Footage showed rebel fighters waving from the shell of a house bombed by U.S. warplanes in 1986, which Gaddafi had preserved as a monument to his own survival, and clambering onto a bronze sculpture of Gaddafi’s fist clutching an F-16 fighter jet, a work commissioned to commemorate the attack. 的翻译是：画面显示叛军于 1986 年，其中卡扎菲作为自己生存的纪念碑被保存了下来，从一家美国战机轰炸壳挥舞着与卡扎菲的拳头抓的 F-16 战斗机工作委托为纪念这次袭击的青铜雕塑到脚跟。 中文翻译英文意思，翻译英语
请在下面的文本框内输入文字，然后点击开始翻译按钮进行翻译，如果您看不到结果，请重新翻译！
Footage showed rebel fighters waving from the shell of a house bombed by U.S. warplanes in 1986, which Gaddafi had preserved as a monument to his own survival, and clambering onto a bronze sculpture of Gaddafi’s fist clutching an F-16 fighter jet, a work commissioned to commemorate the attack.
选择语言：从
罗马尼亚语
罗马尼亚语
录像显示，反叛战士挥舞着从美国战机轰炸在1986年一个房子，这卡扎菲曾作为他自己生存的纪念碑保存的外壳，并到一个卡扎菲的拳头青铜抓着一个F - 16战斗机的喷气雕塑攀登，一个工作委托纪念攻击。
画面显示叛军于 1986 年，其中卡扎菲作为自己生存的纪念碑被保存了下来，从一家美国战机轰炸壳挥舞着与卡扎菲的拳头抓的 F-16 战斗机工作委托为纪念这次袭击的青铜雕塑到脚跟。
画面显示叛军于 1986 年，其中卡扎菲作为自己生存的纪念碑被保存了下来，从一家美国战机轰炸壳挥舞着与卡扎菲的拳头抓的 F-16 战斗机工作委托为纪念这次袭击的青铜雕塑到脚跟。
镜头显示叛军战士挥舞着从船壳的美国众议院的轰炸 战机在1986年,卡扎菲已保持为一个纪念碑,他本身的生存,边歪斜上青铜雕塑的卡扎菲的拳手抱一架F-16战机,委托一工作为纪念的攻击。
英尺长度显示了挥动从美国轰炸的房子的壳的反叛战斗机。 1986年战机，卡扎菲保存了作为纪念碑到他自己的生存和攀登抓住F-16喷气式歼击机，工作的卡扎菲的拳头一个铜雕塑被委任纪念攻击。
相关内容&a汤姆的玩具比彼得的多 Tom's toy compared to Peter many & a李明是谁的朋友 Whose friend is Li Ming & aI waited there until you called me. 我等待了那里，直到您告诉了我。 & a不过，我会努力跟上班上的同学、不让自己成为最后一名、不会让老师、父母失望 I can with go to work diligently on schoolmate, do not enable oneself to become last, cannot let teacher, the parents be disappointed & a适用于建筑物，雕塑，码头，广告牌，广场，停车场及绿化等照明场所 Is suitable in the building, the sculpture, the wharf, the billboard, the square, illumination places and so on parking lot and afforestation & aYou can't change the past, but you can ruin the present by worrying about the future 您不可能改变过去，但是您能通过担心破坏礼物未来 & a我是我爱我家张自忠店的陈新涛，你什么时候过来 I am I love my family Zhang Zizhong shop Chen Xin Tao, when do you come & aindustry's average profit potential 业界平均值赢利潜力 & a空缺的 Vacancy & a很不幸 Very unfortunate & a亲爱的明 Dear bright & aThe gently sigh was regretting sorry 柔和地叹气是后悔抱歉 & aSpectral Line Half-Width 鬼线半宽度 & abuilding17 building17 & akoncked off koncked & a你通常用电脑来做什么 ??? ?? ????? ??? ??? ??? ?? ???? & aSometimes I miss you so much that I can hardly stand it. 有时我非常想念您我可以几乎不站立它。 & afacilitates 促进 & aThanksgiving is not only part of the United States 感恩作为美国的不仅部分 & a每天早晨我都要散步或者慢跑半个小时。 Every morning I all must take a walk or jog for half hour. & agreat changes have taken place there and my hometown has more and more beautiful.where have you been,Jane? 巨大变动发生了那里，并且我的故乡越来越有beautiful.where有您是，珍妮？ & awords fail me 词发生故障我 & a很少有趣的书 Very little interesting book & athe intimate(Ae ln)Full movie 亲密的(Ae ln)充分的电影 & awrite the new words in your notebook 写新的词在您的笔记本 & a散水坡 Apron slope & ainvalid command-line parameter: 手. 无效命令线参量： 手。 & aok tks 好tks & ai'm on my way i'm在我的途中 & a我叫刘丹，英文名字叫COffey，32岁，大专毕业，从事人力资源工作已有7年，其中4年在零售行业-Belle公司，3年在吉的堡少儿英语学校 My name am Liu Dan, English name to call the COffey,32 year old, the technical college to graduate, am engaged in the human resources work to have 7 years, 4 years in retail sales profession - Belle Corporation, 3 years in lucky fort children English school & aI never knew I could hurt like this~ 我未曾知道我可能伤害象this~ & a我爱你白脸 I love your white face & asad and failure 哀伤和失败 & a就算我的超人，也绝不会有justinbieber帅 Even if my superhuman, also cannot have Commander justinbieber & awhat you give up me for 什么您放弃我为 & a谁是我的超人？ Who is my superhuman? & ata em pt 您在pt & a她是一个可爱的女生 She is a lovable female student & aadminister oxygen at 6 to 12 L/minute via mask 执行氧气在6到12 L/minute通过面具 & atwo heads are better than one 二个头比一个好 & aI know love means never say sorry，I am me 我知道爱手段从未认为抱歉，我是我 & a无聊的游戏 Bored game & aApply daily to face and neck. 每日适用于面孔和脖子。 & ai have sent several email to : also including your visa. i物产被送数电子邮件对： 也包括您的签证。 & ashawn listen carefully your target is the smoking guy on the alley entrance. find a way to maim him without killing him 肖恩在胡同入口听仔细地您的目标是抽烟的人。 发现一个方式残废他，不用杀害他 & a我有在两个公司工作，都是从零做起来的，我有足够的信心把工厂做起来，目前缺少的就是订单； I have in two companies work, all is does from the zero, I have the enough confidence to do the factory, at present lac & aAnd our look at English democracy does not prevent us from appreciating the great American democracy. 并且我们的神色在英国民主不防止我们赞赏巨大美国民主。 & a我是野猫 依然 I am the stray cat still & aif the meal is over it is notpolite to leave forat least half an hour ，lest you seem to have come only for the meal 如果膳食是它是留给forat最少半小时的notpolite，唯恐您似乎为膳食仅来了 & aAviod the eye area. Aviod眼睛区域。 & aTripod mount 三脚架登上 & aFootage showed rebel fighters waving from the shell of a house bombed by U.S. warplanes in 1986, which Gaddafi had preserved as a monument to his own survival, and clambering onto a bronze sculpture of Gaddafi’s fist clutching an F-16 fighter jet, a work commissioned to commemorate the attack. 英尺长度显示了挥动从美国轰炸的房子的壳的反叛战斗机。 1986年战机，卡扎菲保存了作为纪念碑到他自己的生存和攀登抓住F-16喷气式歼击机，工作的卡扎菲的拳头一个铜雕塑被委任纪念攻击。 &versión On-line ISSN
J. Soil Sci. Plant Nutr. v.10 n.3 Temuco jul. 2010
http://dx.doi.org/10.-00005
J. Soil. Sci.
Plant Nutr. 10 (3): 268 - 292 (2010)
AND BIOAVAILABILITY OF HEAVY METALS AND METALLOIDS IN SOIL ENVIRONMENTS
A. Violante1*,
V. Cozzolino1, L. Perelomov2, A.G. Caporale1,
and M. Pigna1.
1Dipartimento
di Scienze del Suolo, della Pianta e dell'Ambiente, Universitá di Napoli Federico
II, Portici
(Napoli), Italy. * Corresponding author:
2Pula State University, Lenin Avenue, 92, Pula, 300600,
In soil environments,
sorption/desorption reactions as well as chemical complexation with inorganic
and organic ligands and redox reactions, both biotic and abiotic, are of great
importance in controlling their bioavailability, leaching and toxicity. These
reactions are affected by many factors such as pH, nature of the sorbents, presence
and concentration of organic and inorganic ligands, including humic and fulvic
acid, root exudates, microbial metabolites and nutrients. In this review, we
highlight the impact of physical, chemical, and biological interfacial interactions
on bioavailability and mobility of metals and metalloids in soil. Special attention
is devoted to: i) the sorption/desorption processes of metals and metalloids
on/from soil
ii) their precipitation and reduction-oxidation
reactions in solution and onto surface iii) their chemical
speciation, fractionation and bioavailability.
Keywords: bioavailability,
heavy metals, metalloid, adsorption, desorption, chemical speciation
INTRODUCTION
Characterizing
the factors affecting bioavailability, leaching and toxicity of metals and metalloids
in soil is of paramount importance. Metals and metalloids are significant natural
components of all soils where their presence in the mineral fraction comprises
a store of potentially-mobile metal species as important components of clays,
minerals and iron and manganese oxides that, in turn, have a dramatic influence
on soil geochemistry (Gadd, 2008). Metals are also present in the organic fraction,
frequently as bound forms, with some metal recycling occurring as a result of
organic matter decomposition. The aqueous phase provides a mobile medium for
chemical reactions, metal transfer and circulation through the soil, to organisms,
and also to the aquatic environment.
Heavy metals and
metalloids can be involved in a series of complex chemical and biological interactions.
The most important factors which affect their mobility are pH, sorbent nature,
presence and concentration of organic and inorganic ligands, including humic
and fulvic acids, root exudates and nutrients. Furthermore, redox reactions,
both biotic and abiotic, are of great importance in controlling the oxidation
state and thus, the mobility and the toxicity of many elements, such as Cr,
Se, Co, Pb, As, Ni and Cu. Redox reactions can mobilize or immobilize metals,
depending on the particular metal species and microenvironments.
REDOX REACTIONS
AND METAL TRANSFORMATIONS
The reductive transformation
of some heavy metals may proceed chemically, for example Cu(II) reduction to
Cu(I) by Fe2+ or H2S and reduction of Cu(II), Ag(II),
and Hg(II) to elemental forms by Fe(II)-bearing green-rust (Borch et al,
2010; and references therein) . Microorganisms may directly reduce many
highly toxic metals (e.g., Cr, Hg, U) via detoxification pathways. Microbial
reduction of certain metals to a lower redox state may result in reduced mobility
and toxicity. Such processes may accompany other metal precipitation mechanisms.
Aerobic and anaerobic reduction of Cr(VI) to Cr(III) is widespread in microorganisms
(Gadd, 2008).
Sulfate reduction
in contaminated soils may mobilize Cu, Pb, and Cd through the formation of Cu-rich
sulphide colloids. Sulphate-reducing bacteria (SRB) oxidise organic compounds
or hydrogen coupled with the reduction of sulphate, producing sulphide. The
solubility product of most heavy metal sulphides is very low, so that even a
moderate output of sulphide can remove metals from solution. As reviewed by
Borch et al. (2010) bio-geochemical oxidation processes are driven by
02 entering anoxic systems. Certain oxidation processes such as Fe(II)
oxidation by 02 at neutral to alkaline pH proceed rapidly abiotically,
but many slower processes are facilitated by chemotrophic microorganisms.
Precipitation of
Fe-, Mn-, and Al-(hydr)oxides may efficiently capture dissolved trace metals
(Violante et al, 2008). If the precipitates formed during oxidative precipitation
are nanoparticulate and colloidal, they may greatly enhance the
mobility of associated trace metals in aquatic and subsurface environments.
Biogeochemical
redox processes strongly influence also metalloids mobility such as arsenic
(As) and antimony (Sb). Whereas environmental As chemistry has received great
attention during the last two decades due to the worldwide health impacts of
As contaminated drinking water and soils, Sb may locally represent an important
contaminant, for example in the vicinity of copper and lead ore smelters. Oxidation
of Sb(0) to Sb(III) or Sb(V), as well as sorption to Fe (hydr)oxides, controls
Sb toxicity and mobility at these sites.
The As mobility,
bioavailability, toxicity, and environmental fate are controlled by biogeochemical
transformations that either form or destroy As-bearing carrier phases, or modify
the redox state and chemical As speciation (see
in Borch et al, 2010).
Arsenic in natural
waters and soils is intimately related to the presence of iron (hydr)oxides
being arsenate and arsenite strongly sorbed onto the surfaces of Fe-oxides,
forming inner-sphere complexes. The simultaneous presence of high dissolved
As and Fe(II) concentrations in anoxic groundwater has led to the conclusion
that reductive dissolution of As rich Fe(III) (hydr)oxides mobilizes geogenic
in Borch et al, 2010). More recently,
it has been demonstrated that microbial sulphide formation in As-ferrihydrite
systems leads to transformation and dissolution of the As-bearing ferrihydrite.
Direct reduction of Fe(III) by microorganisms can lead to As sequestration by
sorption of As onto secondary Fe minerals (see
in Borch et al, 2010).
Clays and oxides
often demonstrate the ability to catalyze electron transfer reactions. Abiotic
redox processes occur on the surfaces of humic substances Fe(III) and Mn-oxides
as depicted in . Iron and manganese oxides and layer
silicates with structural Fe(III) are the most active in this regard. Redox
reactions also control the transformation and Fe- reactivity and Mn-oxides in
soils, which are the major sinks for heavy metals and metalloids. The Mn oxides
efficiency as &electron pump& for a wide range of redox reactions
is unique among common soil minerals.
Redox reactive
metals often do have different degrees of toxicity depending on the specific
metal oxidation state. For example, enrómate is toxic to plants, animals and
humans and is a suspected carcinogen, whereas Cr(III) is not toxic to plants
and is necessary in animal nutrition, so that reactions that reduce Cr(VI) to
Cr(III) are of great importance. Furthermore, Cr(VI) is mobile in soils and
readily available. Organic material, sulfides, and ferrous species appear to
be the dominant reductants. Very stable Cr(III)-organic complexes form when
Cr(VI) is reduced by soil organic matter (Fendorf, 1995).
SORPTION / DESORPTION
PROCESSES OF HEAVY METALS AND
METALLOIDS
The soil components
responsible for trace element sorption include, soil humic substances, phyllosilicates,
carbonates and variable charge minerals (constituents such as Fe, Al, Mn and
Ti oxides, short-range ordered aluminosilicates as well as phyllosilicates coated
by OH-A1 or OH-Fe species whose charge varies with the pH of the soil solution).
Evidence on the
sorption of heavy metals on microorganisms has been reported. Soil components
differ greatly in their sorption capacities, their cationand
anion exchange capacities, and the binding energies of their sorption sites.
Sorption of
cationic metal species
Heavy metals show
typical ion exchange behavior on layer silicate clays with permanent charge,
demonstrating essentially the same affinity for exchange sites on the clays
as do alkaline earth metals having the same charge and similar ionic radius.
Surface bonding is electrostatic, dependent only on the charge and hydration
properties of the cation. Trace elements in cationic form are probably not dominantly
sorbed on phyllosilicates because they are always vastly outnumbered by other
cations with which they compete (e.g. Ca). They may be strongly sorbed only
on the edges of the phyllosilicates. However, clay minerals have also an important
role as carriers of associated oxides and humic substances forming organo-mineral
complexes, which present peculiar sorption capacities different from those of
each single soil constituent.
There are two general
surface complexes and are described by the configuration geometry of the adsórbate
at the adsorbent surface. These include inner- and outer-sphere surface
complexes and are defined by the presence, or absence, of the hydration
sphere of the adsórbate molecule upon interaction. When at least one water molecule
of the hydration sphere is retained upon sorption, the surface complex is referred
to as outer-sphere. When the ion is bound directly to the adsorbent without
the presence of the hydration sphere, an inner-sphere complex is formed (Sposito,
1984; Sparks, 2003; Borda and Sparks, 2008).
Humic matter and
metal oxides are much more effective sorbents of heavy metals in cationic form
than even the most efficient sorbent among phyllosilicates, indicating that
specific sorption and other complexation processes are the dominant binding
mechanisms.
Except for some
noncrystalline minerals that have very high specific surface charge density
with highly reactive sites, organic matter appears to have the greatest capacity
for sorption of trace elements in cationic form.
Humic substances
contain a large number of complexing sites, hence they behave as a natural &multiligand&
complexing system (Buffle, 1988). The high selectivity degree of soil organic
matter for most trace elements in cationic form indicates that they form inner-sphere
complexes with the functional groups, often forming an internal five- or six-member
ring on structures (Senesi, 1992; Senesi and Loffredo, 1998; Huang and Germida,
2002; Sparks, 2003).
Complexation reactions
have the following effects: i) metal ions are prevented fro
ii) complexing agents can act as carriers for trace eleme
iii) their toxicity is often reduced by complexation. The stability constant
(K) of trace element complexes with humic acids increases with increasing pH
and decreasing ionic strength.
Unlike phyllosilicates,
the variable charge minerals (crystalline and short range ordered Fe-, Al-,
Mn-oxides, allophanes, imogolite) strongly retain trace heavy metals for their
dependency on pH. On these materials a hydroxylated or hydrated surface, positive
or negative charge is developed by sorption or desorption of H+ or
OH- ions (). The pH at which the
net variable charge on the surfaces of these components is zero is called the
point of zero charge (PZC). The reported PZC of Fe-oxides range from pH 7.0
to 9.5, whereas that of Al-oxides ranges from pH 8.0 to 9.2. (Hsu, 1989; Cornell
and Schwertman, 1996; Violante et al, 2005).
Variable charge
minerals selectively sorb polyvalent cations, even when their surfaces are positively
charged (solution pH values lower than the point of zero charge [PZC] of the
sorbent). Most transition cations (Pb, Cu, Cr, Ni, Co, Zn, Al, Fe, Mn) are often
sorbed more strongly than alkaline earth cations. Spectroscopic techniques such
as electron spin resonance (ESR) and Extended X-ray Absorption Fine Structure
Spectroscopy (EXAFS) have been used for the identification of metal complexes
at the surfaces of Al, Fe or Mn oxides, silicate clays and soil organic matter.
Sorption of metal
cations increases by increasing pH. Sorption, which rises from 0 to 100% of
the added amount over a narrow region of 1-2 pH units, is termed as &sorption
edge&. The pH at which 50% of the total sorption has occurred is called
pH50. The lower the pH50 of a trace element for a sorbent,
the stronger is the element-surface complex (Kinniburgh and Jackson, ;
Sparks, 2003;
Violante et al.,
). Experiments with various synthetic Fe, Al and Mn oxides showed
that the affinity of trace elements for Mn oxide was usually much greater than
that for Fe or Al oxides. However, the nature, crystallinity size of the crystals,
surface charge of metal oxides and mixed metal oxides (e.g., Fe-Al oxides) also
play an important role in the sorption selectivity of trace elements in cationic
form (Kinniburg and Jackson, 1976; McBride, 1982; Violante et al., 2003; Sparks,
2003; Violante et al., 2008). Evidence on the sorption of trace elements on
microorganisms have been reported. Biosorption can be defined as the microbial
uptake of soluble and insoluble organic and inorganic metal species by physico-chemical
mechanisms such as sorption and, in living cells, metabolic activity may influence
this process because of changes in pH, Eh, organic and inorganic
nutrients and metabolite excretion. Biosorption can also be a prelude for the
formation of stable minerals. Cationic metal species can also be accumulated
within cells via membrane transport systems of varying affinity and specificity
(Gadd, 2008). Free-living bacteria and their extra-cellular macromolecular products
(e.g. fibrils)
can accumulate trace elements and may have mineral coatings with bound metals
on their surfaces (Beveridge, 1989a and 1989b; Jackson and Leppard, 2002 and
references there in). All microorganisms contain biopolymers such as proteins,
nucleic acids, and polysaccharides which provide reactive sites for binding
metal ions. Cell surfaces of all bacteria are negatively charged containing
different types of negatively charged functional groups, such as carboxyl, hydroxyl
and phosphoryl that can adsorb metal cations, and retain them by mineral nucleation.
Intact bacterial
cells, live or dead, and their products are also highly efficient in accumulating
both soluble and paniculate metals forms. Therefore, bacteria play an important
role in the speciation, fate and transport of metals, metalloids and radionuclides
in soil and associated environments. Lopez et al. (2000) reported the
following affinity order for the surfaces of microorganisms: Ni & Hg &
U & As & Cu & Cd & Co & Cr & Pb.
Biosorption is
a fast and reversible process for removing toxic metal ions from solution. Many
environmental factors influence the chemical reactivity of the binding sites
on bacterial cell surfaces and the subsequent biosorption of metals. These factors
include pH, ionic strength, temperature, and the presence of other metals and
organic compounds. The ability of bacteria to accumulate toxic metals also varies
with cell age. Many factors such as pH, surface properties of the sorbents,
number of sites available for sorption, nature and charge of Me-L species in
solution influence trace element sorption onto soil components in the presence
of inorganic and biological ligands. In rhizospheric soils, the amount of natural
organic compounds is much higher than in bulk soil at the soil-root interface.
In fact, rhizosphere C flow has been estimated to account for a large fraction
of plant primary production. Up to 20% of the C assimilated through photosynthesis
can be released from roots. In the rhizosphere, root exudates comprise both
high and low-molar-mass substances released by the roots. The most important
high-molecular-mass compounds are mucilage, polysaccharides and ectoenzymes,
whereas the main constituents of the low molecular mass root exudates are carbohydrates,
organic acids, amino acids, peptides and phenolics. Microorganisms produce a
number of extracellular metabolites that can complex metals in solution, including
polysaccharides, pigments, organic acids and siderophores. Citric and oxalic
acids are released by free-living and mycorrhizal fungi, lichens and plant roots.
Inorganic and organic
ligands (e.g., organic acids) which form strong complexes with ions of metals
usually prevent or reverse their association with negatively charged sorbents,
as phyllosilicates, by forming stable dissolved or dispersed negatively charged
complexes with the cations. In contrast, the presence& of& certain
(phyto)siderophores
produced by microorganisms and phytosiderophores exuded by plants may promote
the formation of positive complexes and, consequently, the sorption of trace
elements onto phyllosilicates (Violante et al, 2008; see their .3 and .4). The processes, which affect the sorption
of trace element cations onto variable charge minerals in the presence of complexing
agents, are particularly complex and are different from those onto phyllosilicates
(Violante et al., 2008). Recently, Perelomov et al. (2010) studied the influence
of selected low molecular mass organic ligands on Cu and Pb sorption onto a
shows Cu and Pb amounts (25 mmol added per kg) sorbed at pH 5.0 in the
presence of oxalic (OX), citric (CT) or glutamic (GL) acid. By increasing the
initial organic ligand/Me molar ratio (r) from 0 to 10, the sorption of both
metals on the iron oxide initially increased and then decreased or remained
constant, as referred to their sorption when added alone (r = 0). In the presence
of oxalic acid, the sorption of Cu on the surfaces of goethite increased until
2 and then decreased. Citric acid showed a similar effect on Cu sorption, but
fixed Cu amounts
were lower at CT/Cu molar ratio of 2 and 4. Glutamic acid had the lowest effect
on the sorption of this element onto goethite. Cupper sorption sharply increased
at r = 1 and then grew very slowly. However, at R & 6 a greater Cu amount
was sorbed in the presence of glutamic acid than in the presence of oxalic or
citric acid (Figure 3A). Lead showed a similar behavior, but its sorption was
lower than that of Cu on this sorbent (B).
Adsorption of Cu (A) and Pb (B) on goethite at pH 5 in the presence
of increasing concentrations of Oxalic (OX),citric
(CT) or glutamic acid (initial ligand/Cu or Pb molar ratio from 0 to 10;
25 mmol of Cu or Pb added per kg of goethite).
findings indicate that heavy metals complexed with organic ligands have a greater
affinity for the surfaces of a sorbent at certain ligand/metal molar ratio and
promote the formation of ternary complexes (McBride, 1989). The complexes affinity
is affected by the nature and concentration of the cation and the organic ligand
involved in the complexes, mineralogy and surface properties of the sorbent
and pH (Violante et al, 2008, and references therein). In the presence
of high concentrations of negative charged organic ligand, the sorption of heavy
metals decreases clearly because different processes, which prevent the metal
sorption, may occur. Large amounts of these organic ligands may either occupy
many sorption sites on the iron mineral, which are then not available for the
metals, or may favor complexation in solution and metal desorption relative
to ternary complex formation. In other words, large excess of chelating ligands
may shift the equilibrium in favor of soluble metal complexes and, consequently,
the ternary complex is destabilized.
Competitive
sorption of heavy metals in absence or presence of chelating ligands
Heavy metals compete
for sorption sites onto soil components (Violante et al., ,
2008). Experiments on the competitive Cu and Zn sorption were carried out on
a ferrihydrite, a humic acidlike sample (POL), and a Fe (OH)x-POL complex (Capasso
et al., 2004). On POL as well as on ferrihydrite, Cu was sorbed more selectively
than Zn. In the presence of equimolar Zn concentrations, the sorbed Cu amounts
were only slightly reduced when compared with the amounts sorbed in the absence
of Zn. Conversely, the Cu presence strongly reduced Zn sorption. A similar trend
was observed using
a Fe (OH)x-POL complex as sorbent, but the quantities of sorbed Cu and Zn on
this complex were much lower than those fixed on the humic acid-like sample.
Time of reaction,
surface coverage, sequence of addition of sorbates have also a great influence
on the competitive sorption between trace elements and organic and inorganic
ligands. Most competitive sorption studies have been carried out adding the
ions contemporaneously. However, it is more likely that the ions will come in
contact with a sorbent sequentially in natural environments, i.e., the solid
is exposed to one ion first, with the second ion coming in contact with a solid
at a later time. The sorption of metals and metalloids is strongly affected
by the order of addition of organic and inorganic ligands and trace elements
on the sorbents. It has been demonstrated that larger amounts of selected heavy
metals (Pb, Cu) were sorbed when chelating organic anions (oxalate, tartrate)
were added before or after the metals as referred to the systems where the heavy
metals were added as a mixture with the organic ligands or alone (Violante et
al. 2003).
Competition in
sorption between two or three heavy metals on soil components take into account
attention in the last years (Violante et al., 2008 and references there
in). Recently, Perelomov et al., 2010 studied the competition in sorption
between Pb and Cu in the absence or presence of oxalic acid. Figure 4 shows
the effect of increasing Pb concentrations (25 mmol added per kg) on Cu sorption
(initial Pb/Cu molar ratio [R] ranging from 0 to 10) at pH 5.0 either in the
absence or presence of oxalic acid. Oxalic acid (25 mmol kg-1) was
added as a mixture with the heavy metals (Cu+Pb+OX systems) or one hour before
Cu + Pb addition (OX//Cu + Pb systems).
Lead strongly inhibited
Cu sorption, but its inhibition was affected by the initial Pb/Cu molar ratio
and presence of oxalic acid. In the absence of oxalic acid (Pb + Cu systems),
the Pb inhibition in preventing Cu sorption increased from 11% at R = 1 to 55%
at R = 10 (numbers reported on the curves in Figure 4 indicate the inhibition
in percentage). The presence of the organic ligand enhanced Cu sorption when
added together, more than before the trace elements. In fact, the increase of
Cu sorption ranged from 34 to 43% in Cu+Pb+OX systems and from 18 to 25 in OX//Cu+Pb
systems. In the presence of oxalic acid, Pb was less efficient in preventing
Cu sorption than in the absence of the organic ligand, but it affected Cu fixation
more in OX//Cu + Pb (from 8 to 35 %) than Cu + Pb + OX systems (from 7 to 23%)
(not shown).
Desorption of
trace elements in cationic form
Desorption studies
of heavy metals in cationic form have shown biphasic reaction processes, a fast
reaction followed by a slow reaction (Sparks, 1990).
The presence of
inorganic and organic ligands has a significant impact on the desorption of
trace elements from soils or soil components. Krishnamurti et al. (1997)
demonstrated that LMMOLs have the ability to desorb Cd from soils, with malate,
fumarate, and succinate being the most effective ligands. Helal (2006) found
that small Pb portions less than 16% of the amounts previously sorbed onto Fe-
or Al-precipitates were replaced at pH 7.0 by LMMOLs (oxalate, citrate and taímate).
Lower amounts were desorbed from Fe-oxides. Desorption increased as sorption
density rose and pH decreased.
Another important
aspect that influences desorption of heavy metals in cationic form is the residence
time effect. Some researchers found that trace elements reacted with metal oxides
and pyrophyllite over longer times resulted in either irreversible or reversible
sorption mechanisms. Helal (2006) showed that as residence time sorbed Pb, in
particular at pH 7.0 it was nearly irreversible over a desorption period of
24 h. Violante et al. (2003) studied residence time effect on Zn sorption
onto ferrihydrite in the presence of copper. As copper has a greater affinity
than zinc for the surfaces of ferrihydrite, Cu was added from 1 to 336 hours
after Zn at a Zn/Cu molar ratio of 2. Zn sorption increased, especially when
Cu was added 6 to 336 hours after Zn. A possible explanation of these findings
is that trace elements initially sorbed on the surfaces of variable charge minerals
slowly form precipitates with time and/or penetrate into micropores of the sorbents.
Sorption of
trace elements in anionic form
Trace elements
which exist in anionic form are mainly sorbed at reactive sites of metal oxides
and allophanes and at the edges of phyllosilicates (Kampf et al. 2000; Violante
et al. 2008). Sorption of anions onto variable charge minerals and soils varies
with pH. With increasing pH, within a certain range, sorption decreases (due
to a decrease of positive charge of minerals) or increases to a maximum close
to pKa for anions of monoprotic conjugate acids and then decreases.
Anions may be sorbed
specifically or nonspecifically. Ligands which are specifically sorbed replace
OH- or OH2 groups from the surfaces of variable
charge minerals. These reactions are promoted at a low pH, which causes OH- groups
to accept protons, being OH2 group an easier ligand to displace than
OH& Specific sorption is also termed &inner-sphere sorption&
as reported above. (Violante et al., 2008).
Trace elements,
which form inner-sphere complexes, are molybdate, arsenate, arsenite
(only onto Fe-oxides) and selenite. Particularly, arsenate may form different
surface complexes on inorganic
soil components: monodentate, bidentate-mononuclear and bidentate-binuclear
complexes in different proportion depending on pH and surface coverage (Fendorf
et al. 1997; O'Reilly et al. 2001). Several studies have suggested
that arsenate is sorbed more than arsenite in a wide range of pH. However, literature
studies have found that arsenite is sorbed more than arsenate on ferrihydrite.
The sorption mechanisms of enrómate are unclear. Zachara et al. (1989) suggested
that chromate forms an outer-sphere complex on the surfaces of Fe and
Al oxides. However, spectroscopic studies have shown that chromate forms inner-sphere
complexes on goethite (Fendorf et al. 1997). This anion has a smaller
shared charge than arsenite and arsenate, creating a weaker bond on sorption
(McBride, 1994) and consequently, exhibits a steeper reduced sorption at near
neutral pH values than that of arsenate (Grossl et al. 1997). However,
where more than one type of surface species is present, XAS bulk will detect
only the primary (or average) type of surface product/species in the bulk sample
(i.e., sums over all geometric configurations of the target atom). Consequently,
while it may be concluded that the primary surface complex is inner-sphere,
this does not mean that outer-sphere complexation is not occurring. Recently,
through the use of X-ray scattering measurements to study metal(loid) binding
on single crystal surfaces, Catalano et al. (2008) showed that arsenate
surface complexation was bimodal, with sorption occurring simultaneously as
inner- and outer-sphere species.
Martin et al.
(2009) showed that arsenate and arsenite were differently sorbed onto different
minerals (Figure 5). Arsenate showed a higher affinity for minerals containing
iron or manganese than on minerals containing aluminum (Violante and Pigna,
2002). Pigna et al. (unpublished data) found that arsenite is sorbed
more than arsenate, particularly in alkaline pH on ferrihydrite ().
Competitive
The competitive
sorption between trace elements in anionic form has received attention. However,
a systematic investigation of the relative competition for sorption onto variable-charge
minerals and soils among various anions with different binding affinities is
rather limited.
Manning and Goldberg
(1996) studied pH effects and competing molybdate and arsenate ions onto goethite
and gibbsite. Molybdate at 50% of surface coverage decreased the sorption of
arsenate only at pH&6.0, whereas arsenate reduced molybdate sorption within
a wider pH range (2.0 to 9.0 for goethite and 2.0 to 8.0 for gibbsite). Their
data suggested that arsenate occupies a fraction of the pH-dependent molybdate
sorption sites on both goethite and gibbsite and that another distinct fraction
of sites has a higher affinity for molybdate than arsenate at low pH.
Recently, Pigna
et al. (2010) carried out experiments on the competition in sorption
between As(III) or As(V) with organic and inorganic ligands onto ferrihydrite.
Table 1 shows the amounts of As(III) and As(V) sorbed at 6.0 pH on ferrihydrite
in the presence of inorganic anions (sulfate, selenate and selenite) and organic
(oxalate, malate, tartrate and citrate) added simultaneously with As(III) or
As(V) at the initial ligand / As (III) or As(V) molar ratio = 1. The parameter
used for comparing the behaviour of the various anions to compete with As(III)
or As(V) was the efficiency of the ligands in preventing the sorption of the
metalloid calculated according to the following expression:
Efficiency = [(As
added alone - As sorbed in the presence of a ligand) / As alone] x 100
In respect to the
behaviour of individual inorganic anions to compete with arsenic for sorption
sites on ferrihydrite the following sequence was observed:
Se04 ≈S04
In particular,
sulphate and selenate did not compete with As(V) or showed a negligible capacity
to compete with As(III), whereas selenite strongly competed with both As(III)
and As(V) ().
The organic anions
had a great influence in inhibiting As sorption according to this sequence:
Citrate & Tartrate
≈ Malate & Oxalate
The greater capacity
of citrate to compete with arsenic compared with other organic ligands must
be attributed to its high capacity to be sorbed on the surfaces of ferrihydrite
and to form strong complexes with Fe.
Desorption of
trace elements in anionic form
In contrast to
sorption studies, relative little information is available on desorption of
trace elements in anionic forms from soil or soil components (Violante et
al., 2008). Desorption of metalloids from soils is particularly affected
by the residence time. O Really et al. (2001) studied the residence time
effect on arsenate desorption by phosphate at a 3- phosphate/arsenate molar
ratio from goethite at different pH values. Initially, the desorption of arsenate
was very fast (35% of As(V) desorbed within 24 h), and then it slowed down.
Total desorption
increased with time, reaching about 65% total desorption after five months.
O'Reilly et al. (2001) did not find significant role of residence time
in preventing arsenate desorption. In contrast, Arai and Sparks, (2002) demonstrated
than the longer the residence time (3 days to 1 year), the greater the decrease
in arsenate desorption by phosphate from a bayerite. Similar results, were found
by Pigna et al. (unpublished
data) who studied the effect of residence time on arsenate desorption by phosphate
(phosphate/arsenate molar ratio of 3) from an Andisol sample containing 42%
of allophanic materials (Vacca et al, 2002). The amounts of arsenate
desorbed by phosphate after one week decreased from 53% to 35% and 22% when
phosphate was added 1, 5, or 15 days, respectively, after arsenate addition
CHEMICAL FRACTIONATION
AND BIOAVAILABILITY OF HEAVY METALS AND METALLOIDS
The biogeochemical
cycle of heavy metals and metalloids has been greatly accelerated by human activities.
Accumulation of metal ions and metalloids in different compartments of the biosphere,
and their possible mobilization under environmentally changing conditions induce
a perturbation of both the structure and function of the ecosystem and might
cause adverse health effects to biota (Fedotov and Miró, 2008). Heavy metals
and metalloids enter an agroecosystem through both natural and anthropogenic
processes. Some soils have been found to have a high background of some trace
elements, which are toxic to plants and wild life, due to extremely high concentrations
of these elements in the parent materials. Anthropogenic processes include
inputs of heavy metals through use of fertilizers, organic manures, and industrial
and municipal wastes, irrigation, and wet and/or dry deposits. These processes
contribute with variable amounts of heavy metals to the agro-ecosystem.
Only a small portion
of heavy metals and metalloids in soil is bioavailable. The mobility and availability
of these pollutants are controlled by many chemical and biochemical processes
such as precipitation-dissolution, sorption-desorption, complexation-dissociation,
and oxidation-reduction. Not all the processes are equally important for each
element, but all of them are affected by soil pH and biological processes. Therefore,
it is crucial to understand some major reactions in soils that control the release
of a specific trace element in the soil and the environment in order to overcome
problems related to deficiency and contamination of these elements.
The determination
of total soil metal content alone is not a good measure of bioavailability and
not a very useful tool to quantify contamination and potential environmental
and human health risks. Total concentrations of metals in soils are a poor indicator
of metal toxicity since metals exist in different solid-phase forms that can
vary greatly in terms of their bioavailability (Krishnamurti et al, 1995a
and 1995b; Krishnamurti et al, 1996; Krishnamurti and Naidu, ;
Huang and Gobran, 2005). Most regulations or guidelines used for protecting
soil from metal pollution (to set maximum& & & & &
permissible &metal concentrations
used for definition of contaminated sites, for setting remediation clean-up
goals, and for limiting loading of metals to soils in fertilizers and wastes)
are still based on assessing the total concentration of metal present in the
soil. Recently, some countries have started to introduce the concept of bioavailability
in their regulations concerning environmental protection and remediation. The
bioavailability of heavy metals and metalloids, their biological uptake, and
their eco-toxicological effects on the soil biota can be better understood in
terms of their chemical speciation. The mobility and bioavailability, and hence
potential toxicity of metal in the soil depend on its concentration in soil
solution, the nature of its association with other soluble species, and soil
ability to release the metal from the solid phase to replenish that removed
from soil solution by the plants (Krishnamurti and Naidu, 2002; Huang and Gobran,
2005; Krishnamurti et al, 2007).
The identification
and quantification of metals associated with predefined phases or soil components
is defined as &fractionation analysis& according to JUPAC recommendation
(Templeton et al, 2000). The extractants used in the sequential
extraction procedures (SEPs) are intended to simulate conditions under which
metals associated with certain mineralogical phases can be released (Fedotov
and Miró, 2008). Fractionation of heavy metal cations into operationally defined
forms under the sequential action of given reagents with increasing aggressiveness
is a common approach for distinguishing various species of trace elements according
to their physicochemical mobility and potential bioavailability () (Ure and Davison, 2001; Hlavay et al. 2004; Fedotov and Miró,
2008). Fractionation of elements existing in anionic forms needs obviously a
specific sequence of leaching reagents, as discussed later.
Sequential extraction
schemes for metal partitioning in environmental solid samples have recently
been summarized in a comprehensive review by Filgueiras et al. (2002)
and in a JUPAC technical report by Hlavay et al. (2004).
A more sophisticated
SEP, which significantly improved the specificity and efficiency of extraction,
by a carefully designed combination of various extractants in order to identify
the specific species contributing to bioavailability was proposed by Krishnamurti
et al. (1995a and b).
The differentiation
of the metal-organic-complex-bound metal species, as distinct from the other
organically bound species was the innovation in the selective sequential extraction
scheme suggested by these authors. The particulate-bound metal species in soils
were fractionated as exchangeable, carbonate-, metal-organic complex, easily
reducible metal oxide-, organic material, amorphous mineral colloid, crystalline
Fe oxide-bound and residual. The metal-organic complex-bound was selectively
extracted using 0.1 M sodium pyrophosphate (pH 10) as the extractant in the
sequential extraction scheme.
Later, Krishnamurti
and Naidu (2000) modified the sequential extraction scheme developed by Krishnamurti
et al. (1995a) sub-fractionating the trace element bound to metal-organic
complexes as extracted by 0.1 M sodium pyrophosphate (). This fraction may contain metal associated with both humic and fulvic
acid fractions of soil organic matter which is bound in metal-organic complexes.
These authors have shown that the modified sequential extraction scheme consisting
of 8 steps of extractions subdivides the metal-organic complex-bound Cd into
metal-fulvate complex-bound and metal-humate complex-bound Cd. Therefore, the
scheme fractionates the solid components into specific &species& with
operationally defined binding mechanisms, i.e. &exchangeable, specifically&sorbed/carbonate-bound,
metal-fulvate complex-bound, metal-humate complex-bound, easily reducible metal
oxide-bound, organic site-bound, amorphous metal oxide-bound, crystalline Fe
oxide-bound, and residual (aluminosilicate lattice-bound). Their data indicate
that the extraction scheme could be used for identifying the specific -operationally&defined
&species& contributing to the bioavailability of the trace element.
Details of the extraction scheme are presented in .
The distribution
of solid-phase fractions of Cu, Zn and Cd of a few typical surface soils in_South
Australia, was carried out by Krishnamurti and Naidu (2000) following the fractionation
scheme reported in The trace elements in these
soils were dominantly (on an average 40% of Cu, 52.4% of Zn and 33.4% of Cd)
associated with the alumino-silicate mineral lattices, identified as residual
fraction in the scheme, followed by the fraction associated with organic sites
(on an average 32.4% of Cu, 28.0% of Zn and 28.5% of Cd) (see Krishnamurti and
Naidu, 2008; their Figures 11.3 and 11.4).
Mobility and plant
uptake of trace elements proceed through the solution phase. However, plant
uptake of an element depends not only on its activity in the solution, but also
on the relation existing between solution ions and solid-phase ions.
Plants take up
essential and non-essential elements from soils in response to concentration
gradients inducted by selective uptake of ions by roots, or by diffusion of
elements in the soil. The accumulation level of heavy metals differs between
and within species (Huang and Cunningham, 1996; McGrath et al. 2002).
Baker (1981) suggested that plants could be classified into three categories:
(1) excluders: those that grow in metal-contaminated soil and maintain
shoot concentration at low level up to a critical soil value above which relatively
unrestricted root-to-shoot transport result, (2) accumulators: those
that concentrate metals in the aerial part, and (3) indicators: the uptake
and transport of metals to the shoot is regulated so that internal concentration
reflects external levels, at least until toxicity occurs. A number of biochemical
reactions occur in plants stressed by heavy metal/metalloids. Most of these
reactions are produced by the displacement of protein cationic centres or the
increase of reactive oxygen species (). Those plants
with better ability to adjust to toxicity effects are able to survive in heavy
metal/metalloids impacted sites and are better candidates for phytoremediation
An attempt to understand
the importance of solid-phase fractions in assessing phytoavailability of Cu,
Zn and Cd using multiple regression analysis was made by Krishnamurti and Naidu
(2002). Phytoavailable Cu, Zn and Cd were found to be significantly correlated
with the metal-fulvate complex-bound Cu (r=0.944, p&0001), exchangeable Zn
(r=0.832, p=0.002) and the metal-fulvate complex-bound Cd (r = 0.824, p=0.002).
It was observed that fulvic complex Cu could explain 89% of variation in phytoavailable
Cu, whereas the metal-fulvate complex-bound element together with exchangeable
element could explain 79% and 92% of variation in phytoavailable Zn and Cd,
respectively. Inclusion of solution element concentration in the regression
analysis was not found to improve the predictability of phytoavailable element.
The regression analysis indicated that phytoavailable Cu, Zn and Cd in these
soils are mainly from solid-phase fractions.
Figure 8. Heavy
metal toxicity in plants. Purple spheres indicate redox active metals and red
and blue are redox inactive metals. The green sphere is a metal centre that
is displaced by a heavy metal (red). The affinity for heavy metals will alter
the activity of the protein and create imbalances and disruption that will lead
to macromolecular damage. However, the cell may adjust to the toxic metals and
signal for reaction to prevent damage.
Recently, Krishnamurti
et al. (2007) have studied copper mobility and phytoavailability on wheat
durum (Triticum durum) grown in polluted and unpolluted Italian soils.
The study was conducted to determine the solid phase distribution of copper
in representative soils of Italy with wide differences in chemical and physicochemical
properties. Selected sites varied in location as well as in current vegetation
and land use. Samples with a high Cu load (132-253 mg kg&1)
from a vineyard cultivation area were also included to study the contamination
effect on Cu distribution among solid-phases. The solid phase fractionation
of Cu in the soils was determined using a 8-step selective sequential extraction
method (. The results indicated that Cu was dominantly
associated with organic binding sites 62.6-74.8%. The relative importance of
solid-phase fractions in assessing Cu phytoavailability by durum wheat
in a greenhouse setting and the effectiveness of two soil tests, the DTPA-TEA
and NH4C1 extraction method for predicting the phytoavailable Cu
was studied. Most Cu was retained by roots with very limited translocation to
the upper plant parts of wheat. A significant correlation (r = 0.960, P = 0.0001)
was found between plant Cu content and the Cu associated with the metal-fulvate
complexes, indicating that phytoavailable Cu was mainly from metal-fulvate complexes.
The contaminated soils had a significantly higher Cu proportion (77%) associated
with organic binding sites, in comparison with that of uncontaminated soil (21.3%),
resulting in higher proportion of phytoavailable Cu.
Special sequences
of extracting reagents are needed for the fractionation of metalloids such as
arsenic and selenium as a consequence of their different oxidation states in
soils. Several steps in fractionation analysis are performed under oxidizing
or reducing conditions, whereby the original oxidation state can be altered.
Although the extraction results might be questioned, they are definitely appropriate
to assess the potential bioavailability of metalloids under environmental changing
conditions (Gleyzes et al. 2002).
Wenzel et al.
(2001) have developed an innovative and simple method for arsenic sequential
extraction which provides the following five extraction steps with reagents
gradually stronger ():
(1) 0.05 M (NH4)2S04,
20&C - 4h; (2) 0.05 M NH4H2P04, 20&C - 16h; (3)
0.2 M NH4+-oxalate buffer in the dark. pH 3.25, 20&C
- 4h; (4) 0.2 M NH4+-oxalate buffer+ ascorbic acid pH
3.25, 96&C - 0.5 (5) HN03/H202 microwave
digestion.
These fine As fractions
appear to be primarily associated with (1) non- (2) sp (3) amorphous and poorly-crystalline hydrous Fe and A (4) well
crystallized hydrous Fe and Al oxides 1; and (5) residual phases. These authors
conducted a study using 20 Austrian soils differing in the level of As contamination
(from 96 to 218 mg As kg-1) and soil characteristics.
The results showed that As was
most prevalent in the two oxalate extractable fractions (30% from step 3 and
27%) from step 4), indicating that As is primarily associated with amorphous
and crystalline Fe and Al oxides. The As fraction extracted by NH4H2PO4 represented
about 10%o of total As and may be useful in providing a relative measure of
specifically-sorbed As in soils that may be potentially mobilized due to changes
in pH or P addition. The amount of readily labile.
As extracted by
(NH4)2S04 is generally small (0.3%), but may
represent the most important fraction related to environmental
risks and has been shown to correlate well with As concentrations in field -collected
soil solutions.
Recently, Branco&(2008),
using Wenzel
et al. (2001) method, have studied the sequential As extraction from
polluted Italian soils collected in Scarlino (Tuscany, Italy), which showed
an arsenic content ranging from 100 to 190 mg As kg-1.
shows As distribution in La Botte soil.
Arsenic was mostly
recovered in the crystalline oxides (59.8%) and short-range Fe and Al oxides
(20%). The As fraction extracted with NH4H2P04
was about 7%, whereas the non-specifically sorbed (easily exchangeable) fraction
that forms outer-sphere complexes onto the mineral surfaces was negligible
(&0.2%). The scarce residual As fraction (11-13%) suggested a low presence
of primary minerals rich in this metalloid.
CONCLUSIONS
Sorption-desorption
processes of trace elements on or from soil components is affected by many factors,
such as pH, nature of the sorbents, redox reactions, and presence of organic
and inorganic ligands. The behavior of foreign ligands on the sorption of trace
elements in cationic form is quite different from that toward elements in anionic
form. In fact, complexation reaction of trace elements in cationic form with
organic and inorganic ligands have an important role to play in their sorption-desorption
processes as well as in their toxicity and phytoavailability, whereas competition
for available sites and/or reduction of the surface charge of the sorbents between
foreign ligands and trace elements in anionic form affect primarily their mobility.
Time of reaction and surface coverage have a great influence on the competitive
sorption between trace elements and organic and inorganic ligands. Special sequences
of extracting reagents are needed for the fractionation of heavy metals and
metalloids in order to identify the species that are more available for plants
and microorganisms.
ACKNOWLEDGMENTS
This research was
supported by the Italian Research Program of National Interest (PRIN), year
2006. DiSSPAPA Number
REFERENCES
Arai, Y., Sparks,
D.L. 2002. Residence time effects on arsenate surface speciation at the
aluminum oxide-water interface. Soil Sci.. 167, 303-314.&&&&&&&&[  ]Baker, A.J.M.
1981. Accumulators and excluders strategies in the response of plants to
heavy metals. J. Plant Nutr. 3, 643-654.&&&&&&&&[  ]Beveridge, T.J.
1989a. Metal ions and bacteria. In: T.J. Beveridge, R.J. Doyle (eds). Metal
Ions and Bacteria, R.J. Wiley, New York, pp: 1-29.&&&&&&&&[  ]Beveridge, T.J.
1989b. Role of cellular design in bacterial metal accumulation and mineralization.
Ann. Rev. Microbiol. 43, 147-171.&&&&&&&&[  ]Borch, T., Kretzschmar,
R., Kappler, A., Van Cappelen, P., Ginder-Vogel, M., Voegelin, A., Campbell,
K. 2010. Biogeochemical redox processes and their impact on contaminant
dynamics. Environ. Sci. & Technol. 44, 15-23.&&&&&&&&[  ]Borda, M.J.,
Sparks, D.L. 2008. Mobility of trace elements in soil environments. In:
A. Violante., P.M. Huang, G.M. Gadd (eds). Biophysico-Chemical Processes of
Metals and Metalloids in Soil Environments, John Wiley & Sons, Hoboken,
NJ, pp: 97-168.&&&&&&&&[  ]Branco, A. 2008.
Mobility and bioavailability of arsenic in italian polluted soils. PhD thesis,
University of Naples (Italy) Federico II.122 p.&&&&&&&&[  ]Buffle, J. 1988.
Complexation reactions in aquatic systems: an analytical approach. Ellis
Horwood, Chichester, UK.&&&&&&&&[  ]Capasso, R.,
Pigna, M, De Martino, A., Pucci, M, Sannino, F., Violante, A. 2004. Potential
remediation of waters contaminated with Cr(III), Cu and Zn by sorption on the
organic polymeric fraction of olive mill wastewater (polymerin) and its derivatives.
Environ. Sci. & Technol. 38, .&&&&&&&&[  ]Catalano, J.G.,
Park, C, Fenter, P., Zhang, Z. 2008. Simultaneous inner- and outer-sphere
arsenate sorption on corundum and hematite. Geochim. Cosmochim. Acta 72, (8)
.&&&&&&&&[  ]Cornell, R.M.,
Schwertmann, U. 1996. The Iron Oxides. Structure, Properties, Reactions
and Uses. VCH Publ., New York.&&&&&&&&[  ]Fedotov, P.S.,
Miró, M. 2008. Fractionation and mobility of trace elements in soils and
sediments. In: A. Violante, P.M. Huang, G.M. Gadd. (eds). Biophysico-Chemical
Processes of Heavy Metals and Metalloids in Soil Environments. Wiley-Jupac Series,
Vol 1 John Wiley & Sons, Hoboken, NY, pp: 467-520.&&&&&&&&[  ]Fendorf, S.E.
1995. Surface reactions of chromium in soils and waters. Geoderma 67, 55-71&&&&&&&&[  ]Fendorf, S.E.,
Eick, M.J., Grossl, P.R, Sparks, D.L. 1997. Arsenate and chromate retention
mechanisms on goethite. I. Surface structure. Environ. Sci. Technol. 31, 315-320.&&&&&&&&[  ]Filgueiras,
A.V., Lavilla, I., Bendicho, C. 2002. Chemical sequential extraction for
metal partitioning in environmental solid samples. J. Environ. Monit. 4, 823-857.&&&&&&&&[  ]Gadd, J.M. 2008.
Transformation and mobilization of metals, metalloids, and radionuclides
by microorganisms. In: A. Violante, P.M. Huang, G.M. Gadd. (eds). Biophysico-Chemical
Processes of Metals and Metalloids in Soil Environments. Wiley-Jupac Series,
Vol 1 John Wiley & Sons, Hoboken, NY pp: 53-96.&&&&&&&&[  ]Gleysez, C,
Tellier, S., Astruc, M 2002. Fractionation
studies of trace elements in contaminated soils and sediments: a review of sequential
extraction procedures. Trends Anal. Chem. 21,451-466.&&&&&&&&[  ]Grossl, P.R,
Eick, M.J., Sparks, D.L., Goldberg, S., Ainsworth, C.C. 1997. Arsenate and
chromate retention on goethite. II. Kinetic evaluation using a p-jump relaxation
technique. Environ. Sci. & Technol. 31, 321-326.&&&&&&&&[  ]Helal, M.I.D.
2006. Sorption/desorption characteristics of heavy metal on amorphous Al
and Fe hydroxides formed in the presence of oxalate, citrate and tannate ligands.
J. Agronomy 5, 440-450.&&&&&&&&[  ]Helavay, J.,
Prohaska, T., Weisz, M, Wenzel, W.W., Stingeder, G.J. 2004. Determination
of trace elements bound to soils and sediment fractions. JUPAC Tech. Rep. Pure
Appl. Chem. 76,415-442.&&&&&&&&[  ]Hsu, P.H. 1989.
Aluminum hydroxides and oxyhydroxides. In: J.B. Dixon, S.B. Weed (eds).
Minerals in Soil Environments. 2n edition, pp: 331-378.&&&&&&&&[  ]Huang, J.V.,
Cunningham, S.D. 1996. Lead phytoextraction: species variation in lead uptake
and translocation. New Phytol. 134, 73-84.&&&&&&&&[  ]Huang, P.M.,
Cobran, G.R 2005. Biogeochemistry of trace elements in the rhizosphere.
Elsevier B.V. Amsterdam.&&&&&&&&[  ]Huang, P.M,
Germida, J.J. 2002. Chemical and biochemical processes in the rhizosphere:
metal pollutants. In: P.M. Huang, J.-M. Bollag, N. Senesi (eds). Interactions
Between Soil Particles and Microorganisms: Impact on the Terrestrial Ecosystem.
John Wiley & Sons, New York, pp: 381-438&&&&&&&&[  ]Jackson, T.A.,
Leppard, G.G. 2002. Energy dispersive x-ray microanalysis and its applications
in biological research. In: A. Violante, P.M. Huang, J.M. Bollag, L. Gianfreda
(eds). Soil Mineral-Organic-Matter-Microorganism Interactions and Ecosystem
Health: Dynamics, Mobility and Trasformations of Pollutants and Nutrients, Developments
in Soil Science, Vol. 28. Elsevier, New York, pp: 219-260.&&&&&&&&[  ]Kampf, N., Scheinost,
A.C., Schultze, D.G. 2000. Oxide Minerals. In; M.E. Sumner (ed). Handbook
of Soil Science, ed., CRC Press, Boca Raton, FL, pp: F125-F168.&&&&&&&&[  ]Kinniburgh,
D.G., Jackson, ML. 1976. Adsorption of alkaline earth, transition and heavy
metal cations by hydrous oxides gels of iron and aluminum. Soil Sci. Soc. Am.
J. 40, 796-799.&&&&&&&&[  ]Kinniburgh,
D.G., Jackson, M.L. 1981. Cation adsorption by hydrous metal oxides and
clays. In: M.A. Anderson, A.S. Rubin (eds). Adsorption of Inorganics at Solid-Liquid
Interfaces. Ann Arbor Science, Ann Arbor, MI, pp: 91-160.&&&&&&&&[  ]Krishnamurti,
G.S.R, Naidu, R 2008. Chemical speciation and bioavailability of trace metals..
In: A. Violante, P.M. Huang, G.M. Gadd (eds). Biophysico-Chemical Processes
of Heavy Metals and Metalloids in Soil Environments. Wiley Interscience, New
York, pp: 419-466.&&&&&&&&[  ]Krishnamurti, G.S.R.,
Pigna, M., Arienzo, M., Violante, A. 2007. Solid-Phase Speciation and Phytoavailability
of Copper in a Few Representative Soils of Italy. Chem. Spec.& Bioav. 19,
57-67.&&&&&&&&[  ]Krishnamurti,
G.S.R., Naidu, R. 2002. Solid-solution speciation and phytoavailability
of copper and zinc in soils. Environ. Sci. Technol. 36,.&&&&&&&&[  ]Krishnamurti,
G.S.R, Naidu, R 2000. Speciation and phytoavailability of cadmium in selected
surface soils of south Australia. Aust. J. Soil Res. 38, 991-1004.&&&&&&&&[  ]Krishnamurti,
G.S.R., Cieslinski, G, Huang, P.M., Van Rees, KC.J. 1997. Kinetics of cadmium
release from soils as influenced by organic acids: implication in cadmium availability.
J. Environ. Qual. 26, 271-277.&&&&&&&&[  ]Krishnamurti,
G.S.R, Huang, P.M., Van Rees, KC.J. 1996. Studies on soil rhizosphere: speciation
and availability of Cd. Chem. Speciat. Bioavailab. 8, 23-28.&&&&&&&&[  ]Krishnamurti,
G.S.R, Huang, P.M., Van Rees, K C. J., Kozak, L. M., Rostad, H. P.W. 1995a.
Speciation of particulate-bound cadmium of soils and its bioavailability.
The Analyst 120, 659-665.&&&&&&&&[  ]Krishnamurti,
G.S.R, Huang, P.M., Van Rees, KC.J., Kozak, L.M, Rostad, H.P.V. 1995b. A
new soil test method for the determination of plant available cadmium in soils.
Commun. Soil Sci. Plant Anal. 26, 857-2867.&&&&&&&&[  ]Lopez, A., Lázaro,
N., Priego, J.M., Marques, A.M.
2000. Effect of pH on the adsorption of nickel and other heavy metals by
Pseudomonas fluorescens 4F39. J. Ind Microbiol. Biotechnol. 24, 146-151.&&&&&&&&[  ]Manning, B.A.,
Goldberg, S. 1996. Modelling competitive adsorption of arsenate with phosphate
and molybdate on oxide minerals. Soil Sci. Soc. Am. J. 60, 121-131.&&&&&&&&[  ]Martin, M.,
Celi, L., Barberis, E., Violante, A., Kozac, L.M., Huang, P.M. 2009. Effect
of humic acid coating on arsenic adsorption on ferrihydrite-kaolinite mixed
systems. Canadian J. of Soil Sci. 89,421-434.&&&&&&&&[  ]McBride, MB.
1982. Cu2+-adsorption characteristics of aluminum hydroxide and
oxyhydroxides. Clays Clay Miner. 30, 21-28.&&&&&&&&[  ]McBride, MB.
1994. Environmental Chemistry of Soils, Oxford University Press, New York&&&&&&&&[  ]McKenzie, R.M
1980. The adsorption of lead and other heavy metals on oxides of manganese
and iron. Aust. J. Soil Res. 21, 505-513.&&&&&&&&[  ]McGrath, S.P.,
Zhao, F.J., Lombi, E. 2002. Phytoremediation of metals, metalloids, and
radionuclides. Adv. Agrie. 75, 1-56.&&&&&&&&[  ]O'Reilly, S.E.,
Strawn, D.G, Sparks, D.L. 2001. Residence time effects on arsenate adsorption/desorption
mechanisms on goethite. Soil Sci. Soc. Am. J. 65, 67-77.&&&&&&&&[  ]Perelomov, L.,
Cozzolino, V., Pigna, M., Violante, A. 2010. Adsorption of Cu and Pb on
goethite in the presence of low molecular mass aliphatic acids. Geomicrobiol.
J. in press.&&&&&&&&[  ]Pigna, M., De
Santis, R., Cozzolino, V., Violante, A. 2010. Influence of organic and inorganic
ligands on the adsorption of arsenate and arsenite onto ferrihydrite. Geoderma
in Press.&&&&&&&&[  ]Senesi, N.
1992. Metal-humic substance complexes in the environment. Molecular and
mechanistic aspects by multiple spectroscopic approach.. In: D.C. Adriano (ed).
Biogeochemistry of Trace Elements. Lewis , Boca Raton, FL, pp: 429. 451.&&&&&&&&[  ]Senesi, N.,
Loffredo, E. 1998. The chemistry of soil organic matter. In; D.L. Sparks
(ed). Soil Physical Chemistry. 2n edition, CRC Press, Boca Raton,
FL, pp: 239-370.&&&&&&&&[  ]Sparks, D.L.
2003. Environmental Soil Chemistry. 2n edition. Academic Press,
San Diego.&&&&&&&&[  ]Sparks, D.L.
1990. Kinetics of Soil Chemical Processes. Academic Press, San Diego, CA.&&&&&&&&[  ]Sposito, G.
1984. The surface chemistry of soils, Oxford Univ. Press, 234 p.&&&&&&&&[  ]Templeton, D.M.,
Ariese, F., Cornells, R., Danielson, L.-G., Munau, H., van Leeuwen, H.P., Lobinski,
R. 2000. Guidelines for terms related to chemical speciation and fractionation
of elements. Definition, structural aspects and methodological& approaches&
(JUPAC Recommendation
2000). Pure Appl. Chem. 72, .&&&&&&&&[  ]Ure, A.M.,
Davison, CM. 2002. Introduction. In: A.M. Ure, CM. Davison (eds). Chemical
Speciation in the Environment, 2n edition, ed, CM., Blackwell Science,
Oxford, pp: 1-5.&&&&&&&&[  ]Vacca, A.,
Adamo, P., Pigna, M., Violante, P. 2002. Properties and classification
of selected soils from the Roccamonfina volcano, Central-Southern Italy. Soil
Sci. Soc. Am. J. 67, 198-207.&&&&&&&&[  ]Violante, A.,
Krishnamurti, G.S.R., Pigna, M. 2008. Mobility of Trace Elements in Soil
Environments. In: A. Violante, P.M. Huang, G.M. Gadd (eds). Biophysico-Chemical
Processes of Metals and Metalloids in Soil Environments. John Wiley & Sons,
Hoboken, NJ, pp: 169-213.&&&&&&&&[  ]Violante, A.,
Ricciardella, M., Pigna, M., Capasso, R 2005. Effects of organic ligands
on the sorption of trace elements onto metal oxides and organo-mineral complexes.
In: P.M. Huang, G.R. Gobran (eds). Biogeochemistry of Trace Elements in the
Rhizosphere, Elsevier B.V, pp: 157-182.&&&&&&&&[  ]Violante, A,
Ricciardella, M., Pigna, M. 2003. Adsorption of Heavy Metals on Mixed Fe-Al
Oxides in the Absence or Presence of Organic Ligands. Water Air and Soil Poll.
145, 289-306.&&&&&&&&[  ]Violante, A,
Pigna, M. 2002. Competitive sorption of arsenate and phosphate on different
clay minerals and soils. Soil Sci. Soc. Am. J. 66, .&&&&&&&&[  ]Wenzel, W.W.,
Kirchbaumer, N., Prohaska, T., Stingeder, T., Lombi, E., Adriano, D.C. 2001.
Arsenic fractionation in soils using an improved sequential extraction procedure.
Anal. Chim. Acta 436, 309-323.&&&&&&&&[  ]Zachara, J.M.,
Ainsworth, C.C., Cowan, C.E., Resch, C.T. 1989. Adsorption of chromate by
subsurface soil horizons. Soil Sci. Soc. Am. J. 53,418-428.&&&&&&&&[  ]
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