in the xylem sap when the fern is supplied with arsenate. Such a
finding would disagree with previous conclusions for P. vittata (Wang et al., 2002) and
Arabidopsis thaliana (Quaghebeur and Rengel, 2004). However, if this were the case,
i.e., arsenite, was not preferentially transported in the xylem sap, then the fact that greater
arsenate was observed in the roots than that in the hydroponics solution may suggest
oxidation of arsenite to arsenate inside the roots. Also because arsenate and phosphate
are similar, it is conceivable that this difference may be due to arsenate being taken up
via the phosphate uptake system.
A weak correlation was found between arsenic concentrations in the xylem sap and
arsenic concentrations in the fronds (r = 0.50). This implies that the amount of arsenic
accumulated in the fronds was affected by arsenic species, in addition to arsenic
concentration. A slightly stronger correlation was found between arsenic concentrations
in the xylem sap and arsenic concentrations in the roots (r = 0.66), i.e., greater root
arsenic concentrations, resulting in greater xylem sap arsenic concentrations.
Interestingly, the arsenic concentration in the fronds treated with 50 mg 1-1 MMA
was the greatest (627 mg kg- Fig. 3-1). However, the arsenic concentration in the xylem
sap of the ferns treated with 50 mg 1-1 As(V) was the greatest (60.7 mg 1- Table 3-1).
Therefore the highest arsenic concentration in the xylem sap of ferns treated with
arsenate did not translate into the highest arsenic concentration in the fronds.
In experiment A, the predominant form of arsenic transported in P. vittat xylem
sap appeared to be arsenate, regardless of the species supplied in the external nutrient
solution (Fig. 3-3 A). This was consistent with the root data, where most of the arsenic
was present as arsenate, i.e., 84-100% even in plants treated with 50 mg 1-1 As (III) (Fig.
3 -2). The arsenate concentrations in the roots were correlated with the arsenate
concentrations in the xylem sap with r = 0.74. This suggests that greater arsenate
concentrations in the roots resulted in greater arsenate in the xylem sap.
Although 84% of the arsenic was present as arsenate in the fern roots treated with
50 mg 1-1 As(III) (Fig. 3-2), only 59% of the arsenic in the xylem sap was present as
arsenate (Fig. 3-3 A), indicating that proportionally more arsenite than arsenate was
present in the xylem sap. This, however, contradicted the fact that the highest arsenic
concentration was observed in the xylem sap of plants treated with 50 mg 1-1 arsenate
instead of arsenite (Table 3-1). This may be explained by the fact that some of the
arsenate was reduced to arsenite during the transport. This is supported by the fact that,
though all of the arsenic in the roots treated with 50 mg 1-1 As(V) was present as arsenate
(Fig. 3-2), approximately 5.3% of the arsenic in the xylem sap was present as arsenite
(Fig. 3-3 A); this suggests that a small amount of arsenic reduction occurred during the
transport in P. vittat. However, most of the arsenic reduction occurred mainly in the
pinnae of the fern. Because arsenate can compete for phosphate sites, such as ATP,
within the plant, it is important for the arsenic reduction to occur. It is thought that thiol-
containing compounds may sequester some arsenite and shuttle it to the frond vacuoles to
limit the arsenic toxicity (Lombi et al., 2002; Webb et al., 2003). No methylated forms of
arsenic were detected in the xylem sap when arsenite or arsenate was supplied. Therefore,
arsenic methylation does not occur prior to or during arsenic transport in P. vittat.
In experiment B, the species of arsenic transported was strongly correlated with the
arsenic species that was supplied to the fern (Fig. 3-3 B). For example in those ferns
supplied with DMA, arsenic was transported mainly as DMA. There were also low
concentrations of arsenite and arsenate detected in the sap when DMA and MMA were
fed to the fern. In general, inorganic forms of arsenic, such as arsenite and arsenate, are
considered more toxic than organic forms (Tamaki and Frakenberger, 1992). Also,
monosodium methanearsonate (MSMA) has been shown to be quickly absorbed by plant
leaves and move into the symplast (Wauchope, 1983). Therefore, it may be easier for the
femn to transport arsenic in methylated form, rather than to demethylate it for transport.
Overall, the fern appears to transport arsenic in the form least harmful to itself, regardless
of the species in which it is supplied. In previous studies, little or no methylated species
have been detected in P. vittata fronds when supplied with DMA or MMA (Tu et al.,
2003), suggesting that demethylation of the arsenic may be occurring in the pinnae.
Since phosphate and arsenate are chemical analogues, it is reasonable to expect
competition between the two during their transport in the femn (Tu and Ma, 2003). The
study by Wang, et al. (2002) showed that arsenic concentrations in the fronds and roots of
P. vittata decreased with increasing phosphate concentrations in the nutrient solution.
Therefore, it was thought that a similar phenomenon would take place with phosphorus
the in the xylem sap, when various concentrations of arsenic were supplied to the femn. Tu
and Ma (2003) found that phosphate might mitigate the phytotoxicity of arsenic in the
femn. At a concentration of 2.67 mM As kg-l of soil, arsenate even increased phosphate
uptake. At a higher arsenic concentration, 5.34 mM As kg-l of soil, phosphate
concentrations decreased. However, in the xylem sap, the presence of arsenic did not
seem to affect phosphorus concentration, and vice versa (Fig. 3-4). Therefore,
phosphorus and arsenic are probably not competing for transport within the xylem sap at
the concentrations used in this study. The concentration of phosphorus in 0.20-strength
Hoagland-Arnon solution is 120 mg 1- Therefore, the ratio of phosphorus to arsenic was
12: 1 and 2.4: 1 for the 10 and 50 mg 1-1 arsenic treatments, respectively. Lower ratios of
phosphorus to arsenic may have resulted in competition between the two elements and
lower concentrations of phosphorus in the xylem sap.
EFFECTS OF ARSENIC ON GLUTATHIONE REDUCTASE AND CATALASE INT
THE FRONDS OF Pteris vittat L.
It has been well established that P. vittat is able to accumulate very high
concentrations of arsenic in its fronds. However, it is presently unclear as to how this
fern tolerates such high concentrations of arsenic. Arsenic-exposure in plants may result
in the production of reactive oxygen species (ROS), such as superoxide anions, hydrogen
peroxide (H202) and hydroxyl radicals, which can generate significant cell damage
(Weckx and Clijsters 1996; Conklin, 2001; Hartley-Whitaker et al., 2001b).
A study on enzymatic antioxidants found that catalase (CAT), the antioxidant
enzyme that converts H202 to water and oxygen, was induced in the fronds of P. vittat
exposed to arsenic. However, under the same conditions they are not induced in the
fronds of Pteris ensiformis, a non-arsenic hyperaccumulator (Srivastava et al., 2005).
This further suggests a role for antioxidant enzymes in arsenic tolerance and/or
hyperaccumulation by P. vittat. Also, the reduction of arsenate to arsenite can result in
the production of ROS, such as H202, pOssibly resulting in the need for increased CAT
Glutathione reductase (GR) is the enzyme that, in conjunction with NADPH,
catalyzes the reduction of glutathione disulfide (GSSG) to glutathione (GSH) (Carlberg
and Mannervik, 1985). This antioxidant enzyme has been detected in bacteria, yeast,
plants and animals. It is essentially responsible for maintaining cellular GSH levels.
Glutathione is composed of the amino acids, glutamate, cysteine and glycine. Its
significance lies mostly in its role as a reductant, as well as in its ability to detoxify
harmful components within a cell. Glutathione is a source of non-protein thiols, and it is
also a precursor for phytochelatins (PC) (Kneer and Zenk, 1992; Zenk, 1996; Pawlik-
Showronska, 2001). Therefore, GSH has been implicated in aiding plants to cope with
various environmental stresses, either directly by binding and detoxifying, or indirectly
through conversion into PCs.
Glutathione reductase and GSH have both been the subj ect of numerous
experiments investigating their roles in plant tolerance of various environmental
conditions, such as photooxidation, water, temperature and heavy metal stresses (Aono et
al., 1997; Gupta et al., 1999; Vitoria et al., 2001; Jiang and Zhang, 2002; Keles and
Oncle, 2002; Piquey et al., 2002). A study conducted by Aono et al. (1997) indicated that
transgenic tobacco (Nicotiana tabacum L.) plants with high GR activity exhibited a
decreased sensitivity to photooxidative stress as a result of exposure to the herbicide
paraquat. Similarly, Zea mays L. subj ected to water stress showed higher GR activity, as
well as other antioxidant enzymes, when compared to non-stressed plants (Jiang and
Zhang, 2002). Vitoria et al. (2001) studied the effects of cadmium on radish (Raphanus
sativus L.) GR activity. It was determined that GR activity increased in the radishes after
24 h exposure to cadmium. The authors concluded that the main response of the radish to
cadmium was in its activation of the ascorbate-GSH cycle to remove H202, and that an
alternative response may be to make GSH available for cadmium-binding protein
synthesis. Glutathione reductase activity was found to also increase with exposure to
copper in Pha~seohes vulgaris L. (Gupta et al., 1999). Temperature stress (Keles and
Oncle, 2002) and salt stress (Bor, et al., 2003) were also shown to result in an increased
response of GR.
Because of the production of ROS in the presence of arsenic, it is essential to
understand the role that these important antioxidants play in P. vittat plants subj ected to
arsenic stress. An increase or stimulation of GR activity may lead to an increase in the
GSH levels in cells, thereby contributing to the ability of a plant to interact with free
radicals, detoxify heavy metals or contaminants and/or overcome other generally
unfavorable environmental conditions. However, it is critical to know that there are
many antioxidants and antioxidant enzymes, such as ascorbate, ot-tocopherol, catalase,
xanthophylls and carotenoids, which also aid plants in dealing with environmental
The obj ectives of this study were 1). to determine and compare the apparent
enzyme kinetics (Km and Vmax) of antioxidant enzymes GR and CAT in the fronds ofP.
vittat and P. ensiformis; 2). to determine if the presence of arsenic inhibits the activities
of these enzymes in both Pteris species; and 3). to determine if these enzymes are
induced in the fronds of P. vittat upon arsenic exposure.
Materials and Methods
Plant and Chemical Materials
Pteris vittat and P. ensiformis ferns were used in the following experiments. All
ferns were produced in the laboratory to ensure uniformity. All chemicals were supplied
by Fisher Scientific (Pittsburgh, PA USA) or Sigma (St. Louis, MO USA), unless
Fresh frond material was homogenized in a chilled mortar containing sea sand and
extraction buffer (100 mM Tris-HCI pH 8.0, 2 mM EDTA, 5 mM DTT, 10% glycerol,
100 mM sodium borate, 4% w/v insoluble PVPP and protease inhibitors: 0.5 mM
leupeptin, 20 mM AEBSF, 100 CIM pepstatin A, 100 C1M bestatin, 100 C1M E-64 and 100
mM 1, 10 phenathrolin). The homogenate was filtered through cheesecloth and
centrifuged for 20 min at 20,000g and 4oC. The crude supernatant was collected, its
volume estimated and 2 ml were reserved. To the crude supernatant, 5% PEG (w/v) was
added. The solution was incubated for 20 min and centrifuged for 20 min at 20,000g and
4oC. The 5% PEG supernatant was collected and its volume recorded. Additional PEG
was added to obtain a final concentration of 20% (w/v). After incubation for 20 min, the
20% PEG fraction was centrifuged for 40 min at 20,000g and 4oC. The 20% PEG
fraction pellet was redissolved in redissolve buffer (50 mM Tris-HCI pH 8.0, 5 mM DTT
and 10% glycerol). All protein fractions were stored at -80oC until analysis.
Protein and Enzymatic Activity Determinations
Protein concentrations were estimated in the various fractions using the method of
Lowry et al. (1951) as modified by Peterson (1977). Bovine serum albumin (BSA) was
used as a standard.
Glutathione reductase (EC 22.214.171.124) activity was assayed by following NADPH
activity at 340 nm on a UV-spectrophotometer (Beckman DU@-520 UV/VIS
Spectrophotometer, Beckman Coulter, Inc. Fullerton, CA, USA) for 5 min in 1 ml of an
assay mixture. The assay contained 50 mM potassium phosphate buffer (pH 7.0), 2 mM
EDTA, 0. 15 mM NADPH, 0.5 mM GSSG and 50 CIM of enzyme extract. The reaction
was initiated by the addition of NADPH. Glutathione reductase activity was calculated
using the extinction coefficient of NADPH at 340 nm (6.2 mM1 cm- ) (Jiang and Zhang,
Catalase (EC 126.96.36.199) activity was assayed by following the decrease in
absorbance, or degradation of H202, at 240 nm for 3 min with a UV-spectrophotometer.
The assay contained 50 mM potassium phosphate buffer (pH 7.0), 88 CIM H202 and
approximately 50 Clg protein. The reaction was initiated by the addition of H202-
Catalase activity was calculated using the extinction coefficient of H202 at 240 nm (40
mM1 cm- ) (Chance and Machly, 1955).
Enzyme Induction Study
Pteris vittat ferns of similar age/size (approximately 90 d old plants) were placed
in a hydroponics system and acclimated for 7 d using 0.2 strength Hoagland-Arnon
solution (Hoagland and Arnon, 1938). The ferns were kept in a controlled environment
with 65% humidity and day and night temperatures of 25oC and 20oC, respectively. The
ferns were subj ected to an 8 h light period with a light intensity of 3 50 Clmoles m-2 S-1
Arsenic in the form of sodium arsenate (Na2HAsO4 7H20), was added to a 0.2 strength
Hoagland-Arnon solution with final concentrations of0O and 10 mg 1- The experimental
design was a randomized complete block which consisted of three replications.
After 3 d, fern fronds were harvested, flash frozen in liquid nitrogen and stored at
-80oC until analysis. Protein extraction, determination and GR and CAT enzymatic
assays (same as above) were performed on the 20% PEG fractions of the 0 and 10 mg L^1
treatments. Activities of both GR and CAT were also determined in mixed samples using
approximately equal amounts of frond tissue from both 0 and 10 mg 1-1 plants.
The induction or lack of induction of GR was further confirmed through
immunoblotting. A SDS-PAGE was performed in a 12% (w/v) separation gel using the
methods of Laemmli (1970). Coomassie Brilliant Blue stain was used to visualize
proteins. Following activity staining, proteins were transferred from the 12% SDS-
polyacrylamide gel to nitrocellulose electrophoretically (Mini Trans-Blot@
Electrophoretic Transfer Cell, Bio-Rad Laboratories). After blocking, the blot was
incubated with a 1/1500 dilution [IgG fraction diluted in Tris-Buffered Saline + Tween
20 (TBST)] of a Zea mays L. cystolic GR antibody (Pastori et al., 2000) for 15 h at 4oC.
The blot was then washed four times with TB ST and incubated for 1 h at 4oC with a
1/3000 dilution of alkaline phosphatase conjugated with of anti-rabbit IgG. The
phosphatase conjugate was detected by nitro-blue tetrazolium (NBT) and 5-bromo-4-
chloro-3-indoyl phosphate (BCIP).
Determination of Apparent Kinetics
The apparent Michaelis-Menten enzyme kinetic parameters, Vmax and Km were
determined for both GR and CAT in P. vittata and P. ensiformis using the 20% PEG
fractions. The assay procedures for GR and CAT were similar to those described above
except that substrate concentrations were varied as indicated. All assays were performed
The apparent kinetics for GR were determined for both GSSG and NADPH. The
NADPH concentration was fixed at a saturating concentration during GSSG kinetics
determination, and the GSSG concentration was fixed at a saturating concentration during
NADPH kinetics determination. For CAT, the apparent kinetics ofH202 were
Data were plotted using Lineweaver-Burk plots (double reciprocal plots). The
apparent kinetic parameters were derived from x (-1/Km) and y (1/Vmax) intercepts of the
Determination of Arsenic Effects on Enzyme Activities
Inhibition and/or activation of GR and CAT activities by arsenic in P. vittata and
P. ensiformis were examined. Various concentrations of arsenic were added directly to
assays immediately prior to initiation of the enzymatic reaction. For GR, both arsenate,
as sodium arsenate, and arsenite, as sodium arsenite, were examined. However, for CAT,
only arsenate, in the form of sodium arsenate, was used to examine inhibition or
activation. This is because arsenite will be oxidized to arsenate upon exposure to H202
(Aposhian et al., 2003). Therefore, the addition of arsenite would give false activities
and/or yield similar results to arsenate. Effects of arsenate on the CAT activity of
purified protein CAT positive control (bovine liver) was also examined for comparison.
Glutathione Reductase and Catalase Induction Study
Spectrophotometric assays of GR activity indicated that GR in P. vittat was not
induced upon exposure to arsenic (Fig. 4-1). These results were confirmed with the GR
immunoblot (Fig. 4-2). However, CAT activity increased approximately 1.5 times when
ferns were exposed to arsenate (Fig. 4-3).
Figure 4-1. Glutathione reductase activity in P. vittata plants exposed to 0 and 10 mg I~
arsenic, and the GR activity of an extraction mixture of consisting of equal
amounts of frond tissue of both arsenic treatments. Values represent means +
std. dev. (n = 3).
Figure 4-2. Immunoblot of GR activity in (A) crude extract of arsenic treated P. vittata
(B) crude extract of control P. vittat and (C) crude extract of Zea mays.
Arrows indicate GR bands.
0 As 10 As Mix
Figure 4-3. Catalase activity in P. vittata plants exposed to 0 and 10 mg 1Y arsenic, and
the CAT activity of an extraction mixture of consisting of equal amounts of
frond tissue of both arsenic treatments. Values represent means + std. dev. (n
Glutathione Reductase and Catalase Apparent Kinetics
The GR activities exhibited Michaelis-Menten kinetics with respect to the substrate
saturation response. The responses for varying concentrations of GSSG and NADPH are
shown in Figures 4-4 A and 4-6 A for P. vittata and Figures 4-5 A and 4-7 A for P.
ensiformis, respectively. Although the reactions catalyzed by H202 for CAT did appear
to exhibit Michaelis-Menten kinetics for the substrate concentrations used, substrate
saturation was not reached for either species (Fig. 4-8 A and 4-9 A). Higher H202
concentrations could not be used accurately in the spectroscopic assays.
There were no significant differences found between the apparent kinetic constant,
Km, of P. vittat and P. ensiformis for the substrates, as determined from the Lineweaver-
Burk plots (Fig. 4-4 B, 4-5 B, 4-6 B, 4-7 B, 4-8 B and 4-9 B). The values for Vmax of GR
were also comparable between the two species. However, the Vmax of CAT activity in P.
O 50 100 150 200 250 300
ensiformis was approximately an order of magnitude greater than that in P. vittata. The
Km and Vmax values are summarized in Table 4-1.
Figure 4-4. Apparent kinetic analysis of substrate, GSSG, for GR activity in P. vittat.
(A) Direct plot showing the dependence of GR velocity on GSSG
concentration. (B) Lineweaver-Burk (double reciprocal) plot. Substrate
GSSG concentrations varied between 5 and 300 CIM. The NADPH
concentration was maintained at 200 CIM. Values represent means + std. dev.
(n = 3).
0 50 100 150 200 250 300
y = 0.2047x + 0.0166
R2 = 0.9493
Figure 4-5. Apparent kinetic analysis of substrate, GSSG, for GR activity in P.
ensiformis. (A) Direct plot showing the dependence of GR velocity on GSSG
concentration. (B) Lineweaver-Burk (double reciprocal) plot. Substrate
GSSG concentrations varied between 8 and 300 CIM. The NADPH
concentration was maintained at 50 CIM. Values represent means + std. dev.
(n = 3).
0.2 0.4 0.6 0.8
Figure 4-6. Apparent kinetic analysis of substrate, NADPH, for GR activity in P. vittat.
(A) Direct plot showing the dependence of GR velocity on NADPH
concentration. (B) Lineweaver-Burk (double reciprocal) plot. Substrate
NADPH concentrations varied between 1 and 50 CIM. The GSSG
concentration was maintained at 100 CIM. Values represent means + std. dev.
(n = 3).
Figure 4-7. Apparent kinetic analysis of substrate, NADPH, for GR activity in P.
ensiformis. (A) Direct plot showing the dependence of GR velocity on
NADPH concentration. (B) Lineweaver-Burk (double reciprocal) plot.
Substrate NADPH concentrations varied between 1 and 50 CIM. The GSSG
concentration was maintained at 100 CIM. Values represent means + std. dev.
(n = 3).
0 10 20 30 40 50
y = 0.0583x + 0.0)156
R2 = 0.9929
y =3.304x+ 0.0393
R2 = 0.9873
0 0.02 0.04 0.06 0.08 0.1 0.12
Figure 4-8. Apparent kinetic analysis of H202 for CAT activity in P. vittat. (A) Direct
plot showing the dependence of GR velocity on H202 COncentration. (B)
Lineweaver-Burk (double reciprocal) plot. Substrate H202 COncentrations
varied between 8.8 and 220 mM. Values represent means + std. dev. (n = 3).
0 50 100 150 200 25
0.00 0.02 0.04 0.06 0.08 0.10 0.12
Figure 4-9. Apparent kinetic activity of H202 for CAT activity in P. ensiformis. (A)
Direct plot showing the dependence of GR velocity on H202 COncentration.
(B) Lineweaver-Burk (double reciprocal) plot. Substrate H202 COncentrations
varied between 8.8 and 220 mM. Values represent means + std. dev. (n = 3).
Table 4-1. Summary of apparent kinetic parameters for GR and CAT. Values were
derived from Lineweaver-Burk (double reciprocal) plots for GR substrates
(GSSG and NADPH) and CAT substrate (H202) meaSured in P. vittata and P.
Pteris vittata Pteris ensiformis
Km Vmax Km Vmax
(C1M) (Cpmoles mg protein' (C1M) (Cpmoles mg protein'
Glutathione reductase min ) min )
GSSG 13.8 51.0 12.3 60.2
NADPH 4.7 78.1 3.7 64.1
H202 84.7 25.4 86.7 185.2
Effect of Arsenic on Enzyme Activities
Single replicate assays over a range of arsenate and arsenite concentrations, O to
500 mM, did not reveal inhibition or activation of GR activity in either P. vittat or P.
ensiformis fronds (data not shown). Significant inhibition was not observed in either
plant species until 1 mM arsenite was added to the assay. At 1 mM arsenite, GR activity
was inhibited approximately 64% in both P. vittat and P. ensiformis. Arsenate
concentrations up to 3 mM did not inhibit GR activity. To briefly confirm the lack of
inhibition by arsenite, triplicate values of three concentrations, 0, 25 and 250 C1M, were
assayed for both species (Fig. 4-10).
Arsenate did not inhibit CAT activity in P. vittata, P. ensiformis or bovine liver
positive control (Fig. 4-11i, 4-12 and 4-13). However, the addition of arsenate did appear
to activate CAT activity in P. vittat (Fig. 4-11). Catalase activity increased 175%,
relative to the control in P. vittat, at 10 C1M sodium arsenate (Fig. 4-14). Activity
returned to a similar velocity as the control assay when a concentration of 20 C1M sodium
arsenate was added. Activity increased again and reached a maximum, approximately
300% that of the control, at a concentration of 200 CIM sodium arsenate. The CAT
activity returned to a similar level to the control upon the addition of 500 CIM sodium
Pteris ensifornzis (Fig. 4-12) and the bovine liver (CAT positive control) (Fig. 4-
13) showed similar patterns as P. vittat. However, P. ensifornzis and bovine liver
maximum relative increases in activity were only 133% and 120%, respectively (Fig. 4-
14). The maximum activity for P. ensifornzis was obtained at 200 CIM sodium arsenate,
which was also the case for P. vittat. The bovine liver reached its maximum CAT
velocity with the addition 100 C1M sodium arsenate.
oa I-*- P. vittata
58 ..-P. ensiforinis
0 50 100 150 200 250 300
Arsenite concentration (CLM)
Figure 4-10. Effect of arsenite on GR activity in P. vittata and P. ensifornzis. Values
represent means + std. dev. (n = 3).
Sodium arsenate (mM)
Figure 4-11. Effect of sodium ar senate on CAT activity in P. vittat 20% PEG protein
fraction. Values represent means f std. dev. (n = 3).
Sodium arsenate concentration (pLM)
Figure 4-12. Effect of sodium ar senate on CAT activity in P. ensiformis 20% PEG
protein fraction. Values represent means f std. dev. (n = 3).
C m 1000
Sodium arsenate (CLM)
Figure 4-13. Effect of sodium arsenate on CAT activity in bovine liver (CAT positive
control). Values represent means f std. dev. (n = 3).
- +- -Bovine liver
0 200 400 600
Sodium arsenate concentration
Figure 4-14. Comparison of the percent change in CAT activity in P. vittata P.
ensiformis and bovine liver (CAT positive control) upon exposure to arsenate.
Percent change is based on the control (no sodium arsenate) assays for each
protein. The control CAT activities for P. vittata, P. ensiformis and bovine
liver were 16, 134 and 3271 Clmoles min' mg protein- Values represent
means f std. dev. (n = 3).
Antioxidant enzymes play an important role in plant responses to environmental
stress, such as exposure to arsenic. Changes in antioxidant enzyme activities upon stress
can give insight into a plant' s ability to tolerate stress and mediate its effects. Singh et al.
(2005, unpublished data) found that the GSH:GSSG ratio did not change significantly in
P. vittata exposed to arsenic. One reason for this was thought to be the induction by or
the efficiency of GR. However, GR activity was not induced in P. vittata upon arsenic
exposure. These results indicate that while GSH is an important antioxidant, the
recycling of it from GSSG to GSH by GR does not play an important role in P. vittat' s
ability to maintain the GSH: GSSG ratio. The same study indicated that GSH
concentrations also increase in P. vittat exposed to arsenic (Singh et al., 2005,
unpublished data). Therefore, it is possible that one or both of the GSH synthesizing
enzymes (gamma-glutamyl cysteinyl synthetase and glutathione synthetase) may be
induced by arsenic, causing an increase in GSH concentration and maintaining the
It was originally thought that although GR may not be induced in P. vittata that it
may be more efficient (in terms of Km) compared to a non-arsenic hyperaccumulator, P.
ensiformis. However, kinetics studies produced Km constants similar in both ferns.
These Km values were also comparable to those previously reported for GR in other
plants (Hausladen and Alscher, 1994; Griffith et al., 2001). The direct plots for NADPH
(Fig. 4-6 A and 4-7 A) indicated that higher concentrations of the substrate inhibits GR
activity. This was more evident in P. vittat.
It is interesting that arsenic did not inhibit or activate GR activity in P. vittat or P.
ensiformis. Because arsenite can readily bind to compounds with thiol groups, such as
GSH, it was thought that its presence could possibly impact GR activity by altering the
GSH:GSSG ratio. Although arsenite did inhibit GR activity in both species, it did so at a
concentration (1 mM) that was at least an order of magnitude greater than the substrates,
GSSG and NADPH. A lack of inhibition of GR activity by arsenic suggests that GR may
be compartmentalized away from arsenic, or that arsenic does not bind to GR and cause it
to be inhibited. The results of the induction, kinetics and inhibition studies suggest that
GR does not play an active role in the ability of P. vittat to hyperaccumulate arsenic.
Unlike GR, CAT was induced 1.5 times in the fronds of P. vittat plants exposed to
arsenic. Catalase activities were also found to be induced in the fronds P. vittata and P.
ensiformis in a study conducted by Srivastava et al. (2005). Catalase is the enzyme
responsible for the degradation of the ROS, H202, to water and oxygen. Pteris vittat has
been shown to contain mostly arsenite in its fronds (Chapter 3). This is in spite of the
results shown in Chapter 3 that most of the arsenic, when supplied as inorganic arsenic,
taken up by the fern is transported as arsenate. It is thought that the reduction of arsenate
to arsenite in plants increases the concentration of ROS in plants (Hartley-Whitaker et al.,
2001a; Meharg and Hartley-Whitaker, 2002). Therefore, it is possible that the induction
of CAT activity may be a result of H202 prOduced, directly or indirectly, from arsenate
reduction in the fronds. Superoxide dismutase (SOD) activity was also found to be
induced in P. vittata fronds (Srivastava et al., 2005). The SOD enzyme dismutates
superoxide anions, and it produces H202 in prOcess (Fridovich, 1978; Fridovich, 1986;
Fridovich, 1995). Therefore, the increase in CAT activity may also be affected by the
induction of SOD.
Catalases have been known to be extremely efficient in the degradation of H202.
The determination of the Km constants from the CAT kinetics assays did not indicate that
CAT enzymes in P. vittata were more efficient than P. ensiformis. The 7 to 8-fold
difference in the Vmax values for both species is interesting. This difference could be due
to the presence of an inhibitor in the P. vittat extract. It may also be simply due to
differences between the two plants. Switala and Loewen (2002) found that the observed
Vmax values of CAT in some bacteria varied 2 to 10 times for species within the same
genus. The differences are thought to be a result of inactivation of smaller-subunit
catalases by H202 damage. Small-subunit catalases reach their maximum velocity around
200 mM, which was similarly found in P. vittata. However, larger sub-unit catalases
reached a maximum velocity around 1 M. Pteris ensiformis CAT activity still appeared
to be fairly linear, but with some leveling off at a velocity of 220 mM H202.
Determinations of CAT activity with H202 COncentrations much greater than 200 mM are
not possible using the spectrophotometric method, as effervescence by H202 does not
allow for linear decrease in absorbance. It is presently unclear why P. vittat appears to
be more sensitive to damage caused by H202. It would be expected that the arsenic
hyperaccumulating fern would be less sensitive to such damage. However, because the
proteins were not purified, the accuracy of these Vmax values are rather uncertain and
The study by Switala and Loewen (2002) also concluded that the traditional
Michaelis-Menten kinetics terms, Km and Vmax, cannot be directly used for catalases.
Catalases do not follow Michaelis-Menten kinetics over the H202 COncentration range
because of the two-step CAT reaction. Therefore, the kinetic parameters should be
considered to be theoretical. This is especially true for concentrations greater than 200
mM. The same study did find a better correlation for lower substrate concentrations,
such as those used in this experiment.
Catalase activation by arsenate has not yet been reported in plants. However,
sodium arsenate appears to activate CAT activity in P. vittat at two concentrations, 10
and 200 mM (Fig. 4-14). The percent activation at a concentration of 200 mM sodium
arsenate was approximately 1.8 times greater than the activity found at 10 mM sodium
arsenate. The increase in CAT activity observed at the two concentrations may be a
result of the activities of two different CAT isozymes. Different CAT isozymes have
been shown to respond differently to the same conditions or stresses (i.e., Horvath et al.,
2002). As previously mentioned, almost all of the arsenate taken up by this fern is
reduced to arsenite in the fronds. This reduction likely produces ROS. It is possible that
the presence of arsenate activates some CAT isozymes in preparation for the pending
arsenate reduction and the subsequent production of H202. Similar activation patterns,
although not to the same extent, were observed in P. ensiformis and the CAT positive
control. Therefore, these results suggest that activation of CAT by arsenate may
constitute an important role in the ability of P. vittat to hyperaccumulate arsenic.
PHYTOREMEDIATION OF AN ARSENIC-CONTAMINATED SITE USING Pteris
Remediation of contaminated soils has traditionally focused on engineering-related
methods (Cunningham et al., 1997). Many of these methods, such as excavation, can be
expensive, while containment remediation techniques, such as capping, do not actually
remove the contaminants) from the soil. Recently, phytoextraction has emerged as a
potential in situ remediation alternative to these traditional remediation methods.
Phytoextraction is the use of plants to remove pollutants from the soil and/or water
matrices (Raskin and Ensley, 2000; Lasat, 2002; McGrath et al., 2002). Commonly,
hyperaccumulating plants are employed for phytoextraction purposes. By definition, the
aboveground dry matter of hyperaccumulators is comprised of greater than 0. 1% of the
element of interest. Ideally, a hyperaccumulator used for phytoextraction should have the
following characteristics: high rates of accumulation and translocation, fast growth and a
high production of biomass (Wantabe, 1997).
There is evidence that phytoremediation has a promising future role in soil and
water remediation. As such, interest in phytoremediation as a viable remediation
technology has significantly increased over recent years. Phytoremediation is still in its
infancy, but it is being used in some in-field remediation. Currently there are no sites that
have been completely remediated using phytoremediation (Schoor, 2002; USEPA,
Therefore, this raises several issues. One of which is cost. It is widely claimed that
phytoremediation is a much more economical remediation technology compared to most
other remediation techniques. There are estimated costs for the use of phytoremediation.
Using these figures, the costs do appear to be significantly lower than those for
conventional techniques. However, the costs can vary widely depending on the site
factors and the plants) being used to remediate the site. Another issue is time. The
amount of time needed to fully remediate a site is, again, very dependent on the plant and
site characteristics (Schoor, 2002)
Pteris vittat was the first arsenic-hyperaccumulating plant to be identified (Komar
et al., 1998; Komar, 1999; Ma et al., 2001). It is a relatively fast-growing perennial plant
that prefers alkaline soil. Most of the arsenic that is taken up by the fern is translocated
and accumulated in its aboveground biomass. It was shown to have relatively high
production of root and frond biomass. Further, P. vittat was found to have high
bioconcentration factor (BF) and translocation factor (TF) of arsenic, indicating its ability
to not only take up high amounts of arsenic, but also to translocate much of the arsenic to
its fronds (Tu et al., 2002), which can subsequently be harvested and taken off-site for
disposal. In a greenhouse study by Tu et al. (2002), 26% of the initial soil arsenic was
depleted using P. vittat after 20 wk of growth.
Because of its fast growth, relatively large biomass production and ability to
hyperaccumulate and translocate arsenic, P. vittata does exhibit the potential for use in
phytoextraction of arsenic-contaminated soils. One study focused on using P. vittat and
Indian mustard (Bra~ssica juncea) to phytoremediate a soil contaminated with arsenic and
lead (Pb) (Salido, et al., 2003). This study concluded that eight years would be needed to
decrease the acid-extractable soil arsenic concentration from an average of 82 to 40 mg
kg l. However, additional studies and field-related data are needed before this fern can be
used effectively for phytoextraction.
The main objectives of this field study were: 1) to determine the ability and
efficiency of P. vittat in accumulating arsenic from an arsenic-contaminated site; 2) to
determine the ability of P. vittata in decreasing total arsenic concentrations in the arsenic-
contaminated soils; and 3). To determine the most appropriate harvesting practices in
order to obtain maximum arsenic removal from the soil.
Materials and Methods
The field site, located in North central Florida, was previously used to pressure
treat lumber with CCA from 1951-1962. The lumber was pressure treated in a cylinder
using a CCA-solution containing arsenic pentoxide, copper sulfate and either sodium or
potassium chromate (Woodward-Clyde, 1992). Van Groenou, et al. (1951) found that
this solution commonly had a composition of 11% As, 33% Cu and 56% Cr. The past
pressure-treatment of lumber has lead to the current contamination found at this site.
The soil at the site is an Arrendondo-urban complex. The taxonomic classification
is a loamy, siliceous, hyperthermic Grossarenic Paleudult. Previous analysis found the
soil to have a pH of 7.5 and organic matter content of 0.5 to 0.8%. The particle size
distribution of the soil at this site was 88% sand, 8% silt and 4 % clay-sized particles
Planting and Plot Maintenance
In September 2000, a 30.3 mZ plot was prepared at the site. The plot was hand-
weeded, and non-porous black plastic mulch was placed on the experimental area. No
tilling was performed prior to transplanting P. vittat into hand-excavated holes (10.2 cm
wide by 10.2 cm deep). The planting density was 0.09 mZ per fern, for a total of 324
ferns. At the time of planting each fern was supplied with 13 g of STA-GREEN@ time-
released fertilizer brand (12-4-8). Due to late planting, high mortality over the winter
from frost and cold injury occurred, resulting in 314 ferns being replaced in April 2001.
The plot was hand-weeded approximately every two weeks as needed, and it was
watered daily with spray irrigation. No additional fertilizers or soil amendments were
added during the 2001 or 2002 growing seasons. During January and February 2002, the
ferns were covered with black plastic to prevent frost injury. However, due to some frost
injury and lack of water, 1 11 ferns died in 2001. They were replaced in April 2002 with
ferns of similar size. This plot was discontinued in October 2002 due to expansion of the
current business at the site.
In October 2002 plot I was essentially paved over. Therefore, another
experimental plot was initiated in a different area at the same site. Both the size (30.3
m2) and plant spacing (0.09 m2) Of plot 2 were identical to plot 1. Black plastic mulch
was placed on the experimental area, and plants were directly transplanted from plot 1 to
plot 2. No fertilizer was added at the time of planting. However, fertilizer (15-5-15) was
applied each year at a rate of 100 lb N yr in two split applications. The plot was hand-
weeded approximately every two weeks, or as needed, and it was watered every other day
with spray irrigation. Ferns were covered during the winter seasons using shade cloths.
Approximately 6 ferns died each winter and were replaced the following April.
No plant harvests were performed during the 2000 growing season. However, four
harvests were performed in 2001. Senescing fronds were removed at ground level by
hand in August, September and October 2001. Frond samples were taken from each
plant, and were grouped according to a pre-established grid of the site (36 total samples,
9 plants per sample). Fronds that were dead (brown and dry) and were close to
senescence (little green-colored tissue or most of the frond area was yellow, brown with
greatly mottled color) were removed from each plant at ground level by hand. In
December 2001, all plants were harvested, totaling 324 samples. With the exception of
fiddleheads and one to two live fronds, all fronds were removed from the ferns at ground
level. These exceptions were made to help facilitate survival of the ferns during the
winter season. Figure 5-1 A and B are photographs of the site during the 2001 season.
The ferns were harvested differently in 2002 to determine if harvesting frequency
and/or method affected the amount of arsenic removed from the site. Three harvesting
treatments were planned: senescing fronds harvested once a month (DDI); all fronds
harvested once a year (Alx); and all fronds harvested twice a year (A2x). However, due
to the termination of the plot in October 2002, the harvest treatments were not fully
implemented. As a result, senescing fronds in 1/3 of the plot were harvested in August,
September and October, and all fronds in 1/3 of the plot were harvested in August. They
were then extrapolated to the whole plot. Because of the very slow recovery of the ferns
from the winter season and the replacement of the dead ferns, many of the ferns were not
of adequate size to harvest until August.
The experimental design was a randomized complete block. The plot was divided
into 3 blocks. There were 6 subplots per block, for a total of 18 subplots. Subplots
contained 18 ferns. Each of the three harvest treatments was randomly assigned to two
subplots in each block; therefore, each harvest treatment was replicated six times in the
plot. Figure 5-1 C and D show photographs of plot 1 during the 2002 season.
Figure 5-1i. Photographs of P. vittat growing in the first experimental plot (200 1 to
2002). (A) Photograph of plot 1 in April 2001. Ferns were recovering from
the winter season. (B) Photograph of plot 1 in August 2001. (C) Photograph
of plot 1 in August 2002 prior to harvest, and (D) immediately following
harvest of the DDI1 and A2x treatments.
No plant harvests were performed in 2002. In 2003, three plant harvests were
made. All ferns were harvested to a height of 15 cm in July, September and November.
Harvesting treatments were implemented in 2004 to evaluate the effects of harvesting
frequency. The femns were harvested to a height of 15 cm one (lx), two (2x) or four (4x)
times that year. Femn borders were also in place around the harvesting treatments. The
experimental design was a randomized complete block with four replications. Each
replicate contained 20 P. vittata plants, for a total of 80 femns per treatment. The borders
consisted of a total of 84 ferns. Ferns were harvested in May (2x, 4x and borders), July,
(4x and borders), September (4x and borders) and November (lx, 2x, 4x and borders).
-~~p~a ~ EWECB~U;PF C C mm llll =irZgl~ e9 D
Figure 5-2. Photographs of P. vittat growing in the second experimental plot (2003 to
2004). (A) Photograph of plot 2 taken in April 2003. (B) Photograph of plot 2
taken in June 2003. (C) Photograph taken in 2004 after harvest in July 2004.
(D) Close up of femns showing that ferns were taller than 3 ft.
Determination of Frond Biomass and Arsenic Concentrations
For all harvests, frond samples were placed into a 600 C oven for approximately 48
h. Soil particles were removed from the dried fern samples, as necessary. The samples
were then weighed for dry biomass. Dried samples were ground through a 1 mm mesh
Wiley Mill screen. The ground samples (0.25 g) were subjected to HNO3/H202 digestion
(USEPA Method 3051) on a hot block (Environmental Express, Ventura, CA). The
digested plant samples were analyzed for total arsenic concentration using graphite
furnace atomic absorption spectroscopy (GFAAS, Perkin Elmer SIMMA 6000, Perkin-
Elmer Corp., Norwalk, CT).
In 2001 and 2002, all frond samples from the August to October dead and dying
harvests were analyzed for total arsenic concentration. However, due to the large number
of samples harvested in December 2001, only 72 of the 324 fern samples harvested were
analyzed for total arsenic concentration. The samples that were chosen for analyses were
those of the median dry mass weight. All frond samples collected in 2003 and 2004 were
analyzed for total arsenic concentration.
Soil samples were extracted in September 2000, December 2001 and October 2002.
In September 2000 and December 2001, 36 surface (0-15 cm) soil samples were
systematically taken (1 sample per 0.09 m2). In addition to the surface samples, 9 soil
profile samples were extracted (15-30 cm and 30-60 cm). Three sets of profile soil
samples were taken for every 12 surface samples. Due to extreme difficulty in extracting
the soil samples, only 10 random surface samples and 5 random profile samples at each
depth were taken in October 2002.
Soil samples were extracted in December 2002, 2003 and 2004. Within the plot
area, 49 surface (0-15 cm) and profile (15-30 cm and 30-60 cm) soil samples were taken
systematically. In addition, 12 sets of soil samples were extracted from outside the plot
area to compare the difference in soil arsenic concentrations inside and outside the plot,
as affected by P. vittat (Fig. 5-3).
Determination of Total Soil Arsenic
All soil samples were air dried and sieved to pass through a 2 mm mesh screen.
The sieved soil samples (0.5 g) were subj ected to hot block (Environmental Express,
Ventura, CA) digestion using USEPA Method 3051 (HNO3 H202) for arsenic analysis.
The digested soil samples were then analyzed for total arsenic concentration using
Sequential Soil Arsenic Fractionation
A sequential soil arsenic fractionation was performed on all soil samples extracted
from plot 1 (2000 to 2002) using the method developed by Wenzel, et al. (2001). This
sequential extraction method, which represents a functional fractionation, contains five
arsenic fractions (decreasing in availability): non-specifically bound, specifically bound,
amorphous hydrous oxide-bound, crystalline hydrous oxide-bound and residual.
Using approximately 1.0 g of air-dried soil from each soil sample, arsenic was
extracted Wenzel, et al (2001). The non-specifically bound fraction was extracted by
shaking for 4 h at 20oC with 25 ml of 0.05 M ammonium sulfate. The specifically bound
fraction was extracted by 16 h of shaking at 20oC upon the addition of 25 ml of 0.05 M
ammonium phosphate. Twenty-five ml of 0.2 M ammonium oxalate buffer (pH 3.25)
was then added. Samples were shaken in the dark for 4 h at 20oC. The samples were
then washed with 12.5 ml of 0.2 M ammonium oxalate buffer (pH 3.25) by shaking for
10 min in the dark. This fraction was labeled as the amorphous hydrous oxide-bound
fraction. The crystalline hydrous oxide-bound fraction was extracted by mixing the
samples with 25 ml 0.2 M ammonium oxalate buffer and 0.1 M ascorbic acid (pH 3.25)
and placing them in a 96oC water bath for 30 min. Each sample was washed byl2.5 ml
of 0.2 M ammonium oxalate buffer (pH 3.25) by shaking for 10 min in the dark. The
residual fraction was extracted by acid digestion using USEPA Method 3051
(HNO3 H202). Samples from each fraction, with the exception of the residual fraction,
were centrifuged at 3500 rpm for 15 min and 20oC after each extraction and/or wash.
The supernatants were collected. The supernatants of each fraction were filtered through
Whatman 42 filter paper and analyzed for arsenic concentration using GFAAS.