Introduction
Igneous rocks, as granite, have low As concentrations
(< 5 mg kg-1), and background levels in soils are between 5 and
10 mg kg-1 (Smedley and Kinniburgh, 2002), although As levels are much
higher in certain polluted soils. As pollution can be very relevant in mine
sites where oxidation of sulfides such as pyrite takes place and in
areas treated with certain biocides and fertilizers (Matschullat, 2000). As
is an element that can accumulate in living beings and may cause severe
affectations, especially when it is in inorganic form (Smith et al., 2000;
Ghimire et al., 2003), with the potential to provoke environmental and
public health issues. In fact, the recommended threshold level for As in
drinking water is 10 µg L-1 (WHO, 2011).
When As-based products are spread on soils or spoils with the aim of
fertilizing, controlling plagues or promoting re-vegetation, risks of soil
and water pollution, and subsequent transfer to the food chain, must be taken
into account. As indicated in previous works, the use of wood preservative
compounds including arsenic or of As-based herbicides could cause arsenic
pollution episodes in forest areas (Smith et al., 1998) and cultivation soils
(Gur et al., 1979), in both cases increasing risks of soil and water
pollution (Clothier et al., 2006). In this way, it is interesting to
determine As retention capacity corresponding to solid substrates receiving
the spreading of the pollutant, both individually or treated with
complementary materials that can affect As retention/release potential. In
this regard, some previous works have investigated the effectiveness of
mussel shell waste amendment to increase As retention on diverse solid
materials (Seco-Reigosa et al., 2013a, b; Osorio-López et al., 2014), and
this amendment could also be useful to increase As retention on granitic
substrates (such as mine spoils or exposed C horizons), which has not been
studied up to now.
As concentration in natural waters is mainly controlled by interactions
between solids and solution, as adsorption/desorption, which are affected by
pH and other environmental parameters. Clays, organic matter and Fe, Al and
Mn oxyhydroxides can protonate or deprotonate as a function of pH,
facilitating retention of anions such as arsenate when they are positively
charged and promoting progressive anions release when pH rises and
surface charge becomes increasingly negative (Smith et al., 1999; Fitz and
Wenzel, 2002); however, at high pH values and in the presence of sulfate and
carbonate, co-precipitation of As with oxyhydroxides and sulfates, or even
as calcium arsenate, may occur (García et al., 2009). This could explain
that certain soils show maximum As adsorption at pH near 10.5 (Goldberg and
Glaubig, 1988). In this way, Zhang and Selim (2008) indicate that carbonate
can play an important role in arsenate retention in solid substrates having
high pH value. In fact, calcite has been related to As retention in
calcareous soils and carbonate-rich environments due to
adsorption/precipitation of CaCO3 and As forming inner sphere complexes
(Alexandratos et al., 2007; Mehmood et al., 2009; Yolcubal and Akyol, 2008;
Zhang and Selim, 2008), which could be relevant in granitic materials that
were amended with mussel shell to promote As retention.
The study of risks of soil and water As pollution, and the investigation of
potential means to diminish it are just a part of global concerns affecting
soil (and, subsequently, other environmental compartments). In the last
years, numerous studies have indicated that restoration needs to recover soil
functionality, and this call is taking place all over the world (Ahmad et
al., 2013; Johnston et al., 2013; Mao et al., 2014; Moreno et al., 2014;
Novara et al., 2014; Roy and McDonald, 2015; Sacristán et al., 2015;
Sadeghi et al., 2015; Srivastava et al., 2014). Some authors indicate that
this task should be accomplished with a broad view (Brevik et al., 2015) by
considering how soils can interfere with human health (Brevik and Sauer,
2015).
In view of that, the objectives of this work are (a) to determine As(V)
retention/release capacity corresponding to a granitic material, fine mussel
shell and coarse mussel shell, as well as to the granitic material amended
with 12 or 24 t ha-1 fine mussel shell, for different As(V)
concentrations and pH values; (b) to examine fitting of adsorption data to
the Langmuir and Freundlich models; and (c) to determine the fractions where
the adsorbed As(V) was retained, which is in relation with stability of
retention. As far as we know, no equivalent studies were made previously
with the combination of materials here used.
Results and discussion
Characterization of the solid materials
Table 1 shows that the granitic material had low C and N percentages
(indicating low organic matter content) and acid pH (5.7), whereas pH was
alkaline for fine and coarse mussel shell (9.4 and 9.1, respectively). Total
Ca and Na contents were higher for fine and coarse mussel shell, whereas the
granitic material presented the lowest effective eCEC
(eCEC < 4 cmol kg-1) as well as high Al saturation (64.5 %)
and total Al concentrations. Regarding Al forms, amorphous Alo
compounds were clearly more abundant in the granitic material, whereas those
bound to organic matter (Alp) had low presence in all of the
studied materials, with most of the amorphous Al being in inorganic form
(Alop). Similarly, the low organic-C content of the granitic
material and coarse and fine mussel shells justified that most Fe was bound
to inorganic forms (Feop). Additionally to that shown in Table 1,
the particle size distribution of the granitic material was 60 % sand,
23 % clay and 17 % silt.
Adsorption/desorption as a function of added As(V)
concentration
Figure 1a shows that As(V) adsorption was equivalent on granitic material and
fine mussel shell and higher than on coarse mussel shell. The different
behavior for both mussel shell materials (higher As adsorption on fine than
on coarse mussel shell) can be in relation with the higher surface area of
fine shell (1.4 m2 g-1) than that of coarse shell (1 m2
g-1), as previously stated by Peña-Rodríguez et al. (2013).
Figure 1b indicates that As(V) adsorption increased when granitic material
was amended with mussel shell. Adsorption curves in Fig. 1 show type C layout
(Giles et al., 1960) for granitic material and fine and coarse mussel shell
(Fig. 1a), exhibiting a rather constant slope when the added arsenic
concentration was increased. This kind of adsorption curve is generally
associated with the existence of a constant partition between the adsorbent
surface and the equilibrium solution in the contacting layer or to a
proportional increase of the adsorbent surface taking place when the amount
of adsorbed arsenic increases, as indicated by Seco-Reigosa et al. (2013b),
who found the same type of adsorption curve studying arsenic retention on
pine sawdust and on fine mussel shell. The granitic material treated with
mussel shell shows adsorption curves that are near C type (Fig. 1b).
Figure 2 shows that percentage adsorption progressively decreased on granitic
material when the As(V) concentration added was > 10 mg L-1. The
24 t ha-1 mussel shell amendment caused slightly increase in
percentage adsorption, whereas the 12 t ha-1 amendment did not result
in systematic increased percentage adsorption.
Fitting of the adsorption results to the Freundlich and Langmuir
models.
Freundlich
Langmuir
KF
n
R2
KL
Xm
R2
(Ln kg-1 mmol(1-n))
(dimensionless)
(L mmol-1)
(mmol kg-1)
Fine shell
10.8 ± 0.8
0.86 ± 0.08
0.987
–
–
Coarse shell
38.7 ± 11.4
3.14 ± 0.55
0.991
–
–
GM
9.0 ± 0.5
0.68 ± 0.06
0.991
1.0 ± 0.6
16.7 ± 6.0
0.978
GM+12 t ha-1
7.7 ± 0.9
0.41 ± 0.09
0.938
9.2 ± 8.0
6.9 ± 1.6
0.866
GM+24 t ha-1
10.8 ± 1.0
0.61 ± 0.08
0.977
1.6 ± 1.3
16.1 ± 7.5
0.951
GM: granitic material; 12 and 24 t ha-1: doses of the fine mussel
shell amendments; - fitting was not possible due to estimation errors being
too high.
Regarding desorption, Table 2 shows released As(V) concentrations and
percentages (referred to the amounts previously adsorbed). The highest
desorption percentage (49 %) corresponded to coarse mussel shell when
25 mg L-1 As(V) were added. When 100 mg L-1 As(V) were added,
percentage desorption was always < 19 %. Mussel shell amendment (12
and 24 t ha-1) increased As(V) desorption, which could be in relation
with the fact that arsenate bind strongly to the surface of oxides and
hydroxides in clearly acid environments (pH between 3.5 and 5.5; Silva et
al., 2010), whereas increased pH values (from above 5 for clay minerals to
above 12 for calcite) favor desorption (Golberg and Glaubig 1988). Any case,
most of the adsorbed As(V) did not desorb, indicating notable irreversibility
of the process.
Relationship between added As(V) (mg L-1) and As(V)
percentage adsorption for the unamended and shell-amended (12 or 24 t ha-1) granitic material. Average values for three replicates, with
coefficients of variation always < 5 %.
(a) Time-course evolution of pH for the solid materials as a
function of the various molar concentrations of added HNO3 and NaOH; (b)
relationship between adsorption (mg kg-1) and pH value for fine shell
and the unamended and shell-amended granitic material. Average values for
three
replicates, with coefficients of variation always < 5 %.
Adsorption data were adjusted to the Freundlich and Langmuir models
(Table 3), finding that the unamended and shell-amended granitic material
fitted well to both models, whereas fine and coarse mussel shell can be
fitted only to the Freundlich model. Maji et al. (2007) found satisfactory
adjustment to both Freundlich and Langmuir models studying As(V) adsorption
on lateritic substrates, while Yolcubal and Akyol (2008) obtained better
fitting to the Freundlich model using carbonate-rich solid substrates.
As(V) adsorption/desorption as a function of pH
Adsorption
Figure 3 shows the repercussion on As(V) adsorption of adding different
HNO3 and NaOH molar concentrations to fine mussel shell and to the
unamended and shell-amended granitic material. The acid concentrations added
to fine shell were not permitted to reach pH < 7 (Fig. 3a), whereas the
addition of alkaline solutions was allowed to achieve pH values near 12 for this
material. The granitic material exhibited the lowest buffer potential
(possibly related to its low colloids content), presenting pH values between
2 and 10. Mussel shell amendment increased the buffer potential of this
granitic material, especially when the 24 t ha-1 dose was used.
Figure 3b shows that As(V) adsorption on the granitic material (expressed in
mg kg-1) progressively decreased from pH 4 as a function of increasing
pH value, whereas the mussel shell amendment increased As(V) adsorption. The
granitic material contains variable charge compounds (such as Fe and Al
oxyhydroxides, kaolinite-type clays and organic matter), positively charged
at acid pH, facilitating retention of H2AsO4- and
HAsO42- (Smedley and Kinniburgh, 2002; Xu et al., 2002; Yan et al.,
2000) but suffering progressive de-protonation and increase of negative
charge as pH increases, which can lower As(V) adsorption (Fitz and Wenzel,
2002). However, the effect of lowering As(V) adsorption due to pH increase
did not occur when granitic material was amended with mussel shell, which
must be related to the additional As(V) adsorption capacity associated
with calcium carbonate present in mussel shell, establishing cationic bridges
when pH values are higher (Alexandratos et al., 2007). Salameh et al. (2015)
found that arsenic was completely removed by charred dolomite samples
(another alkaline material) over a wide range of pH (2–11). Our granitic
material suffered just slight changes in As(V) adsorption in the pH range 3.5
to 6.9, which can be related to the effective adsorption that As(V)
experience in a wide range (4–11) (Stanic et al., 2009).
Expressing As(V) adsorption as percentage with respect to the amount added,
the maximum for the unamended granitic material (66 %) took place at
pH < 6, progressively decreasing from that point as a function of
increasing pH value. Fine mussel shell adsorbed As(V) notably on the pH range
6–12, with maximum value of 83 %. When the granitic material was amended
with fine mussel shell, As(V) adsorption reached 99 % at pH near 8 and then
progressively decreased as pH increased.
In the case of the shell-amended granitic material, significant
(p < 0.005) statistical correlations existed between adsorbed As(V) and
pH (r=0.926 and r=0.880 for the 12 and 24 t ha-1 mussel shell
doses, respectively), whereas no correlation was found between both
parameters in the case of mussel shell by itself. The latter can be due to
the absence of anionic exchange with OH- groups when As(V) anions adsorb on
mussel shell, contrary to that happening to other anions on different
adsorbent materials (Arnesen and Krogstad, 1998; Bower and Hatcher, 1967;
Gago et al., 2012; Huang and Jackson, 1965). However, anions other than
OH- can be released, as is the case for SO42-, PO43- or
organic anions, which is in concordance with the correlations found between
adsorbed As(V) and DOC (r = 0.810, for fine shell, and r = 0.919
and r = 0.913, for the granitic material amended with 12 and
24 t ha-1 mussel shell, respectively, p< 0.005). Moreover, other
mechanisms that can be responsible for anion retention (such as retention on
calcite or H and van der Waals bindings) do not implicate OH- release
(Boddu et al., 2003). Different authors remark on the influence of pH on As(V)
adsorption (Maji et al., 2007; Partey et al., 2008; Stanic et al., 2009), but
in the case of our granitic material, Al, Fe, Alo,
Feo, organic matter and organoaluminum complexes, contents must
also be relevant.
Fine and coarse mussel shell presented alkaline pH (9.39 and 9.11,
respectively, Table 1), making the dominant As species
HAsO42- (Yan et al., 2000), which can bind to the surface of
carbonates such as calcite by means of inner sphere complexes with octahedral
Ca (Alexandratos et al., 2007).
Relationship between As(V) desorption (%) and pH value for
fine shell and for the granitic material (average values for three replicates,
with coefficients of variation always < 5 %) when 100 mg L-1
As(V) were added to the adsorbents.
Desorption
Figure 4 shows that, when a concentration of 100 mg L-1 As(V) was
added, As(V) desorption from fine shell and granitic material was always
< 26 % of the amount previously adsorbed, considering the whole
pH range studied (2–12). Two different behaviors took place: (a) As(V)
desorption from granitic material clearly increased as pH increased between 4
and 6, and (b) As(V) desorption from mussel shell clearly decreased as pH
increased between 4 and 6. Moreover, As(V) desorption from mussel shell
continued to be low at pH > 6, slowly decreasing, whereas release from
the granitic material further increased when pH > 6.
As(V) desorption from mussel shell clearly increased at pH < 6 in
accordance with that detected by Goldberg and Glaubig (1988), who found that
As adsorption on calcite increased from pH 6 to 10 (then decreasing release),
attaining maximum adsorption at pH between 10 and 12 and then decreasing at
higher pH values. Di Benedetto et al. (2006) indicated that As(V) can be
incorporated to calcite in alkaline conditions by preventing its
mobilization even in situations where oxyhydroxides do not exhibit
adsorption potential. Alexandratos et al. (2007) found that arsenate
anions have great affinity for calcite at pH around 8, establishing strong
bindings due to inner sphere complexes with AsO43- binding to the
mineral surface through Ca cationic bridges. All these facts are in
accordance with the low As(V) release suffered by our mussel shell samples at
pH > 6 (Fig. 4).
Percentages of the various fractions of As(V) adsorbed after 24 h
(a), 1 week (b) and 1 month (c) of incubation. Average values for
three
replicates, with coefficients of variation always < 5 %.
Fractionation of the As(V) adsorbed at three different
incubation times
Figure 5 shows that the As(V) soluble fraction (exchangeable and bound to
carbonates) is quantitatively the most important in all samples (especially
in the unamended and shell-amended granitic material), representing at 24 h
of incubation contents that ranged between a minimum of 69 % in fine
mussel shell and a maximum of 88 % in the 12 t ha-1 shell-amended
granitic material. The soluble fraction corresponds to the most mobile As(V),
which is weakly retained mainly due to anionic exchange mechanisms (Keon et
al., 2001) and which is associated to high risks of toxicity. Moreover,
Taggart et al. (2004) indicate that As(V) derived from anthropogenic
pollution incorporates to the most mobile fractions of solid substrates in
great percentage. In our materials, the As(V) reducible fraction (associated
to Al and Fe oxides and oxyhydroxides) represented between 9 and
19 % of the As(V) adsorbed at 24 h of incubation (Fig. 5), whereas the
As(V) residual fraction (that incorporated to the structure of minerals) always
constituted < 16 % of the amount adsorbed. Finally, the As(V)
oxidizable fraction (associated to organic matter and as sulfides) was always
< 2.6 % (Fig. 5), attributable to the low organic content of the solid
materials here studied. The increase of incubation time from 24 h to 1 week
and to 1 month, as well as the 12 t ha-1 shell amendment of the
granitic material, did not cause statistically significant modifications in
the percentage content of each fraction of the adsorbed As(V) (Fig. 5).
The As(V) reducible fraction (bound to Al and Fe oxides and oxyhydroxides)
correlated positively with DOC (r=0.957 at 24 h, and r=0.954 at 1 week
incubation time, p< 0.005), suggesting that arsenate compete with organic
groups to bind on oxides and oxyhydroxides. Additionally, the As(V) residual
fraction correlated with total Fe (r=0.980 at 24 h, and r=0.973 at 1
month incubation time, p< 0.005), suggesting the existence of
re-adsorption and co-precipitation processes with Fe minerals.