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UNCORRECTED PROOF
Chemosphere xxx (2018) xxx-xxx
Contents lists available at ScienceDirect
Chemosphere
journal homepage: www.elsevier.com
Iron and sulfur cycling in acid sulfate soil wetlands under dynamic redox conditions:
A review
Niloofar Karimian∗, Scott G. Johnston, Edward D. Burton
Southern Cross GeoScience, Southern Cross University, Lismore, NSW, 2480, Australia
ARTICLE INFO
Article history:
Available online xxx
Keywords:
Acid sulfate soils
Iron
Sulfur
Dynamic redox
Arsenic
Antimony
ABSTRACT
Acid sulfate soils (ASS) contain substantial quantities of iron sulfide minerals or the oxidation reaction prod-
ucts of these sulfidic minerals. Transformation of iron (Fe) and sulfur (S) bearing minerals is an important
process in ASS wetlands with fluctuating redox conditions. A range of potentially toxic metals and metalloids
can either be adsorbed on or incorporated into the structure of Fe and S bearing minerals. Therefore, transfor-
mation of these minerals as affected by dynamic redox conditions may regulate the mobility and bioavailabil-
ity of associated metals/metalloids. Better understanding of the interaction between Fe/S biogeochemistry and
trace metal/metalloid mobility under fluctuating redox conditions is important for assessing contaminant risk
to the environment. This review paper provides an overview of current knowledge regarding cycling of Fe,
S and selected trace metal/metalloids in ASS wetlands under fluctuating redox conditions and outlines future
research challenges and directions on this subject.
© 2017.
1. Introduction
A complex interplay between hydrology, redox conditions and iron
(Fe) and sulfur (S) mineralogy leads to diverse reactions that deter-
mine contaminant fate and general water quality in acid sulfate soil
(ASS) wetlands. While many of the Fe and S- bearing minerals that
commonly occur in ASS-affected environments are capable of be-
ing potent scavengers for potentially-toxic trace metals and metalloids
under oxidising acidic conditions, many of these minerals are also
metastable and prone to change under reducing conditions. Some trace
metals and metalloids such as arsenic (As) may have a large impact
on the environment. Therefore, a sound understanding of metastable
host-mineral phases and the corresponding behaviour of associated
trace metal/metalloid contaminants under fluctuating redox conditions
is essential for ASS wetlands management.
This review focuses on the formation, characteristics and remedia-
tion of ASS wetlands with a particular emphasis on key Fe and S min-
erals and their behaviour under both oxidising and reducing conditions
and the consequences for potentially toxic trace metals and metalloids
behaviour under dynamic redox conditions.
1.1. Overview of acid sulfate soils
Acid sulfate soils generally contain substantial quantities of Fe sul-
fide minerals or the oxidation reaction products of these sulfidic min-
erals (e.g. acidity and secondary Fe(III)-containing minerals) (Dent,
1986; Dent and Pons, 1995; Fanning et al., 2002).
∗Corresponding author.
Email address: niloofar.karimian@scu.edu.au (N. Karimian)
The global extent of ASS has been estimated at ∼50 million ha
and includes both inland and coastal ASS in Asia, Australia, Scandi-
navia and west Africa (Fig. 1) (Ljung et al., 2009). Although pyrite
(FeS2) is the typically most prevalent Fe sulfide mineral in undis-
turbed ASS materials, other sulfide minerals such as greigite, marca-
site and iron monosulfides such as mackinawite, may also be abundant
in ASS environments (Bush et al., 2004; Dent, 1986; Van Breemen,
1973). Undisturbed, unoxidised sulfide-rich soils which have the po-
tential to become acidic are commonly termed potential acid sulfate
soils (PASS) and they are usually stable under anoxic and waterlogged
conditions (Sullivan et al., 2009). However, a wide range of anthro-
pogenic activities (e.g. drainage and excavation) and natural phenom-
ena (e.g. drought) can cause exposure and oxidation of PASS and gen-
erate significant quantities of acid (Dent, 1986; Sullivan et al., 2008).
The formation of actual acid sulfate soils (AASS) occurs when the
magnitude of sulfuric acid surpasses the buffering capacity of the sed-
iments (Boman et al., 2010; Johnston et al., 2009a; Pons et al., 1982).
Therefore, one of the foremost ASS management strategies is to iden-
tify the PASS materials to avoid exposure, oxidation and further en-
vironmental problems, and then to develop appropriate management
options for these soils (Broughton, 2008; Kraal et al., 2013; Morgan et
al., 2012; Sullivan et al., 2009).
1.2. Environmental hazards of ASS
Oxidation of reduced inorganic sulfur (RIS) species and formation
of AASS has an enduring and complex effect on the soil and sur-
rounding environment (Corfield, 2000; Rosicky et al., 2006). Some
of the major environmental hazards associated with ASS that are re-
lated to the oxidation of RIS include acute acidification of soils and
waters, aluminium (Al) toxicity, trace metal/metalloid mobilisation,
https://doi.org/10.1016/j.chemosphere.2018.01.096
0045-6535/© 2017.
UNCORRECTED PROOF
2 Chemosphere xxx (2018) xxx-xxx
Fig. 1. Worldwide distribution of coastal ASS [after (Ljung et al., 2009)] and the probability of the ASS occurrence in Australia (Source: S. Marvanek, CSIRO).
de-vegetation of the soil, deoxygenation of water bodies and the for-
mation of toxic sulfur gases (e.g. H2S) (Huang et al., 2011).
The adverse environmental and ecological impacts of acute acidifi-
cation and associated metals mobilisation on agricultural crops, terres-
trial, and aquatic environments have been documented in many former
studies (Callinan et al., 2005; Sammut et al, 1996; Wilson et al., 1999;
Mosley et al., 2014). Severe acidification of ASS and resultant de-
crease in soil pH can significantly enhance the solubility and mobility
of many metals following re-wetting of oxidised sediments [e.g. Al,
Fe, zinc (Zn), nickel (Ni), copper (Cu) and vanadium (V)] (Åström,
2001; Macdonald et al., 2004; Simpson et al., 2010; Fitzpatricket al.,
2017; Michaelet al., 2017).
For instance, Al solubility is highly pH-dependent, and potentially
toxic amounts of dissolved Al are common in ASS drainage waters be-
low pH 5.0. High aqueous concentrations of Al can disrupt plant root
systems and can affect plant growth (Sammut et al., 1996). Some eco-
logical hazards of acute aquatic acidification include direct fish mor-
tality and sub-lethal impacts on gilled aquatic organisms (Callinan et
al., 2005; Sammut et al., 1996), a decrease in native fauna and flora
diversity (Sammut et al., 1996) and a decrease or complete loss of ben-
thic organisms (Corfield, 2000). The acidity generated from ASS can
also result in the corrosion of concrete and steel infrastructure (e.g.
tunnels, buildings, bridges and pipelines) which can cause significant
financial losses (Ljung et al., 2009).
Given the broad range of environmental and ecological risks asso-
ciated with ASS, detailed investigations and understanding of the ki-
netics of acidification events and development of effective manage-
ment strategies are important.
2. Formation of acid sulfate soils
2.1. Reduced inorganic sulfur formation
One of the defining characteristics of PASS is the presence of ap-
preciable concentrations ofRIS. Reduced inorganic sulfur species in-
clude, pyrite (FeS2) (Bloomfleld and Coulter, 1974; Van Breemen,
1973), iron monosulfides (FeS) (Bush et al., 2000), greigite (Fe3S4)
(Bush and Sullivan, 1997) and elemental sulfur [S(0)] (Burton et al.,
2006a). Under anoxic conditions and ambient temperature, forma-
tion and accumulation of RIS depends on dissimilatory SO42− reduc-
tion (Fanning et al., 2002). The key components required to form
RIS through microbially-mediated SO42− reduction include a supply of
SO42− (which is usually supplied from seawater), organic matter and
Fe (generally provided by Fe-containing minerals in the sediments)
(Eq. (1)) (Dent, 1986). This reaction is then catalysed by SO42− reduc-
ing bacteria (SRB) (Dent, 1986; Johnston et al., 2004):
Reduced inorganic S can occur in various mineral phases. How-
ever, two forms of Fe sulfide minerals are of particular interest from
an environmental perspective: Fe monosulfides and pyrite (Kraal et
al., 2013). Iron monosulfide minerals (e.g. mackinawite) can form via
the reaction of H2S with free Fe2+ (Eq. (2)) (Marnette et al., 1993;
Rickard, 1995). Some other intermediate species, such as greigite, can
also be formed following partial oxidation of mackinawite (Eq. (3))
(Boursiquot et al., 2001).
(1)
UNCORRECTED PROOF
Chemosphere xxx (2018) xxx-xxx 3
Formation of FeS2has been investigated extensively due to its im-
portant role in coupling the biogeochemical cycling of Fe, S and car-
bon (C), and its implications in the bioavailability of toxic metals in
anoxic marine sediments (Hunger and Benning, 2007; Luther, 1991).
Many studies consider FeS minerals to be important precursors to the
formation of the more thermodynamically stable pyrite (Benning et
al., 2000; Hurtgen et al., 1999; Neretin et al., 2004). However, FeS
phases are not the only necessary prerequisites for FeS2formation, as
FeS2may also form via the reaction of sulfide with other Fe-contain-
ing phases such as fougerite (green rust), goethite or hematite (Rickard
and Luther, 2007). Mackinawite (FeS) is often an important precursor
for pyrite formation in Fe-rich environments due to the more rapid for-
mation of FeS than pyrite in these systems (Rickard and Luther, 2007).
Two key pathways are proposed for FeS2formation from FeS in
sedimentary environments. One of these pathways involves the reac-
tion of FeS (i.e. mackinawite) with H2S (Eq. (4)) (Drobner et al., 1990;
Rickard, 1997). Hydrogen sulfide may also be oxidised and form ze-
rovalent sulfur [S(0)], which reacts with FeS to form FeS2(Eq. (5))
(Coles et al., 2000; Postma and Jakobsen, 1996; Rickard and Luther,
2007):
Nucleation is another possible pathway for pyrite formation which
usually occurs via dissolution and precipitation of some thermody-
namically less stable FeS phases in the sediments (Morse and Luther,
1999; Rickard and Luther, 2007). Contrary to pyrite crystal growth,
nucleation is a slow process which only occurs in FeS-rich systems
that can provide proper active surfaces and high supersaturation for
FeS2nucleation (Rickard and Luther, 2007).
2.2. Oxidation of pyrite and iron monosulfides
In ASS environments pyrite is generally the most abundant iron
sulfide mineral and often represents the largest pool of potential acid-
ity in these systems (Burton et al., 2010; Claff et al., 2011). The oxida-
tion of pyrite may occur at various pH ranges utilizing two oxidants;
oxygen (O2) or Fe3+ (Lin et al., 1998) (Eqs (6) and (7)):
In low pH conditions (<4.0), abundant Fe3+ can catalyse pyrite ox-
idation and acidity generation (Eq. (8)):
About 4mol of H+is generated for each mole of pyrite oxidised
(Lin et al., 1998). However, acute acidification and low pH only oc-
curs when the H+generated via oxidation reactions overwhelms the
acid neutralising capacity (ANC) of the soil.
Although pyrite oxidation is often considered to be the principal
source of potential acidity in ASS environments (Boman et al., 2010;
Claff et al., 2011), iron monosulfides (e.g. nanocrystalline mackinaw-
ite) may also be oxidised very rapidly following exposure to oxi-
dising conditions (Benning et al., 2000; Burton et al., 2009a). Oxi-
dation of iron monosulfides can result in the formation of S(0) and
Fe(III)-oxyhydroxides (e.g. FeOOH) (Eq. (9)) (Burton et al., 2009a).
However, the oxidation and transformation of monosulfide minerals
to S(0) involves the formation of highly reactive intermediate poly-
sulfide species. These compounds can rapidly dissociate to form S(0)
precipitates (Burton et al., 2009a):
Oxidation of S(0) may proceed with O2(Eq. (10)) or alternatively
Fe(III) (Eq. (11)) as electron acceptors, which can result in the simul-
taneous release of acidity (H+) and SO42− (Burton et al., 2006b, 2009a,
2009b):
Under low pH conditions (pH<4.0), Fe3+ (formed following fur-
ther oxidation of Fe2+) can precipitate in the form of a range of
secondary Fe(III) minerals (Lowson, 1982), such as schwertmannite
(Fe8O8(OH)6SO4) and jarosite (KFe3(SO4)2(OH)6) (Eqs (12) and (13))
(Bigham and Nordstrom, 2000; Dent, 1986; Fitzpatrick et al., 2010):
These metastable secondary Fe(III) minerals can be hydrolysed to
thermodynamically more stable minerals (e.g. goethite). This hydrol-
ysis reaction generates a substantial amount of additional acidity, and
thereby sustain the acidity discharge from AASS (Eqs (14) and (15)).
The risk of surface water acidification following the first substantial
rainfall event in Fe(III)-rich ASS-affected environments can be signif-
icant (Ahern et al., 2004; Bigham et al., 1996a):
(2)
(3)
(4)
(5)
(6)
(7)
(8)
(9)
(10)
(11)
(12)
(13)
UNCORRECTED PROOF
4 Chemosphere xxx (2018) xxx-xxx
3. Iron and sulfur behaviour under oscillatory redox conditions
in ASS environments
In regions prone to extreme variations in seasonal rainfall, ASS
wetlands can experience large, periodic redox oscillations as water
levels fluctuate. These redox oscillations generally result in accumula-
tion of various species of Fe-bearing minerals in sediments (Bigham et
al., 1996a; Johnston et al., 2014; Murad and Rojik, 2004). The occur-
rence and stability of these Fe minerals are pH and redox-dependent.
For example, Fe(II) minerals are more stable under reducing condi-
tions, whereas Fe(III) minerals are typically more thermodynamically
stable under oxidising conditions (Lin and Herbert, 1997).
Iron minerals are prevalent in ASS environments and make an un-
deniably important contribution to the geochemical behaviour of trace
metals/metalloids (Stumm and Sulzberger, 1992). Both oxidised (pri-
mary and secondary Fe(III) minerals) and reduced Fe(II) minerals are
extensively involved in the redox equilibria of soil. Iron reduction is
one of the major geochemical changes following inundation of wet-
land sediments. The geochemical consequences of Fe(III) reduction
may include: increased pH, increased dissolved Fe2+ concentrations,
increased exchangeable cations and stimulated formation of new min-
eral phases (Ponnamperuma, 1972). The importance of microorgan-
isms such as dissimilatory iron reducing bacteria (DIRB), is substan-
tial under anaerobic non-sulfogenic conditions where they oxidise or-
ganic matter to reduce Fe(III) oxides; whereas H2S plays an impor-
tant role as a reductant of Fe(III) oxides in sulfide-rich environments
(Lovely, 1991; Dos Santos Afonso and Stumm, 1992).
In sedimentary environments, especially those under anoxic condi-
tions, Fe-S biogeochemical cycling and mineralisation processes can
control trace metal/metalloid bioavailability and acidity generation
(Burton et al., 2011; Caille et al., 2003). For example, microbial reduc-
tion of SO42− and the consequent formation of iron sulfides (following
re-flooding) can immobilise Fe2+ and sequester toxic trace metal/met-
alloids (particularly chaclophiles) and is considered an effective re-
mediation approach in some ASS environments (Burton et al., 2008a,
2007; Johnston et al., 2010a).
However, several environmental factors can constrain dissimila-
tory SO42− reduction including: pH [sulfate reducing bacteria (SRB)
are relatively inactive at pH < 5.0] (Blodau, 2006), redox conditions,
the availability of labile C and SO42−, and SO42− competing with
other electron acceptors [e.g. Fe(III) and nitrate (NO3)] (Muyzer and
Stams, 2008; Burton et al., 2011). Given all these influencing factors,
re-flooding of oxidised ASS sediments may not result in the re-es-
tablishment of SO42− reducing conditions under some circumstances
(Leyden et al., 2016; Mosley et al., 2017). Therefore, a sound under-
standing of the main factors controlling dissimilatory SO42− reduction,
and how they may apply given the characteristics and site-specific cir-
cumstances of a specific location, is essential prior to the development
of effective remediation strategies for oxidised ASS wetlands sedi-
ments.
Under oxic conditions, oxidation of Fe2+ can produce poorly crys-
talline Fe(III) minerals which can rapidly precipitate and generate
substrate for Fe(III) reducing bacteria (Megonigal et al., 2003). Aer-
obic oxidation of Fe2+ generally occurs at the anaerobic and aero-
bic boundaries in waterlogged sediments (Schwertmann and Cornell,
2000). The rate of Fe2+ oxidation is controlled by different factors,
such as, pH, O2levels and the delivery rate of Fe2+ (Neubauer et
al., 2005). Seasonal variations in hydrology and redox conditions of
ASS wetlands can promote reductive dissolution of Fe(III)-containing
phases (Johnston et al., 2011, 2005), which can cause rapid mineral
transformations and influence trace element mobilisation (Burton et
al., 2006c; Johnston et al., 2012).
Reduction and dissolution rates of the Fe(III) oxides depends on
different factors such as their degree of crystallinity, mineralogy, sur-
face area, microbial activity and the possible existence of chelating
agents (Burton et al., 2010; Lovely, 1991). Iron(II)-induced transfor-
mation of different Fe oxides can take place via various reactions in-
cluding isomorphic substitution, surface sorption of Fe2+ to Fe(III) ox-
ides, electron flow between Fe2+ and Fe3+ and the consequent forma-
tion of thermodynamically more stable Fe oxides (Burton et al., 2007,
2008b; Burton and Johnston, 2012; Hansel et al., 2005; Pedersen et al.,
2005; Williams and Scherer, 2004).
3.1. Iron and sulfur minerals under oxidising conditions
In ASS environments Fe(III) minerals consist mainly of either ox-
ides, hydroxides or oxy-hydroxides, oxyhydroxysulfates (Pedersen,
2006). Iron oxides, due to their high specific surface area, are impor-
tant scavengers for trace metals and some of the toxic trace metal-
loids {e.g. arsenic (As)] (Appelo and Postma, 2005; Pedersen, 2006;
Rozan et al., 2002). Therefore, Fe oxides can control the concentra-
tions of these dissolved trace metals and metalloids, and thereby, in-
fluence surface/sub-surface water chemistry.
Although Fe(III) is the dominant oxidation state, Fe(II) (e.g. in FeO
and Fe(OH)2) and mixed-valency (e.g. green rust (GR) and magnetite)
oxides are also found in nature (Cornell and Schwertmann, 2003).
Poorly crystalline minerals such as ferrihydrite and schwertmannite
are common in ASS and acid mine drainage (AMD) environments
(Cornell and Schwertmann, 2003). This variability in the degree of
Fe oxide crystallinity is highly dependent on the geochemical char-
acteristics of the matrix in which the minerals were formed (Cornell
and Schwertmann, 2003). In ASS landscapes, Fe(III) minerals such as
jarosite, ferrihydrite, schwertmannite and goethite, are prevalent (Bur-
ton et al., 2006c; Claff et al., 2011).
The Eh-pH diagram of Fig. 2 shows the stability of iron phases and
dissolved species in SO42- rich systems. The diagram shows that Fe2+
is soluble under lower redox levels and up to pH ∼8. However, Fe(III)
is soluble at low pH (less than ∼3) and oxidising conditions and it
percipitates in the form of some Fe(III)-bearing compounds such as,
jarosite (at pH = < ∼3.0), schwertmannite, lepidocrocite, ferrihydrite
(between pH ∼4.5 to ∼6.0) and green rust sulfate and goethite (pH
=>6.0) (see Fig. 2) (Bigham et al., 1996b; Karimian et al., 2017a).
Goethite is a thermodynamically stable phase and other mineral
phases tend to transform to this mineral over time, but the metastable
phases (e.g. green rust, schwertmannite and ferrihydrite) are often the
initial precipitates (Burton et al., 2007, 2008b; Jones et al., 2009;
Hansel et al., 2005; Karimian et al., 2017a).
3.1.1. Jarosite
Jarosite [KFe3(SO4)2(OH)6] is a metastable oxyhydroxysulfate
mineral with yellow-brown colour and rhombohedral morphology
(<500nm) (Bigham and Nordstrom, 2000; Fanning et al., 2002),
which forms under SO42--rich and acidic conditions such as ASS
and AMD wetlands (Bigham and Nordstrom, 2000; Johnston et al.,
(14)
(15)
UNCORRECTED PROOF
Chemosphere xxx (2018) xxx-xxx 5
2011a; Sánchez España et al., 2007). Jarosite is a crystalline mineral
with well-developed long-range order (Scheinost and Schwertmann,
1999) and is often diagnostic of disturbed ASS landscapes (Fitzpatrick
et al., 2009; Johnston et al., 2009b, 2009c). Furthermore, jarosite for-
mation and dissolution can influence the potential and actual pools of
acidity in ASS (McElnea and Ahern, 2004).
Various cations and anions can be substituted into the structure of
jarosite, which may then be released following jarosite dissolution. As
a result, jarosite affects environmental bioavailability and mobility of
trace elements (Welch et al., 2007). For example, jarosite is an im-
portant host phase for both As and antimony (Sb) in wetland sedi-
ments (Johnston et al., 2010b; Tighe et al., 2013). It may undergo var-
ious mineralogical transformations under Fe2+-rich and reducing con-
ditions, which consequently influences As and Sb mobility (Tighe et
al., 2013; Johnston et al., 2010b, 2012). There are different factors that
control trace metals release from the jarosite structure upon dissolu-
tion, the most important of which are: pH, organic ligands, saturation
index of the solution, surface absorption, and the solubility of the ions
(Welch et al., 2007).
3.1.2. Ferrihydrite
Ferrihydrite is a reactive, metastable, reddish brown and poorly
crystalline Fe(III) oxy-hydroxide (Fe5O3(OH)9) with a large surface
area (>200m2g−1) which exists in many soils and sediments (Cornell
and Schwertmann, 2003). Ferrihydrite usually aggregates in the form
of sub-spherical to hexagonal nanoparticles. Given the relatively
higher solubility and surface area of ferrihydrite, compared with other
Fe(III) minerals, it is usually considered as the most readily available
Fe(III) phase for dissimilatory reducing microorganisms (Roden and
Wetzel, 1996). This mineral usually accumulates and can be the most
abundant Fe oxyhydroxide in ASS and AMD environments follow-
ing pyrite oxidation (Fitzpatrick and Self, 1997; Sammut et al., 1995).
Formation and accumulation of ferrihydrite also occurs in the surface
sediments of seawater inundated ASS wetlands under circumneutral
pH (Johnston et al., 2011a).
Ferrihydrite is thermodynamically unstable and its transformation
(either biotically or abiotically) to more stable phases (e.g. goethite)
can be catalysed by Fe2+ and cysteine (Fredrickson et al., 2003; Hansel
et al., 2005; Pedersen et al., 2005; Yee et al., 2006; Burton et al.,
2011). The Fe2+catalysed transformation of ferrihydrite produces
goethite, lepidocrocite and magnetite (Hansel et al., 2003; Pallud et
al., 2010; Burton et al., 2011). Some of the main factors influencing
the formation and stability of each of these transformation end prod-
ucts are: pH (limited stability below pH 6.0), Fe2+ concentration and
ligand composition (Hansel et al., 2003; Pallud et al., 2010; Tufano et
al., 2009; Zachara et al., 2002).
3.1.3. Goethite
Goethite (α-FeOOH) is one of the most thermodynamically sta-
ble secondary Fe(III) oxides under oxic conditions, ambient temper-
ature and acidic environments such as in ASS and AMD landscapes
(Schwertmann and Taylor, 1989; Sullivan and Bush, 2004; Burton et
al., 2006d). Goethite forms via either in situ transformations of ferri
Fig. 2. Eh-pH diagram of Fe-containing phases in the Fe2+-H2O-SO4-K system at 25 °C
and 1.013bars total pressure. ∑{Fe2+}= 10−4 (green), 10−3 (black), 10−2 (red) and 10−1
(blue), {SO4}=10−2 and {K} =10−2. Diagrams a, b and c show the sequential distribu-
tion of metastable phases and precursors as the more thermodynamically stable phases
are suppressed successively (Taken from Karimian et al., 2017a). (For interpretation of
the references to colour in this figure legend, the reader is referred to the web version of
this article.)
UNCORRECTED PROOF
6 Chemosphere xxx (2018) xxx-xxx
hydrite, schwertmannite and lepidocrocite, or oxidation of dissolved
Fe2+ in soils and sediments (Burton et al., 2007, 2008b; Cornell and
Schwertmann, 2003). This mineral usually has acicular or nee-
dle-shaped crystals with yellowish-brown colour (Cornell and
Schwertmann, 2003) (Fig. 3). In coastal lowland acid sulfate soils
(CLASS), the burial of sediments can cause the transformation of
poorly crystalline schwertmannite to goethite, in a wide range of pH
(Burton et al., 2007; Bigham et al., 1996b; Acero et al., 2006). A range
of metals such as Al, nickel (Ni), chromium (Cr) and copper (Cu) can
either substitute for Fe in goethite (Kaur et al., 2009; Singh et al.,
2000) or be adsorbed to its surface (Singh et al., 2000).
3.1.4. Lepidocrocite
Lepidocrocite (γ-FeOOH) is the solid polymorph of ferric oxy-hy-
droxides such as goethite, and is typically less prevalent than goethite
(Raiswell and Canfield, 2012; Schwertmann and Cornell, 2000). This
mineral forms in circumneutral pH and in environments with redox
oscillations such as tidally inundated ASS (Fitzpatrick and Self, 1997;
Johnston et al., 2011b). Lepidocrocite thermally transforms to
maghemite (γ-Fe2O3) or magnetite (Fe3O4). Transformation of lepi-
docrocite to hematite has also been observed in previous studies, how-
ever, this process is reported to be very slow (Raiswell and Canfield,
2012).
3.1.5. Schwertmannite
Schwertmannite (Fe8O8(OH)6SO4) is a poorly crystalline Fe(III)
oxy-hydroxysulfate. It is a brownish yellow or orange precipitate
and is often widespread in ASS (Burton et al., 2006d; Fitzpatrick
et al., 2008; Sullivan and Bush, 2004) and acidic environments such
as AMD waters (Cornell and Schwertmann, 2003). Natural schwert-
mannite has a pin cushion or hedgehog morphology (acicular crystals
200–400nm in diameter) (Bigham and Nordstrom, 2000) (Fig. 4).
The specific surface area of schwertmannite can reach
∼100–200m2g−1, which is higher than many other iron oxides (ex-
cluding ferrihydrite). This high surface area contributes to the effi-
cacy of schwertmannite as a scavenger for a wide range of trace met-
als in natural settings (Schwertmann and Cornell, 2000). Formation
of schwertmannite in natural conditions requires low pH (3.0–4.5),
the high concentrations of SO42− and Fe(III) and microorganisms
(e.g. Thiobacillus ferrooxidans) (Murad and Rojik, 2004). The min-
eral structure of schwertmannite is similar to that of akaganeite (β-
Fig. 3. A TEM image of needle-shaped crystals and selected area electron diffraction
(SAED) pattern of goethite (inset), formed during Fe2+-induced transformation of syn-
thetic K-jarosite at pH 7.0 (image N. Karimian).
Fig. 4. A SEM image of hedgehog shaped crystals of schwertmannite, formed during
oxic incubation of surface sediments from a freshwater re-flooded ASS wetland (image
N. Karimian).
FeOOH), but akaganeite forms ASS with in high chloride concen-
trations. Schwertmannite transforms to goethite and releases acidity
(Bigham et al., 1990; Regenspurg et al., 2004). Transformation rates
of schwertmannite to goethite are relatively slow under low pH con-
ditions (Burton et al., 2008b; Jönsson et al., 2005). However, under
near-neutral, anoxic conditions, schwertmannite rapidly transforms to
goethite following catalytic reaction with dissolved Fe2+ (Burton et al.,
2007). Given SO42− in schwertmannite can undergo some structural
substitutions by, for example, arsenate and other oxy-anions, it can
be considered an important host phase for toxic trace metals and met-
alloids, and therefore, is an important control on their environmental
bioavailability (Acero et al., 2006; Burton et al., 2008b; Fukushi et al.,
2004; Regenspurg and Peiffer, 2005).
3.1.6. Green rust (GR)
Green rusts are highly reactive, oxygen sensitive Fe2+/Fe3+ layered
double hydroxides with greenish-blue colour and a common formula
of [Fe2+(1-x) Fe3+x(OH)2]x+[(x/n)An−, mH2O]x− (Ahmed et al., 2010;
Bingjie et al., 2014; Randall et al., 2001) (Fig. 5) Iron (III) can be sub-
stituted for Fe2+ in the trioctahedral layer of GR which can positively
charge the hydroxide (OH−) layer of Fe(OH)2. This positive charge
can be balanced by hydrated anions such as SO42−, chloride(Cl−) or
carbonate (CO32−) (Mitsunobu et al., 2008). Green rust can forme as
an intermediary phase during the Fe2+-induced transformation of a
range of iron oxides (e.g. lepidocrocite, magnetite and goethite) via
biotic and abiotic pathways under circumneutral to slightly alkaline
reducing conditions (Mitsunobu et al., 2008). Given their high reac-
tive surface area and capability of reducing trace metals and metal-
loids [e.g. Sb(V), Cr(VI) and U(VI)], GR can influence trace metal
mobility and environmental bioavailability under suboxic conditions
(Loyaux-Lawniczak et al., 2000; Mitsunobu et al., 2008; O'Loughlin
et al., 2003).
3.2. Iron and sulfur minerals under reducing conditions
3.2.1. Pyrite
Pyrite is the most ubiquitous sulfide in the earth's surface and ac-
cumulates in sediments under anaerobic conditions. Although pyrite is
not the sole sulfide mineral in these environments, its reactivity is a
cardinal point in detrimental environmental problems associated with
UNCORRECTED PROOF
Chemosphere xxx (2018) xxx-xxx 7
Fig. 5. A TEM image and selected area electron diffraction (SAED) pattern of (a)
GR-SO4displaying the hexagonal particles and GR-SO4diffraction patterns; and (b)
SEM image of GR-SO4(Ahmed et al., 2010).
ASS and AMD -affected environments (Murphy and Strongin, 2009;
Postma, 1982; Rickard and Luther, 2007).
Australian coastal lowlands typically contain varying amounts of
framboidal pyrite and organic matter (Rigby et al., 2006). The rate
of pyrite oxidation and acidity generation depends on factors such as,
the availability of dissolved O2(which depends directly on microbial
activity and sediment organic matter content), the surface area and
the mineral morphology (Lowson, 1982; Pugh et al., 1981). In natural
ASS environments, pyrite exists in various morphologies, including
fine-grained crystals, euhedral particles, framboids or in-filled fram-
boids (Fig. 6).
Pyrite is a sink for trace metals and toxic metalloids (e.g. As) under
reducing conditions (Claff et al., 2011; Murad and Rojik, 2004). The
onset of high redox conditions leads to oxidation of pyrite with con-
comitant discharge of these trace metals to the environment, thereby
posing serious environmental risks (Åström, 1998b; Nordmyr et al.,
2008a; Sohlenius and Öborn, 2004).
3.2.2. Mackinawite
Mackinawite (FeSm) is a nano-crystalline, tetragonal, black-col-
ored, iron monosulfide that belongs to the acid volatile sulfur (AVS)
fraction of reduced S compounds (Bush and Sullivan, 1997; Coles
et al., 2000; Rickard and Luther, 2007; Schwertmann and Cornell,
2000). Mackinawite is a reaction product of microbial SO42− reduc-
tion in Fe-rich systems (Burton et al., 2009b; Wilkin and Ford, 2006).
As reported in several studies, the black colour of some ASS sed-
iments comes from iron monosulfides such as FeSm(Boman et al.,
2008). Under reducing conditions, FeSmis the first precipitate to form
following reaction between dissolved Fe2+ and sulfide during the ini-
tial stages of iron sulfide formation (Benning et al., 2000; Rickard
and Luther, 2007; Rickard and Morse, 2005). In reducing freshwa-
ter environments and marine systems, accumulation of FeSmcan hin-
der the formation of FeS2due to rapid consumption of H2S (Boesen
and Postma, 1988), slow reaction between H2S and FeS (Benning
et al., 2000; Schoonen and Barnes, 1991), competition for organic
matter, and conversion of H2S to HS−at higher pH (Rickard and
Morse, 2005). Mackinawite is also a sink for a wide range of trace
metals, scavenging them via diverse reactions such as the forma-
tion of surface complexes and insoluble metal sulfides (Billon et
Fig. 6. Scanning electron micrographs of some pyrite morphologies (a and b) euheudral crystals; (c) a cluster of euhedral crystals; (d) a single frambiod; (e) a single framboid and (f)
a cluster of framboids (Kraal et al., 2013).
UNCORRECTED PROOF
8 Chemosphere xxx (2018) xxx-xxx
al., 2001; Jeong and Hayes, 2007; Coles et al., 2000; Burton et al.,
2013, 2014; Niazi and Burton, 2016).
3.2.3. Greigite
Greigite (Fe3S4) is a metastable thiospinel of the Fe and the S ho-
molog of magnetite with a cubic structure and metallic blue-black
colour. Under oxic conditions and low pH, mackinawite transforms to
Fe3S4following oxidation of structural Fe2+ (Boursiquot et al., 2001,
2011; Rickard and Luther, 2007). Both FeSmand Fe3S4are gener-
ally considered as precursors for pyrite (Roberts et al., 2011). How-
ever, Fe3S4and FeS2may form via completely different pathways and
greigite may not necessarily be a precursor for pyrite formation. For-
mation of Fe3S4in soil surfaces of tidally re-flooded wetlands and un-
der slightly acidic conditions have been reported in previous studies
(Burton et al., 2011; Bush and Sullivan, 1997). Burton et al. (2011)
found that formation of greigite can be promoted under fluctuating
tide-induced redox oscillations and slightly acid conditions near the
soil surface of a tidally-re-flooded wetland located in north-eastern
Australia. The results of their study show that formation of greigite
was spatially decoupled from that of pyrite in the soil profile.
4. Effects of dynamic redox conditions on the fate of metal in ASS
environments
Although acidity is usually considered the major product of sul-
fidic material oxidation in ASS environments under highly fluctuat-
ing redox conditions, it is the subsequent mobility of potentially toxic
metals and metalloids from these sediments that often poses more se-
rious risks to the environment (Williamson and Rimstidt, 1994). Un-
der frequently flooded and cyclic redox conditions, the bioavailability
of many trace metals and metalloids is decreased following adsorption
to the sediments and/or co-precipitation with reduced sulfides such as
FeS2and other thermodynamically less stable iron monosulfides (un-
der reducing conditions) and (Bush et al., 2004) and Fe(III)-bearing
minerals (under oxidising conditions) (Burton et al., 2011; Caille et
al., 2003).
Metals and metalloids can be released into the aqueous phase
through oxidation of sulfide mineral host-phases, transformation of
Fe(III) oxides, ion exchange processes and aluminosilicate dissolution
(Broughton, 2008). The rates and magnitude of metal release depends
on different factors, such as rates of acidification, pH, variations in
the redox potential of the system and leaching (Burton et al., 2008a;
Rinklebe and Du Laing, 2011; Johnston et al., 2012a), organic matter
and metal content, salinity and presence of carbonates (Frohne et al.,
2011, 2014).
A key factor controlling the behaviour and bioavailability of trace
metal species is their oxidation state, which directly affects their mo-
bility, precipitation, and dominant adsorption mechanisms
(Syrovetnik et al., 2004; Shaheen et al., 2016). There are various tools
and techniques to investigate Fe and S and metal speciation. Some of
the most important techniques are X-ray powder diffraction (XRD)
and scanning electron microscopy (SEM), transmission electron mi-
croscopy (TEM), Mossbauer spectroscopy and X-ray absorption spec-
troscopy (XAS). These techniques can directly and specifically reveal
characteristics such as, mineralogy, speciation and the oxidation states
of metals and their respective Fe and S mineral host-phase(s) (Ross
et al., 2001; Burton et al., 2008a; Prietzel et al., 2009; Schwertmann
and Cornell, 2000) while sequential extraction procedures (SEPs) in-
directly provide insights regarding metal speciation (Burton et al.,
2009a; Claff et al., 2010; Gleyzes et al., 2002; Rao et al., 2008).
Each set of these techniques has its own strengths and limitations.
The most practical way to investigate a complex sample
is the application of several different tools and techniques in combi-
nation i.e. SEP and other techniques such as XRD or XAS. This com-
bined approach can provide unique and complementary information.
Iron(III) oxide dissolution can occur via either reductive or non-re-
ductive pathways. The non-reductive dissolution is induced by lig-
ands or protons (e.g. citrate and oxalate), without changing the oxida-
tion state of Fe(III), whereas reductive dissolution requires an electron
transfer to Fe(III) (from the reductants), which reduces Fe3+ to Fe2+
(Lovley, 1991; Pedersen, 2006). The abiotic reductive dissolution and
transformation of Fe(III) host-phases in Fe2+- and S(-II)-rich systems
(e.g. re-flooded ASS wetlands) influences some trace metal/metalloid
[e.g. arsenic (As) and antimony (Sb)] bioavailability (Johnston et al.,
2010b; Leuz et al., 2006a; Leuz et al., 2006b; Mitsunobu et al., 2006;
Tighe et al., 2013).
Various metals are associated with ASS runoff and have been iden-
tified in ASS-affected environments such as Al, magnesium (Mg),
Cobalt (Co), Ni, boron (B), V, calcium (Ca), potassium (K), man-
ganese (Mn), and cadmium (Cd), Cr and lead (Pb) (Åström, 1998a;
Åström and Spiro, 2000; Fältmarsch et al., 2008; Nordmyr et al.,
2008a; Frohne et al., 2014; Rinklebe et al., 2016). These metals re-
spond in different ways to variations in redox conditions (Du Laing et
al., 2009). Aluminium for instance, is highly soluble and mobile un-
der oxidising-acidic conditions (Åström, 1998b; Schemel et al., 2000;
Powell and Martens, 2005). Generally, mechanisms influencing the
speciation and solubility of Al are highly pH-dependent (Lundström
and Giesler, 1995). Aluminium solubility will be decreased by in-
creasing the pH of the ASS-affected wetlands, either via mixing the
acidic runoff with higher pH water bodies or following the estab-
lishment of the lower redox conditions (e.g. following re-flooding)
Lundström and Giesler, 1995. Precipitation of Al-hydroxide minerals
could be of environmental concern as it can result in further release of
acidity, low productivity in fisheries and aquaculture industries (Dent,
1986; Sammut et al., 1995).
While some micronutrients, such as Cu, are essential for biotic
functions, other metals including As, mercury (Hg), and Sb are toxic
in trace amounts (Gadd, 2010; Kalis et al., 2006; Kapustka et al.,
2004). Some of these toxic metalloids, such as, Cd and Sb, are con-
sidered important contaminants in many lowland aquifers and wet-
lands (Smedley and Kinniburgh, 2002). For example, As can cause
serious health issues including diabetes, lung and bladder cancer
(Coronado-González et al., 2007; Silvera and Rohan, 2007), and Sb
and Cd can cause lung tumours in humans (Beyersmann and Hartwig,
2008; Silvera and Rohan, 2007). Furthermore, both Sb and Cd have
negative impacts on aquatic organisms and can cause gill formation
disorders and growth and productivity issues (Harper et al., 2009;
Wendelaar-Bonga, 1997). Therefore, given the possible health and en-
vironmental risks linked with some of the potentially trace metals and
metalloids, investigating the behaviour of these elements as affected
by Fe and S cycling under fluctuated redox conditions is worthy of
careful consideration.
For example, in natural environments release of As to the aque-
ous phase may result from oxidation of As-bearing sulfides such as ar-
senopyrite and pyrite (Mandal and Suzuki, 2002; Zhang et al., 2016)
or reductive dissolution of As-bearing Fe/Mn hydroxides (Cheng et
al., 2009; Smedley and Kinniburgh, 2002; LeMonte et al., 2017). Both
As(III) and As(V) have a high tendency to adsorb to iron oxides
surfaces (Jain et al., 1999; Pedersen, 2006; Waychunas et al., 1993;
Burton et al., 2009b). The high affinity of many Fe (oxyhydr)/oxides,
such as lepidocrocite, goethite, hematite, schwertmannite and jarosite,
for As in natural settings exerts significant control on immobilisation
of As in these environments (Cornell and Schwertmann, 2003; Maillot
et al., 2013; Smedley and Kinniburgh, 2002).
UNCORRECTED PROOF
Chemosphere xxx (2018) xxx-xxx 9
In ASS and AMD environments with cyclic redox oscillations,
minerals such as schwertmannite and jarosite may encounter high con-
centrations of biogenic S(−II) and Fe2+(aq) produced by microbial re-
duction of SO42− and Fe(III) oxides under a range of pH conditions.
(Burton et al., 2008a, 2008b; Johnston et al., 2012a; Johnston et al.,
2011a). The catalytic effects of S(−II)(aq) or Fe2+(aq) on the reduc-
tive dissolution and transformation of many Fe (oxyhydr)oxides in
anoxic environments and the consequences for repartitioning of some
of the structurally coprecipitated metalloids (e.g. As) have been re-
ported in many previous studies (Burton et al., 2010, 2012, 2013,
2014; Johnston et al., 2012; Karimian et al., 2017a; Kocar et al., 2010;
Saalfield and Bostick,2009; Zhang et al., 2016). For example, Kocar
et al. (2010), reported that As(V) had a higher affinity for newly
formed magnetite and the residual fraction of primary ferrihydrite
compared with a secondary GR phase that formed during biotic sul-
fidisation of ferrihydrite. Johnston et al. (2012) observed a significant
increase in As(V) mobility during S(-II)-induced reductive transfor-
mation of As(V)-bearing jarosite. A study by Burton et al. (2010) re-
ported substantial decreases in the aqueous concentrations of As dur-
ing Fe2+-catalysed transformation of As(III)- and As(V)-sorbed schw-
ertmannite to goethite under reducing conditions.
The geochemical behaviour of Sb highly depends on its oxidation
state. For example, lower toxicity, but higher solubility has been re-
ported for Sb(V), which mostly occurs as Sb(OH)6-, compared with
Sb(III) in oxic waters (Filella et al., 2002; Mitsunobu et al., 2010). Al-
though Sb has many chemical similarities to its group 15 neighbour
As, and is often found in conjunction with As and sulfur in the natural
environment (Mitsunobu et al., 2008; Tighe et al., 2013), it does not
always exhibit similar behaviour to As (Wilson et al., 2010). However,
in contrast to our rapidly-evolving understanding of the geochemical
behaviour of As, very little is known about Sb behaviour in the envi-
ronment.
In natural sediments, Fe, Al and Mn hydroxides are usually consid-
ered the main host phases of Sb (Chen et al., 2003; Mitsunobu et al.,
2006; Scheinost et al., 2006; Wilson et al., 2010). Similar to As, Cd
and Cr in Fe-rich aqueous systems, Sb concentration strongly interacts
with Fe-containing minerals by (co)precipitation and adsorption (Du
Laing et al., 2009; Wilson et al., 2010). For example, in AMD envi-
ronments, very high concentrations of Sb (several thousand ppm) can
occur in jarosite-rich oxidised samples (Courtin-Nomade et al., 2012).
Therefore, mineralogical transformations of Fe-bearing host phase(s)
under dynamic redox conditions can influence the mobility and fate of
Sb in the environment (Mitsunobu et al., 2010).
Association of Sb with common Fe-bearing phases in ASS/
AMD-affected settings such as ferrihydrite, goethite, hematite, GR
and jarosite has been investigated in a series of sorption experiments
(Karimian et al., 2017a, 2018; Leuz et al., 2006b; Mitsunobu et al.,
2013; Mitsunobu et al., 2010). For example, Mitsunobu et al. (2010)
observed immobilisation and incorporation of Sb(V) into the goethite
structure during transformation of ferrihydrite. Structural incorpora-
tion of Sb(V) into crystalline hematite formed by mineralogical trans-
formation of ferrihydrite at pH = 6.0 was also reported by Mitsunobu
et al. (2013). Various factors influence the affinity of Sb with Fe-con-
taining phases in soils. For example, pH, cyclic redox oscillations and
changes in hydrolysis and speciation of Sb (Pierce and Moore, 1982;
Smith, 1989; Tighe et al., 2013).
The geochemical behaviour of Sb during transformation of a range
of Fe(III)-containing minerals has been previously investigated under
various pH and redox conditions, however still very little is known
about Sb behaviour in anoxic Fe(III)- and organic matter-rich
environments co-occur with high concentrations of dissolved Fe2+.
Furthermore, there is a critical gap in the present knowledge of how Sb
interacts with iron sulfides in anoxic low-temperature environments.
In ASS wetlands oscillatory redox conditions, the above-men-
tioned changes in trace metal/metalloids behaviour are intimately cou-
pled with Fe and S speciation. Therefore, investigating the mecha-
nisms controlling the fate of metal(loids) in ASS-affected environ-
ments is essential for a better understanding of their geochemical be-
haviour (e.g. mobility, speciation) under highly fluctuated redox con-
ditions.
5. Management of the ASS wetlands with dynamic redox
conditions
Any type of natural process which results in translocation of the
hypersulfidic materials can expose them to O2and generate signifi-
cant quantities of sulfuric materials. Therefore, one of the foremost
ASS management strategies is to avoid exposure and oxidation of
RIS-rich sediments (Broughton, 2008; Kraal et al., 2013; Morgan et
al., 2012; Sullivan et al., 2009). However, selecting an appropriate
management strategy for ASS-affected environments will depend on
a wide range of factors, including; the position of the ASS in the
landscape (e.g. vicinity to the appropriate source of water to maintain
or generate anoxic conditions); soil physical and chemical properties;
wetland type and hydrological controls (e.g. tides or rainfall); proxim-
ity to sensitive receiving waters; land use and tenure (Johnston et al.,
2003; Fitzpatrick et al., 2009).
In eastern Australia, many freshwater coastal wetlands have been
severely degraded by excessive drainage. Drains were designed to
moderate the impacts of flash-floods and to carry surface runoff away
from agricultural lands. However, despite some desirable social and
economic outcomes, large-scale artificial drainage of ASS wetlands
has also adversely affected wetland hydrology and severely degraded
the water quality of freshwater wetland and adjacent estuaries
(Johnston et al., 2003, 2004). Over-drainage has enabled enhanced
oxidation of a considerable area of sulfidic material and created en-
hanced connectivity between acidic groundwater and surface receiv-
ing waters (de Weys et al., 2011; Johnston et al., 2004). Consequent
generation and the release of sulfuric acid into adjacent waterways
has led to acute acidification of water bodies, with soil surface and
groundwater contamination by critical concentrations of aqueous Fe
and Al (Indraratna et al., 2002; Sammut et al., 1996).
The environmental degradation caused by over-drainage has led
to various attempts to remediate coastal freshwater ASS wetlands
(Johnston et al., 2003). However, many traditional ASS remediation
approaches, such as floodgate opening, drainage re-design, use of
in-drain weirs and neutralisation with alkaline reagents (e.g. Green et
al., 2007) are limited in scope and do not fundamentally alter the un-
derlying geochemistry of acid generation (Johnston et al., 2004).
Reversing the oxidation process in ASS wetlands by re-flooding,
with either tidal-seawater or freshwater, is an alternative remediation
approach which can be relatively cost-effective because it relies on in-
undation of the oxidised sediments re-establishing conditions where
Fe and S reduction can take place in situ (Johnston et al., 2014). Both
seawater and freshwater re-flooding can stimulate the formation of
new Fe2+ and RIS compounds and generate in situ alkalinity (Eqs (16)
and (17)). However, there are primary hydrological and geochemical
differences between these two remediation methods (Fig. 7) (Johnston
et al., 2014)
UNCORRECTED PROOF
10 Chemosphere xxx (2018) xxx-xxx
Fig. 7. Conceptual illustration of contrasting principal hydrological and geochemical
differences between sea water (tidal) and freshwater (non-tidal) wetlands.
For example, highly soluble anions such as SO42− and bicarbonate
(HCO3−), typically depleted during RIS oxidation, can be resupplied
by seawater inundation of ASS. However, in freshwater restored ASS
wetlands, SO42− remains depleted (relative to highly insoluble Fe(III)
minerals which tend to precipitate in situ), causing a decreased ratio
of SO42− relative to bioavailable Fe(III) [SO42−: Fe(III)]. This may re-
sult in inadequate S(-II) formation to react with available Fe(III) fol-
lowing re-flooding (Burton et al., 2006a; White et al., 1997). This
stoichiometric imbalance resulting from the depletion of SO42− rel-
ative to accumulation of reactive Fe(III)-containing phases in fresh-
water wetlands can influence S and C biomineralisation pathways
(Johnston et al., 2014; Heimann et al., 2010; Postma and Jakobsen,
1996), favouring formation of highly redox sensitive end products
(e.g. [S(0)] and nano-crystalline mackinawite) (Burton et al., 2011;
Keene et al., 2011).
Further, in seawater re-flooding management scenarios, wetland
hydrology becomes controlled by tides, whereas freshwater re-flood-
ing relies principally on rainfall (Fig. 8). Therefore, in seawater
re-flooded wetlands, water levels fluctuate within a narrow range and
are thereby immune to climatic, seasonal cyclic redox oscillations.
In contrast, surface sediments of freshwater re-flooded ASS wetlands
are exposed to both droughts and floods and can experience periodic
extreme redox oscillations (Johnston et al., 2014). The frequency of
these cyclic redox oscillations and the frequency of wet-dry transitions
are potentially increasing (Mosely et al., 2017) and are particularly rel-
evant in a dynamic climate, such as experienced in Australia (Fig. 8)
(Johnston et al., 2014).
Fig. 8. Analysis of fluctuations in long-term annual water balance (1880–2012) for the
northern coast of NSW region [taken from Johnston et al. (2014)].
Therefore, the RIS-rich surface sediments of freshwater wetlands
may remain vulnerable to exposure during prolonged dry periods
(Johnston et al., 2014; Karimian et al., 2017b, c). And maintaining the
stability of redox sensitive species during drought episodes, remains a
challenging aspect of freshwater re-flooding of ASS (Johnston et al.,
2014).
Seasonal fluctuations in redox conditions can drive very dynamic
and complex biogeochemical processes in surface sediments of fresh-
water re-flooded ASS which may significantly affect the future water
quality of these systems (Fig. 9). Studying the speciation and trans-
formation of Fe and S minerals in wetlands with large seasonal os-
cillation in water level and consequent changes on trace metals and
metalloids behaviour can assist in refining management strategies for
freshwater re-flooded ASS wetlands with cyclic wet-dry transitions
(Karimian et al., 2017c).
Contrary to seawater re-flooding, restoration with freshwater has
received relatively limited attention, specifically under controlled lab-
oratory conditions whereby key processes can be better elucidated.
Therefore, investigations into the consequences of freshwater re-wet-
ting of oxidised surface sediments of ASS wetlands, subjected to ex-
treme redox oscillations at different stages of re-flooding, is essential
to development of effective management strategies for these systems.
6. Conclusions
This review has focussed on some important aspects of the geo-
chemical cycling of Fe, S and trace metal/metalloids in freshwater
re-flooded ASS wetlands. A major focus was on the potential ef-
fects of seasonal redox oscillations on Fe and S speciation in con-
ditions relevant to ASS wetlands and AMD environments. There is
a need for improved interpretation of the important processes which
are likely to occur in these systems under fluctuating redox condi-
tions in order to better predict prospective environmental outcomes
and choose optimal management strategies. For example, given the
soil organic matter plays a significant role in mitigating sulfide oxida-
tion and minimizing acidity generation during dry episodes (Jayalath
et al., 2016), further research is recommended to determine the ef-
fect of various organic matter forms (e.g. fulvic and humic acids) and
contents on the rates and extent of acidity generation during dry/oxi-
dised periods in freshwater re-flooded ASS wetlands. Another further
research possibility is to consider the effects of vegetation types and
different organic matter forms on the rate and magnitude of pH re
Fig. 9. Schematic illustration of the potential relationships between seasonal oscillations
in water level and the biogeochemistry of freshwater re-flooded ASS wetlands.
(16)
(17)
UNCORRECTED PROOF
Chemosphere xxx (2018) xxx-xxx 11
covery under re-flooded/reduced conditions following a dry-wet tran-
sition.
An additional avenue of experimentation could involve investiga-
tion of the role of microorganisms involved in the principal mech-
anisms controlling sediment pH during wet-dry alterations in these
systems. Further mineralogical data collection [e.g. using XAS] is
needed to conclusively identify the FeS phases which form follow-
ing freshwater re-flooding. This data could aid in determining the for-
mation and transformation mechanisms of these minerals and subse-
quent trace metal and metalloid mobility and bioavailability. Further
laboratory and field-based investigations are also required to study
the biotic and abiotic redox-induced transformations of Fe(III)-bear-
ing mineral phases under extreme redox oscillations and exploring
how these transformations influence the geochemical behaviour of as-
sociated trace metals and metalloids (e.g. As, Sb, Cr and Mo) in Fe, S
and natural organic matter-rich ASS-affected systems.
Uncited references
Åström and Björklund, 1995; Blodau, 2002; Bridgham et al., 2006;
Conrad and Rev, 1996; Fanning et al., 2004; Howitt et al., 2007;
Johnston et al., 2009d; Keene et al., 2010; Macdonald et al., 2004a;
Mosley et al., 2014b; Peiffer et al., 2015; Peretyazhko et al., 2005;
Regenspurg, 2002; Schwertmann and Murad, 1983; Van Breemen,
1982; Wan et al., 2013; Yuan et al., 2015.
References
Acero, P., Ayora, C., Torrentó, C., Nieto, J.-M., 2006. The behavior of trace elements
during schwertmannite precipitation and subsequent transformation into goethite
and jarosite. Geochem. Cosmochim. Acta 70 (16), 4130–4139.
Ahern, C.R., Blunden, B., Sullivan, L.A., McElnea, A.E., 2004. Soil Sampling Han-
dling, Preparation and Storage. Acid Sulfate Soils Laboratory Methods Guidelines.
Ahmed, I.A., et al., 2010. Formation of green rust sulfate: a combined in situ time-re-
solved X-ray scattering and electrochemical study. Langmuir 26 (9), 6593–6603.
Appelo, C.A.J., Postma, D., 2005. Geochemistry, Groundwater and Pollution. CRC
press.
Åström, M., 1998a. Mobility of Al, Co, Cr, Cu, Fe, Mn, Ni and V in sulphide-bearing
fine-grained sediments exposed to atmospheric O2: an experimental study. Envi-
ron. Geol. 36 (3–4), 219–226.
Åström, M., 1998b. Mobility of Al, P and alkali and alkaline earth metals in acid sul-
phate soils in Finland. Sci. Total Environ. 215 (1–2), 19–30.
Åström, M., 2001. Effect of widespread severely acidic soils on spatial features and
abundance of trace elements in streams. J. Geochem. Explor. 73 (3), 181–191.
, .
Åström, M., Spiro, B., 2000. Impact of isostatic uplift and ditching of sulfidic sedi-
ments on the hydrochemistry of major and trace elements and sulfur isotope ratios
in streams, Western Finland. Environ. Sci. Technol. 34 (7), 1182–1188.
Benning, L.G., Wilkin, R.T., Barnes, H., 2000. Reaction pathways in the Fe–S system
below 100 C. Chem. Geol. 167 (1), 25–51.
Beyersmann, D., Hartwig, A., 2008. Carcinogenic metal compounds: recent insight
into molecular and cellular mechanisms. Arch. Toxicol. 82, 493–512.
Bigham, J.M., Nordstrom, D.K., 2000. Iron and aluminum hydroxysulfates from acid
sulfate waters. Rev. Mineral. Geochem. 40, 351–403.
Bigham, J.M., Schwertmann, U., Carlson, L., Murad, E., 1990. A poorly crystallized
oxyhydroxysulfate of iron formed by bacterial oxidation of Fe(II) in acid mine wa-
ters. Geochem. Cosmochim. Acta 54 (10), 2743–2758.
Bigham, J.M., Schwertmann, U., Pfab, G., 1996a. Influence of pH on mineral specia-
tion in a bioreactor simulating acid mine drainage. Appl. Geochem. 11 (6),
845–849.
Bigham, J.M., Schwertmann, U., Traina, S.J., Winland, R.L., Wolf, M., 1996b. Schw-
ertmannite and the chemical modeling of iron in acid sulfate waters. Geochim.
Cosmochim. Acta 60, 2111–2121.
.
Blodau, C., 2006. A review of acidity generation and consumption in acidic coal mine
lakes and their watersheds. Sci. Total Environ. 369 (1–3), 307–332.
Bloomfleld, C., Coulter, J., 1974. Genesis and management of acid sulfate soils. Adv.
Agron. 25, 265–326.
Boesen, C., Postma, D., 1988. Pyrite formation in anoxic environments of the Baltic.
Am. J. Sci. 288 (6), 575–603.
Boman, A., Åström, M., Fröjdö, S., 2008. Sulfur dynamics in boreal acid sulfate soils
rich in metastable iron sulfide-The role of artificial drainage. Chem. Geol. 255
(1–2), 68–77.
Boman, A., Fröjdö, S., Backlund, K., Aström, M.E., 2010. Impact of isostatic land up-
lift and artificial drainage on oxidation of brackish-water sediments rich in
metastable iron sulfide. Geochem. Cosmochim. Acta 74 (4), 1268–1281.
Boursiquot, S., Mullet, M., Abdelmoula, M., Genin, J.M., Ehrhardt, J.J., 2001. The dry
oxidation of tetragonal FeS, mackinawite. Phys. Chem. Miner. 28 (9), 600–611.
, .
Broughton, J.D., 2008. Element Behaviour in Inundated Actual Acid Sulfate Soils,
East Trinity, Cairns. James Cook University.
Burton, E.D., Johnston, S.G., Kocar, B., 2014. Arsenic mobility during flooding of
contaminated soil: the effect of microbial sulfate reduction. Environ. Sci. Technol.
48 (23), 13660–13667.
Burton, E.D., Johnston, S.G., Planer-Friedrich, B., 2013a. Coupling of arsenic mobility
to sulfur transformations during microbial sulfate reduction in the presence and ab-
sence of humic acid. Chem. Geol. 343, 12–24.
, .
Burton, E.D., Bush, R.T., Johnston, S.G., Sullivan, L.A., Keene, A.F., 2011. Sulfur
biogeochemical cycling and novel Fe-S mineralization pathways in a tidally
re-flooded wetland. Geochem. Cosmochim. Acta 75 (12), 3434–3451.
Burton, E.D., Johnston, S.G., 2012. Impact of silica on the reductive transformation of
schwertmannite and the mobilization of arsenic. Geochem. Cosmochim. Acta 96
(0), 134–153.
Burton, E.D., Bush, R.T., Sullivan, L.A., 2006a. Fractionation and extractability of sul-
fur, iron and trace elements in sulfidic sediments. Chemosphere 64 (8),
1421–1428.
Burton, E.D., Bush, R.T., Sullivan, L.A., 2006b. Sedimentary iron geochemistry in
acidic waterways associated with coastal lowland acid sulfate soils. Geochem.
Cosmochim. Acta 70 (22), 5455–5468.
Burton, E.D., Bush, R.T., Sullivan, L.A., 2006c. Acid-volatile sulfide oxidation in
coastal flood plain drains: iron-sulfur cycling and effects on water quality. Envi-
ron. Sci. Technol. 40 (4), 1217–1222.
Burton, E.D., Bush, R.T., Sullivan, L.A., Hocking, R.K., Mitchell, D.R.G., Johnston,
S.G., Fitzpatrick, R.W., Raven, M., McClure, S., Jang, L.Y., 2009a. Iron-monosul-
fide oxidation in natural sediments: resolving microbially-mediated S transforma-
tions using XANES, electron microscopy and selective extractions. Environ. Sci.
Technol. 43, 3128–3134.
Burton, E.D., Bush, R.T., Johnston, S.G., Watling, K., Hocking, R.K., Sullivan, L.A.,
Heber, G.K., 2009b. Sorption of arsenic(V) and arsenic(III) to schwertmannite.
Environ. Sci. Technol. 43, 9202–9207.
Burton, E.D., Bush, R.T., Sullivan, L.A., Johnston, S.G., Hocking, R.K., 2008a. Mo-
bility of arsenic and selected metals during re-flooding of iron- and organic-rich
acid-sulfate soil. Chem. Geol. 253 (1–2), 64–73.
Burton, E.D., Bush, R.T., Sullivan, L.A., Mitchell, D.R.G., 2007. Reductive transfor-
mation of iron and sulfur in schwertmannite-rich accumulations associated with
acidified coastal lowlands. Geochem. Cosmochim. Acta 71 (18), 4456–4473.
Burton, E.D., et al., 2010a. Arsenic effects and behavior in association with the
Fe(II)-Catalyzed transformation of schwertmannite. Environ. Sci. Technol. 44 (6).
Burton, E.D., Bush, R.T., Sullivan, L.A., Mitchell, D.R.G., 2008b. Schwertmannite
transformation to goethite via the Fe(II) pathway: reaction rates and implications
for iron-sulfide formation. Geochem. Cosmochim. Acta 72 (18), 4551–4564.
, .
.
Bush, R.T., Fyfe, D., Sullivan, L.A., 2004. Occurrence and abundance of monosulfidic
black ooze in coastal acid sulfate soil landscapes. Aust. J. Soil Res. 42 (5–6),
609–616.
Bush, R.T., Sullivan, L.A., 1997. Morphology and behaviour of greigite from a
Holocene sediment in Eastern Australia. Aust. J. Soil Res. 35 (4), 853–861.
Bush, R.T., Sullivan, L.A., Lin, C., 2000. Iron monosulfide distribution in three coastal
floodplain acid sulfate soils, eastern Australia. Pedosphere 10, 237–245.
Caille, N., Tiffreau, C., Leyval, C., Morel, J.L., 2003. Solubility of metals in an anoxic
sediment during prolonged aeration. Sci. Total Environ. 301, 239–250.
Callinan, R., Sammut, J., Fraser, G., 2005. Dermatitis, branchitis and mortality in em-
pire gudgeon Hypseleotris compressa exposed naturally to runoff from acid sulfate
soils. Dis. Aquat. Org. 63 (2–3), 247–253.
Chen, J., Gu, B., Royer, R.A., Burgos, W.D., 2003. The roles of natural organic matter
in chemical and microbial reduction of ferric iron. Sci. Total Environ. 307 (1–3),
167–178.
Cheng, H., Hu, Y., Luo, J., Xu, B., Zhao, J., 2009. Geochemical processes controlling
fate and transport of arsenic in acid mine drainage (AMD) and natural systems. J.
Hazard Mater. 165 (1–3), 13–26.
Claff, S.R., Burton, E.D., Sullivan, L.A., Bush, R.T., 2011. Metal partitioning dynam-
ics during the oxidation and acidification of sulfidic soil. Chem. Geol. 286 (3–4),
146–157.
Claff, S.R., Sullivan, L.A., Burton, E.D., Bush, R.T., 2010. A sequential extraction
procedure for acid sulfate soils: partitioning of iron. Geoderma 155 (3–4),
224–230.
UNCORRECTED PROOF
12 Chemosphere xxx (2018) xxx-xxx
Coles, C.A., Rao, R., Yong, R.N., 2000. Lead and cadmium interactions with macki-
nawite: retention mechanisms and the role of pH. Environ. Sci. Technol. 34,
996–1000.
, .
Corfield, J., 2000. The effects of acid sulphate run-off on a subtidal estuarine mac-
robenthiccommunity in the Richmond River, NSW, Australia. ICES (Int. Counc.
Explor. Sea) J. Mar. Sci. 57, 1517–1523.
Cornell, R.M., Schwertmann, U., 2003. The Iron Oxides: Structure, Properties, Reac-
tions, Occurrences and Uses. Wiley VHC, Weinheim, Germany.
Coronado-González, J.A., Del Razo, L.M., García-Vargas, G., Sanmiguel-Salazar, F.,
Escobedo-de la Peña, J., 2007. Inorganic arsenic exposure and type 2 diabetes mel-
litus in Mexico. Environ. Res. 104, 383–389.
Courtin-Nomade, A., et al., 2012. Weathering of Sb-rich mining and smelting residues:
insight in solid speciation and soil bacteria toxicity. Chemie der Erde-Geochem-
istry 72, 29–39.
Dent, D., 1986. Acid Sulphate Soils: a Baseline for Research and Development. ILRI
Publication No. 39, 39. ILRI, Wageningen, Netherlands.
Dent, D.L., Pons, L.J., 1995. A world perspective on acid sulphate soils. Geoderma 67
(3–4), 263–276.
de Weys, J., Santos, I.R., Eyre, B.D., 2011. Linking groundwater discharge to severe
estuarine acidification during a flood in a modified wetland. Environ. Sci. Technol.
45, 3310–3316.
Dos Santos Afonso, M., Stumm, W., 1992. Reductive dissolution of iron(III) (hydr)ox-
ides by hydrogen sulfide. Langmuir 8 (6), 1671–1675.
Drobner, E., Huber, H., Wächtershäuser, G., Rose, D., Stetter, K.O., 1990. Pyrite for-
mation linked with hydrogen evolution under anaerobic conditions. Nature 346,
742–744.
Du Laing, G., Rinklebe, J., Vandecasteele, B., Meers, E., Tack, F.M.G., 2009. Trace
metal behaviour in estuarine and riverine floodplain soils and sediments: a review.
Sci. Total Environ. 407 (13), 3972–3985.
Fältmarsch, R.M., Åström, M.E., Vuori, K.-M., 2008. Environmental risks of metals
mobilised from acid sulphate soils in Finland: a literature review. Boreal Environ.
Res. 13, 444–456.
, .
Fanning, D.S., Rabenhorst, M.C., Burch, S.N., Islam, K.R., Tangren, S.A., 2002. Sul-
fides and Sulfates. Soil Mineralogy with Environmental Applications. 229–260.
Filella, M., Belzile, N., Chen, Y.W., 2002. Antimony in the environment: a review fo-
cused on natural waters: II. Relevant solution chemistry. Earth Sci. Rev. 59,
265–285.
Fitzpatrick, R.W., Powell, B., Marvanek, S., 2008. Atlas of Australian acid sulfate
soils. In: Fitzpatrick, R.W., Shand, P. (Eds.), Inland Acid Sulfate Soil Systems
across Australia. pp. 75–89, CRC LEME Open File Report No. 249. (Thematic
Volume).
Fitzpatrick, R.W., Shand, P., Merry, R.H., 2009. Chapter 3. Acid sulfate soils. In: Jen-
nings, J.T. (Ed.), Natural History of the Riverland and Murraylands. Royal Society
of South Australia (Inc.) Adelaide, South Australia, pp. 65–111.
Fitzpatrick, R., Shand, P., Mosley, L., 2017. Acid Sulfate Soil Evolution Models and
Pedogenic Pathways during Drought and Reflooding Cycles in Irrigated Areas and
Adjacent Natural Wetlands. Geoderma.
Fredrickson, J.K., Kota, S., Kukkadapu, R.K., Liu, C., Zachara, J.M., 2003. Influence
of electron donor/acceptor concentrations on hydrous ferric oxide (HFO) bioreduc-
tion. Biodegradation 14 (2), 91–103.
Frohne, T., Rinklebe, J., Diaz-Bone, R.A., 2014. Contamination of floodplain soils
along the Wupper River, Germany, with As, Co, Cu, Ni, Sb, and Zn and the impact
of pre-definite redox variations on the mobility of these elements. Soil Sediment
Contam.: Int. J. 23 (7), 779–799.
Frohne, T., Rinklebe, J., Diaz-Bone, R.A., Du Laing, G., 2011. Controlled variation of
redox conditions in a floodplain soil: impact on metal mobilization and biomethy-
lation of arsenic and antimony. Geoderma 160 (3–4), 414–424.
Fukushi, K., Sato, T., Yanase, N., Minato, J., Yamada, H., 2004. Arsenate sorption on
schwertmannite. Am. Mineral. 89 (11–12), 1728–1734.
Gadd, G.M., 2010. Metals, minerals and microbes: geomicrobiology and bioremedia-
tion. Microbiol. 156, 609–643.
Gleyzes, 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, 451–467.
Green, R., Waite, T.D., Melville, M.D., 2007. Treatment of acid sulfate soil drainage
by direct application of alkaline reagents. Water Air Soil Pollut. 178 (1–4), 59–68.
Hansel, C.M., Benner, S.G., Fendorf, S., 2005. Competing Fe(II)-induced mineraliza-
tion pathways of ferrihydrite. Environ. Sci. Technol. 39 (18), 7147–7153.
Hansel, C.M., et al., 2003. Secondary mineralization pathways induced by dissimila-
tory iron reduction of ferrihydrite under advective flow. Geochem. Cosmochim.
Acta 67 (16), 2977–2992.
Harper, D.D., Farag, A.M., Hogstrand, C., MacConnell, E., 2009. Trout density and
health in a stream with variable water temperatures and trace element concentra-
tions: does a cold-water source attract trout to increased metal exposure?. Environ.
Toxicol. Chem. 28, 800–808.
Heimann, A., Jakobsen, R., Blodau, C., 2010. Energetic constraints on H2-dependent
terminal electron accepting processes in anoxic environments: a review of observa-
tions and model approaches. Environ. Sci. Technol. 44 (1), 24–33.
, .
Huang, P.M., Li, Y., Sumner, M.E., 2011. Handbook of Soil Sciences: Resource Man-
agement and Environmental Impacts. CRC Press.
Hunger, S., Benning, L.G., 2007. Greigite: a true intermediate on the polysulfide path-
way to pyrite. Geochem. Trans. 8, 1–1.
Hurtgen, M.T., Lyons, T.W., Ingall, E.D., Cruse, A.M., 1999. Anomalous enrichments
of iron monosulfide in euxinic marine sediments and the role of H2S in iron sul-
fide transformations: examples from Effingham Inlet, Orca Basin, and the Black
Sea. Am. J. Sci. 299 (7–9), 556–588.
Indraratna, B., Glamore, W.C., Tularam, G.A., 2002. The effects of tidal buffering on
acid sulphate soil environments in coastal areas of New South Wales. Geotech.
Geol. Eng. 20 (3), 181–199.
Jain, A., Raven, K.P., Loeppert, R.H., 1999. Arsenite and arsenate adsorption on ferri-
hydrite: surface charge reduction and net OH-release stoichiometry. Environ. Sci.
Technol. 33 (8), 1179–1184.
Jayalath, N., Fitzpatrick, R.W., Mosley, L., Marschner, P., 2016. Type of organic car-
bon amendment influences pH changes in acid sulfate soils in flooded and dry con-
ditions. J. Soils Sediments 16 (2), 518–526.
Jeong, H.Y., Hayes, K.F., 2007. Reductive dechlorination of tetrachloroethylene and
trichloroethylene by mackinawite (FeS) in the presence of metals: reaction rates.
Environ. Sci. Technol. 41, 6390–6396.
Johnston, S., Kroon, F., Slavich, P., Cibilic, A., Bruce, A., 2003. Restoring the Bal-
ance: Guidelines for Managing Floodgates and Drainage Systems on Coastal
Floodplains. NSW Agriculture, Wollongbar 1–47.
Johnston, S.G., Burton, E.D., Aaso, T., Tuckerman, G., 2014. Sulfur, iron and carbon
cycling following hydrological restoration of acidic freshwater wetlands. Chem.
Geol. 371 (0), 9–26.
Johnston, S.G., et al., 2010a. Abundance and fractionation of Al, Fe and trace metals
following tidal inundation of a tropical acid sulfate soil. Appl. Geochem. 25 (3),
323–335.
Johnston, S.G., et al., 2012a. Arsenic mobilization and iron transformations during sul-
fidization of As(V)-bearing jarosite. Chem. Geol. 334 (0), 9–24.
Johnston, S.G., et al., 2009a. Changes in water quality following tidal inundation of
coastal lowland acid sulfate soil landscapes. Estuar. Coast Shelf Sci. 81 (2),
257–266.
Johnston, S.G., Hirst, P., Slavich, P.G., Bush, R.T., Aaso, T., 2009b. Saturated hy-
draulic conductivity of sulfuric horizons in coastal floodplain acid sulfate soils:
variability and implications. Geoderma 151 (3–4), 387–394.
Johnston, S.G., Keene, A.F., Burton, E.D., Bush, R.T., Sullivan, L.A., 2011a. Iron and
arsenic cycling in intertidal surface sediments during wetland remediation. Envi-
ron. Sci. Technol. 45 (6), 2179–2185.
, .
Johnston, S.G., et al., 2010b. Arsenic mobilization in a seawater inundated acid sulfate
soil. Environ. Sci. Technol. 44 (6), 1968–1973.
Johnston, S.G., Keene, A.F., Bush, R.T., Burton, E.D., Sullivan, L.A., 2009c. Remedi-
ation of coastal acid sulfate soils by tidal inundation: effectiveness and geochemi-
cal implications. In: Proceedings of the 18th NSW Coastal Conference. p. 18, 3-6
November, Ballina, NSW., Ballina.
Johnston, S.G., et al., 2011b. Iron geochemical zonation in a tidally inundated acid sul-
fate soil wetland. Chem. Geol. 280 (3–4), 257–270.
.
Johnston, S.G., Slavich, P., Hirst, P., 2004. The acid flux dynamics of two artificial
drains in acid sulfate soil backswamps on the Clarence River floodplain, Australia.
Aust. J. Soil Res. 42 (5–6), 623–637.
Jönsson, J., Persson, P., Sjöberg, S., Lövgren, L., 2005. Schwertmannite precipitated
from acid mine drainage: phase transformation, sulphate release and surface prop-
erties. Appl. Geochem. 20 (1), 179–191.
Kalis, E.J.J., Weng, L., Dousma, F., Temminghoff, E.J.M., van Riemsdijk, W.H.,
2006. Measuring free metal ion concentrations in situ in natural waters using the
Donnan Membrane. Technique. Environ. Sci. Technol. 40, 955–961.
Kapustka, L.A., et al., 2004. Ecological Effects of Metals. US EPA. Risk Assessment
Forum, Washington, USA.
Karimian, N., Johnston, S.G., Burton, E.D., 2017a. Antimony and arsenic behavior
during Fe(II)-Induced transformation of jarosite. Environ. Sci. Technol. 51 (8),
4259–4268.
Karimian, N., Johnston, S.G., Burton, E.D., 2017b. Acidity generation accompanying
iron and sulfur transformations during drought simulation of freshwater re-flooded
acid sulfate soils. Geoderma 285, 117–131.
Karimian, N., Johnston, S.G., Burton, E.D., 2017c. Effect of cyclic redox oscillations
on water quality in freshwater acid sulfate soil wetlands. Sci. Total Environ.
581–582, 314–327.
Karimian, N., Johnston, S.G., Burton, E.D., 2018. Antimony and arsenic partitioning
during Fe2+-induced transformation of jarosite under acidic conditions. Chemos-
phere 195, 515–523.
UNCORRECTED PROOF
Chemosphere xxx (2018) xxx-xxx 13
Kaur, N., Gräfe, M., Singh, B., Kennedy, B., 2009. Simultaneous incorporation of Cr,
Zn, Cd and Pb in the goethite structure. Clay Clay Miner. 57, 234–250.
, .
Keene, A.F., et al., 2011. Effects of hyper-enriched reactive Fe on sulfidisation in a
tidally inundated acid sulfate soil wetland. Biogeochemistry 103, 263–279.
Kocar, B.D., Borch, T., Fendorf, S., 2010. Arsenic repartitioning during biogenic sul-
fidization and transformation of ferrihydrite. Geochem. Cosmochim. Acta 74 (3),
980–994.
Kraal, P., Burton, E.D., Bush, R.T., 2013. Iron monosulfide accumulation and pyrite
formation in eutrophic estuarine sediments. Geochem. Cosmochim. Acta 122 (0),
75–88.
LeMonte, J.J., et al., 2017. Sea level rise induced arsenic release from historically cont-
aminated coastal soils. Environ. Sci. Technol.
Leuz, A.-K., Hug, S.J., Wehrli, B., Johnson, C.A., 2006a. Iron-mediated oxidation of
antimony (III) by oxygen and hydrogen peroxide compared to arsenic (III) oxida-
tion. Environ. Sci. Technol. 40 (8), 2565–2571.
Leuz, A.-K., Mönch, H., Johnson, C.A., 2006b. Sorption of Sb (III) and Sb (V) to
goethite: influence on Sb (III) oxidation and mobilization. Environ. Sci. Technol.
40 (23), 7277–7282.
Leyden, E., et al., 2016. Near shore groundwater acidification during and after a hydro-
logical drought in the Lower Lakes, South Australia. J. Contam. Hydrol. 189,
44–57.
Lin, B., Wielicki, B., Minnis, P., Rossow, W., 1998. Estimation of water cloud proper-
ties from satellite microwave, infrared and visible measurements in oceanic envi-
ronments: 1. Microwave brightness temperature simulations. J. Geophys. Res.: At-
mosphere 103 (D4), 3873–3886.
Ljung, K., Maley, F., Cook, A., Weinstein, P., 2009. Acid sulfate soils and human
health–A Millennium Ecosystem Assessment. Environ. Int. 35 (8), 1234–1242.
Lovely, D., 1991. Dissimilatory Fe(III) and Mn(IV) reduction. Microbiol. Rev. 55,
259–287.
Lowson, R., 1982. Aqueous pyrite oxidation by molecular oxygen. Chem. Rev. 82,
461–497.
Loyaux-Lawniczak, S., Refait, P., Ehrhardt, J.-J., Lecomte, P., Génin, J.-M.R., 2000.
Trapping of Cr by formation of ferrihydrite during the reduction of chromate ions
by Fe (II)-Fe (III) hydroxysalt green rusts. Environ. Sci. Technol. 34 (3), 438–443.
Lundström, U.S., Giesler, R., 1995. Use of the aluminium species composition in soil
solution as an indicator of acidification. Ecol. Bull. 44, 114–122.
Luther, G.W., 1991. Pyrite synthesis via polysulfide compounds. Geochem. Cos-
mochim. Acta 55 (10), 2839–2849.
, .
Macdonald, B.C., et al., 2004b. Impacts of runoff from sulfuric soils on sediment
chemistry in an estuarine lake. Sci. Total Environ. 329 (1), 115–130.
Maillot, F., et al., 2013. Structure and reactivity of As(III)- and As(V)-rich schwert-
mannites and amorphous ferric arsenate sulfate from the Carnoulès acid mine
drainage, France: comparison with biotic and abiotic model compounds and impli-
cations for as remediation. Geochem. Cosmochim. Acta 104 (0), 310–329.
Mandal, B.K., Suzuki, K.T., 2002. Arsenic round the world: a review. Talanta 58 (1),
201–235.
Marnette, E.C., Van Breemen, N., Hordijk, K.A., Cappenberg, T.E., 1993. Pyrite for-
mation in two freshwater systems in The Netherlands. Geochem. Cosmochim.
Acta 57 (17), 4165–4177.
McElnea, A.E., Ahern, C.R., 2004. KCl Extractable PH PHKCl and Titratable Actual
Acidity (TAA) - Method Codes 23A and 23F. Acid Sulfate Soils Laboratory Meth-
ods Guidelines.
Megonigal, J.P., Hines, M.E., Visscher, P.T., 2003. Anaerobic metabolism: linkages to
trace gases and aerobic processes. In: Holland, H.D., Turekian, K.K. (Eds.), Trea-
tise on Geochemistry. Pergamon, Oxford, pp. 317–424.
Michael, P.S., Fitzpatrick, R.W., Reid, R.J., 2017. Effects of live wetland plant macro-
phytes on acidification, redox potential and sulphate content in acid sulphate soils.
Soil Use Manag. 33 (3), 471–481.
Mitsunobu, S., Muramatsu, C., Watanabe, K., Sakata, M., 2013. Behavior of antimony
(V) during the transformation of ferrihydrite and its environmental implications.
Environ. Sci. Technol. 47 (17), 9660–9667.
Mitsunobu, S., Takahashi, Y., Sakai, Y., 2008. Abiotic reduction of antimony (V) by
green rust (Fe4(II) Fe2(III)(OH)12 SO4· 3H2O). Chemosphere 70 (5), 942–947.
Mitsunobu, S., Takahashi, Y., Terada, Y., Sakata, M., 2010. Antimony(V) incorpora-
tion into synthetic ferrihydrite, goethite, and natural iron oxyhydroxides. Environ.
Sci. Technol. 44 (10), 3712–3718.
Morgan, B., Burton, E.D., Rate, A.W., 2012. Iron monosulfide enrichment and the
presence of organosulfur in eutrophic estuarine sediments. Chem. Geol. 296–297
(0), 119–130.
Morse, J.W., Luther, G.W., 1999. Chemical influences on trace metal-sulfide interac-
tions in anoxic sediments. Geochem. Cosmochim. Acta 63 (19–20), 3373–3378.
Mosley, L., Zammit, B., Jolley, A.-M., Barnett, L., Fitzpatrick, R., 2014a. Monitoring
and assessment of surface water acidification following rewetting of oxidised acid
sulfate soils. Environ. Monit. Assess. 186 (1), 1–18.
, .
Murad, E., Rojik, P., 2004. Jarosite, Schwertmannite, Goethite, Ferrihydrite and Lepi-
docrocite: the Legacy of Coal and Sulfide Ore mining., 3rd Australian and New
Zealand Soils Conference. University of Sydney, Australia.
Murphy, R., Strongin, D.R., 2009. Surface reactivity of pyrite and related sulfides.
Surf. Sci. Rep. 64, 1–45.
Muyzer, G., Stams, A.J., 2008. The ecology and biotechnology of sulphate-reducing
bacteria. Nat. Rev. Microbiol. 6 (6), 441–454.
Neretin, L.N., et al., 2004. Pyritization processes and greigite formation in the advanc-
ing sulfidization front in the Upper Pleistocene sediments of the Black Sea.
Geochem. Cosmochim. Acta 68 (9), 2081–2093.
Neubauer, S.C., Givler, K., Valentine, S., Megonigal, J.P., 2005. Seasonal patterns and
plant-mediated controls of subsurface wetland biogeochemistry. Ecology 86 (12),
3334–3344.
Niazi, N.K., Burton, E.D., 2016. Arsenic sorption to nanoparticulate mackinawite
(FeS): an examination of phosphate competition. Environ. Pollut. 218 (Suppl. C),
111–117.
Nordmyr, L., Åström, M., Peltola, P., 2008a. Metal pollution of estuarine sediments
caused by leaching of acid sulphate soils. Estuar. Coast. Shelf Sci. 76 (1),
141–152.
Nordmyr, L., Österholm, P., Åström, M., 2008b. Estuarine behaviour of metal loads
leached from coastal lowland acid sulphate soils. Mar. Environ. Res. 66 (3),
378–393.
O'Loughlin, E.J., Kelly, S.D., Cook, R.E., Csencsits, R., Kemner, K.M., 2003. Reduc-
tion of uranium (VI) by mixed iron (II)/iron (III) hydroxide (green rust): formation
of UO2 nanoparticles. Environ. Sci. Technol. 37 (4), 721–727.
Pallud, C., Masue-Slowey, Y., Fendorf, S., 2010. Aggregate-scale spatial heterogeneity
in reductive transformation of ferrihydrite resulting from coupled biogeochemical
and physical processes. Geochem. Cosmochim. Acta 74 (10), 2811–2825.
Pedersen, H.D., 2006. The Transformation of Fe (III) Oxides Catalysed by Fe2+ and
the Fate of Arsenate during Transformation and Reduction of Fe (III) Oxides.
Technical Univ. Denmark, Copenhagen.
Pedersen, H.D., Postma, D., Jakobsen, R., Larsen, O., 2005. Fast transformation of iron
oxyhydroxides by the catalytic action of aqueous Fe(II). Geochem. Cosmochim.
Acta 69 (16), 3967–3977.
.
.
Pierce, M.L., Moore, C.B., 1982. Adsorption of arsenite and arsenate on amorphous
iron hydroxide. Water Res. 16 (7), 1247–1253.
Ponnamperuma, F.N., 1972. The chemistry of submerged soils. Adv. Agron. 24,
29–96.
Pons, L.J., Van Breemen, N., Driessen, P., 1982. Physiography of coastal sediments
and development of potential soil acidity. In: In: Kittrick, J.A., Fanning, D.S.,
Hossner, L.R. (Eds.), Acid Sulfate Weathering, vol. 10, SSSA Special Publication.
Soil Science Society of America, Madison, WI, pp. 1–18.
Postma, D., 1982. Pyrite and siderite formation in brackish and freshwater swamp sedi-
ments (Denmark). Am. J. Sci. 282 (8), 1151–1183.
Postma, D., Jakobsen, R., 1996. Redox zonation: equilibrium constraints on the Fe(III)/
SO42- reduction interface. Geochem. Cosmochim. Acta 60 (17), 3169–3175.
Powell, B., Martens, M., 2005. A review of acid sulfate soil impacts, actions and poli-
cies thatimpact on water quality in Great Barrier Reef catchments, including a case
study on remediation at East Trinity. Mar. Pollut. Bull. 51, 149–164.
Prietzel, J., et al., 2009. Anoxic versus oxic sample pretreatment: effects on the specia-
tion of sulfur and iron in well-aerated and wetland soils as assessed by X-ray ab-
sorption near-edge spectroscopy (XANES). Geoderma 153 (3–4), 318–330.
Pugh, C., Hossner, L., Dixon, J., 1981. Pyrite and marcasite surface area as influenced
by morphology and particle diameter. Soil Sci. Soc. Am. J. 45, 979–982.
Raiswell, R., Canfield, D.E., 2012. The iron biogeochemical cycle past and present.
Geochemical perspectives, 1-1. Europ. Assoc. Geochem. 1, 1–214.
Rao, C.R.M., Sahuquillo, A., Lopez Sanchez, J.F., 2008. A review of the different
methods applied in environmental geochemistry for single and sequential extrac-
tion of trace elements in soils and related materials. Water, Air, Soil Pollut. 189,
291–333.
Randall, S.R., Sherman, D.M., Ragnarsdottir, K.V., 2001. Sorption of As(V) on green
rust (Fe4(II)Fe2(III)(OH)12SO4· 3H2O) and lepidocrocite (γ-FeOOH): surface
complexes from EXAFS spectroscopy. Geochem. Cosmochim. Acta 65 (7),
1015–1023.
. ,
Regenspurg, S., Brand, A., Peiffer, S., 2004. Formation and stability of schwertman-
nite in acidic mining lakes. Geochem. Cosmochim. Acta 68 (6), 1185–1197.
Regenspurg, S., Peiffer, S., 2005. Arsenate and chromate incorporation in schwertman-
nite. Appl. Geochem. 20 (6), 1226–1239.
Rickard, D., 1995. Kinetics of FeS precipitation: Part 1. Competing reaction mecha-
nisms. Geochem. Cosmochim. Acta 59 (21), 4367–4379.
Rickard, D., 1997. Kinetics of pyrite formation by the H2S oxidation of iron (II) mono-
sulfide in aqueous solutions between 25 and 125°C: the rate equation. Geochem.
Cosmochim. Acta 61 (1), 115–134.
Rickard, D., Luther, G.W., 2007. Chemistry of iron sulfides. Chem. Rev. 107 (2),
514–562.
UNCORRECTED PROOF
14 Chemosphere xxx (2018) xxx-xxx
Rickard, D., Morse, J.W., 2005. Acid volatile sulfide (AVS). Mar. Chem. 97 (3–4),
141–197.
Rigby, P.A., Dobos, S.K., Cook, F.J., Goonetilleke, A., 2006. Role of organic matter in
framboidal pyrite oxidation. Sci. Total Environ. 367 (2–3), 847–854.
Rinklebe, J., Du Laing, G., 2011. Factors controlling the dynamics of trace metals in
frequently flooded soils. Dyn. Bioavail. Heavy Metals Root Zone 245–270.
Rinklebe, J., Shaheen, S.M., Schröter, F., Rennert, T., 2016. Exploiting biogeochemi-
cal and spectroscopic techniques to assess the geochemical distribution and release
dynamics of chromium and lead in a contaminated floodplain soil. Chemosphere
150, 390–397.
Roberts, A.P., Chang, L., Rowan, C.J., Horng, C.S., Florindo, F., 2011. Magnetic prop-
erties of sedimentary greigite (Fe3S4): an update. Rev. Geophys. 49 (1).
Roden, E.E., Wetzel, R.G., 1996. Organic carbon oxidation and suppression of
methane production by microbial Fe(III) oxide reduction in vegetated and unvege-
tated freshwater wetland sediments. Limnol. Oceanogr. 41 (8), 1733–1748.
Rosicky, M.A., Slavich, P., Sullivan, L.A., Hughes, M., 2006. Surface and sub-surface
salinity in and around acid sulfate soil scalds in the coastal floodplains of New
South Wales, Australia. Soil Res. 44 (1), 17–25.
Ross, D.S., Hales, H.C., Shea-McCarthy, G.C., Lanzirotti, A., 2001. Sensitivity of soil
manganese oxides: XANES spectroscopy may cause reduction. Soil Sci. Soc. Am.
J. 65, 744–752.
Rozan, T.F., et al., 2002. Iron-sulfur-phosphorus cycling in the sediments of a shallow
coastal bay: implications for sediment nutrient release and benthic macroalgal
blooms. Limnol. Oceanogr. 47 (5), 1346–1354.
Saalfield, S.L., Bostick, B.C., 2009. Changes in iron, sulfur, and arsenic speciation as-
sociated with bacterial sulfate reduction in ferrihydrite-rich systems. Environ. Sci.
Technol. 43 (23), 8787–8793.
Sammut, J., Melville, M.D., Callinan, R.B., Fraser, G.C., 1995. Estuarine acidification:
impacts on aquatic biota of draining acid sulphate soils. Aust. Geogr Stud. 33,
89–100.
Sammut, J., White, I., Melville, M.D., 1996. Acidification of an estuarine tributary in
eastern Australia due to drainage of acid sulfate soils. Mar. Freshw. Res. 47 (5),
669–684.
Sánchez España, J., López Pamo, E., Santofimia Pastor, E., 2007. The oxidation of fer-
rous iron in acidic mine effluents from the Iberian Pyrite Belt (Odiel Basin,
Huelva, Spain): field and laboratory rates. J. Geochem. Explor. 92 (2–3), 120–132.
Scheinost, A., Schwertmann, U., 1999. Color identification of iron oxides and hydrox-
ysulfates: use and limitations. Soil Sci. Soc. Am. J. 63 (5), 1463–1471.
Scheinost, A.C., et al., 2006. Quantitative antimony speciation in shooting-range soils
by EXAFS spectroscopy. Geochem. Cosmochim. Acta 70 (13), 3299–3312.
Schemel, L.E., Kimball, B.A., Bencala, K.E., 2000. Colloid formation and metal trans-
port through two mixing zones affected by acid mine drainage near Silverton, Col-
orado. Appl. Geochem. 15, 1003–1018.
Schoonen, M.A.A., Barnes, H.L., 1991. Reactions forming pyrite and marcasite from
solution: II. Via FeS precursors below 100°C. Geochem. Cosmochim. Acta 55 (6),
1505–1514.
Schwertmann, U., Cornell, R.M., 2000. Iron Oxides in the Laboratory: Preparation and
Characterization.
, .
Schwertmann, U., Taylor, R.M., 1989. Iron Oxides, Minerals in Soil Environments.
SSSA Book Serie, 1. Soil Science Society of America, Madison, WI, 379–438.
Shaheen, S.M., Rinklebe, J., Frohne, T., White, J.R., DeLaune, R.D., 2016. Redox ef-
fects on release kinetics of arsenic, cadmium, cobalt, and vanadium in Wax Lake
Deltaic freshwater marsh soils. Chemosphere 150, 740–748.
Silvera, S.A.N., Rohan, T.E., 2007. Trace elements and cancer risk: a review of the
epidemiologic evidence. CCC (Cancer Causes Control) 18, 7–27.
Simpson, S.L., et al., 2010. Climate-driven mobilisation of acid and metals from acid
sulfate soils. Mar. Freshw. Res. 61 (1), 129–138.
Singh, B., et al., 2000. Structural chemistry of Fe, Mn, and Ni in synthetic hematites as
determined by extended x-ray adsorbtion fine structure spectroscopy. Clay Clay
Miner. 48, 521–527.
Smedley, P.L., Kinniburgh, D.G., 2002. A review of the source, behaviour and distrib-
ution of arsenic in natural waters. Appl. Geochem. 17 (5), 517–568.
Smith, R.J., 1989. Review of Land and Water Management Impacts on Fisheries and
Agricultural Resources in the Lower Macleay - Working Party Report, NSW Agri-
culture.
Sohlenius, G., Öborn, I., 2004. Geochemistry and partitioning of trace metals in acid
sulphate soils in Sweden and Finland before and after sulphide oxidation. Geo-
derma 122 (2–4 SPEC. IIS.), 167–175.
Stumm, W., Sulzberger, B., 1992. The cycling of iron in natural environments: consid-
erations based on laboratory studies of heterogeneous redox processes. Geochem.
Cosmochim. Acta 56 (8), 3233–3257.
Sullivan, L., Fitzpatrick, R., Burton, E., Bush, R., Shand, P., 2008. Assessing Environ-
mental Hazards Posed by Acid Sulfate Soil Materials and Other Inorganic Sul-
fide-containing Soil Materials: Classification Issues.
Sullivan, L.A., Bush, R.T., 2004. Iron precipitate accumulations associated with water-
ways in drained coastal acid sulfate landscapes of Eastern Australia. Mar. Freshw.
Res. 55 (7), 727–736.
Sullivan, L.A., Ward, N.J., Bush, R.T., Burton, E.D., 2009. Improved identification of
sulfidic soil materials by a modified incubation method. Geoderma 149 (1–2),
33–38.
Syrovetnik, K., Puura, E., Neretnieks, I., 2004. Accumulation of heavy metals in Oost-
riku peat bog, Estonia:-site description, conceptual modelling and geochemical
modelling of the source of the metals. Environ. Geol. 45 (5), 731–740.
Tighe, M., Lockwood, P.V., Ashley, P.M., Murison, R.D., Wilson, S.C., 2013. The
availability and mobility of arsenic and antimony in an acid sulfate soil pasture
system. Sci. Total Environ. 463–464 (0), 151–160.
Tufano, K.J., Benner, S.G., Mayer, K.U., Marcus, M.A., Nico, P.S., 2009. Aggre-
gate-scale heterogeneity in iron (hydr)oxide reductive transformations. Vadose
Zone J. 8, 1004–1012.
Van Breemen, N., 1973. Soil forming processes in acid sulphate soils. In: Dost, H.
(Ed.), Acid Sulphate Soils: Proceeding of the International Symposium on Acid
Sulphate Soils, 13-20 August 1972, Wageningen, The Netherlands. ILRI, Wa-
geningen, Netherlands, pp. 66–130.
.
, .
Waychunas, G., Rea, B., Fuller, C., Davis, J., 1993. Surface chemistry of ferrihydrite:
Part 1. EXAFS studies of the geometry of coprecipitated and adsorbed arsenate.
Geochem. Cosmochim. Acta 57 (10), 2251–2269.
Welch, S.A., Christy, A.G., Kirste, D., Beavis, S.G., Beavis, F., 2007. Jarosite dissolu-
tion I - Trace cation flux in acid sulfate soils. Chem. Geol. 245 (3–4), 183–197.
Wendelaar-Bonga, S.E., 1997. The stress response in fish. Physiol. Rev. 77, 591–625.
White, I., Melville, M.D., Wilson, B.P., Sammut, J., 1997. Reducing acidic discharges
from coastal wetlands in eastern Australia. Wetl. Ecol. Manag. 5 (1), 55–72.
Wilkin, R.T., Ford, R.G., 2006. Arsenic solid-phase partitioning in reducing sediments
of a contaminated wetland. Chem. Geol. 228 (1–3), 156–174.
Williams, A.G., Scherer, M.M., 2004. Spectroscopic evidence for Fe (II)− Fe (III)
electron transfer at the iron oxide− water interface. Environ. Sci. Technol. 38 (18),
4782–4790.
Williamson, M.A., Rimstidt, J.D., 1994. The kinetics and electrochemical rate-deter-
mining step of aqueous pyrite oxidation. Geochem. Cosmochim. Acta 58 (24),
5443–5454.
Wilson, B.P., White, I., Melville, M.D., 1999. Floodplain hydrology, acid discharge
and change in water quality associated with a drained acid sulfate soil. Mar.
Freshw. Res. 50 (2), 149–157.
Wilson, S.C., Lockwood, P.V., Ashley, P.M., Tighe, M., 2010. The chemistry and be-
haviour of antimony in the soil environment with comparisons to arsenic: a critical
review. Environ. Pollut. 158 (5), 1169–1181.
Yee, N., Shaw, S., Benning, L.G., Nguyen, T.H., 2006. The rate of ferrihydrite trans-
formation to goethite via the Fe(II) pathway. Am. Mineral. 91 (1), 92–96.
, .
Zachara, J.M., Kukkadapu, R.K., Fredrickson, J.K., Gorby, Y.A., Smith, S.C., 2002.
Biomineralization of poorly crystalline Fe(III) oxides by dissimilatory metal re-
ducing bacteria (DMRB). Geomicrobiol. J. 19 (2), 179–207.
Zhang, S.-L., et al., 2016. Sulfidization of As(V)-containing schwertmannite and its
impact on arsenic mobilization. Chem. Geol. 420, 270–279.
