The Science of the Total Environment 301 (2003) 239–250
Solubility of metals in an anoxic sediment during prolonged
aeration
Nathalie Caillea,b,c,*, Christophe Tiffreaua, Corinne Leyvalc, Jean Louis Morelb
a
CNRSSP, 930 Bd Lahure, BP 537, F-59505 Douai Cedex, France
ˆ de Haye, F-54505 Vandoeuvre-les-Nancy,
`
Laboratoire Sols et Environment ENSAIA-INRAyINPL, 2, avenue de la Foret
France
c
´
` organique dans les sols- CNRS,
Laboratoire des Interactions Microorganismes-Mineraux-Matiere
`
17 rue Notre-Dame des Pauvres, F-54501 Vandoeuvre-les-Nancy,
France
b
Received 18 December 2001; accepted 8 July 2002
Abstract
This work was conducted to study the evolution of the solubility of selected metals during the aeration of an
anoxic sediment. Batch experiments were carried out for 76 days with a metal-polluted dredged sediment. The pH,
Eh and concentration of Al, Cu, Fe, Hg, Pb and Zn were periodically recorded. Results showed that during the early
stages of aeration, the solubility of metals increased rapidly but was then followed by fast re-adsorption. As a
consequence, after 14 days most of the metals excepted Cu and Zn were present at low or undetectable concentrations
in solution. Re-adsorption of Zn was observed to be much slower during the first two weeks, whereas solubilisation
of Cu increased gradually during months after land disposal. According to speciation calculations, Cu solubilisation
was in part due to complexation in solution by carbonates. In the case of Hg, although complexation by dissolved
organic matter (DOM) could be expected, re-adsorption was the dominant process. However, more knowledge about
the behaviour of the DOM present in anoxic sediments is needed in order to make more quantitative statements about
the mobility of heavy metals contained in dredged material.
䊚 2002 Elsevier Science B.V. All rights reserved.
Keywords: Metals; Mercury; Dredged sediments; Aeration; Mobilisation; MINTEQA2
1. Introduction
Waterways are subjected to a constant silting up
as a result of soil erosion. In order to prevent the
risk of flooding, and allow navigation, the accumulated sediments have to be dredged periodically.
Due to the large amount of materials generated
*Corresponding author. Tel.: q33-3-83-59-57-75; fax: q
33-3-83-57-57-91.
E-mail address: [email protected] (N. Caille).
and the cost of decontamination, dredged products
are generally disposed of along streams or in
deposit sites, without any further treatment. However, sediments may contain large amounts of
metallic pollutants, e.g. trace elements such as: Zn,
Pb, Cd, Cu or Hg.
In anoxic sediments, the partitioning of these
elements is generally dominated by the formation
of metal sulfides, i.e. ZnS (sphalerite), PbS (galena), FeS2 (pyrite), FeS2 (marcasite), FeS (trochite,
machinawite), and amorphous FeS (Brennan and
0048-9697/03/$ - see front matter 䊚 2002 Elsevier Science B.V. All rights reserved.
PII: S 0 0 4 8 - 9 6 9 7 Ž 0 2 . 0 0 2 8 9 - 9
240
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
Lindsay, 1996). Mercury and copper, may also be
present in important concentration in the structure
of pyrite (Huerta-Diaz et al., 1998; Morse and
Luther, 1999). Besides, large quantities of organic
matter do generally accumulate in anoxic sediments, due to the fact that microbial activity is
reduced in oxygen-poor environments (Kersten,
1989). These natural organic compounds are also
frequently considered as sinks for metallic
pollutants.
As a consequence, disposal of dredged material
is likely to change the redox potential, pH and the
organic matter decay rate, which are the most
important factors controlling pollutant mobility
¨
(Forstner
and Kersten, 1989; Kelley and Thuovinen, 1989; Calmano et al., 1996). In generals,
experiments show an enhancement of trace elements availability during aeration of anoxic sediments (Lee and Touray, 1997; De Carvalho et al.,
1998; Alastuey et al., 1999; Tack et al., 1999; Wen
and Allen, 1999; Marseille et al., 2000; Singh et
al., 2000).
Nevertheless, if numerous studies deal with the
leaching of metal from a dredged sediment, very
few take into account mercury with other elements.
Besides, mobility of Hg is generally studied during
short-term suspension. Because of the slow oxidation rate of HgS (Biester and Scholtz, 1997;
Cooper and Morse, 1998) and the strong complexing capacity of organic matter or clays contained
¨
in sediments (Wallschlager
et al., 1996; Desauziers
et al., 1997; Celis et al., 2000) long term experiments are required.
Moreover, the major problem is also that most
of the studies dealing with the potential metallic
pollution emanating from dredged sediments rely
on empirical results (extraction schemes, resuspension experiments, etc.) and on qualitative considerations. Although some simple modelling
approaches have been proposed to model the
release of basic organic compounds from dredged
material (De Rooij and Gerrits, 1995; Blackburn,
1997; Schroeder and Aziz, 1999), very few
attempts have been made to quantitatively model
the potential release of heavy metals in disturbed
dredged material.
Consequently, this work was undertaken to
determine the solubility of mercury during a pro-
longed aeration and to compare the evolution of
Hg in solution with Cu, Pb and Zn. The results
obtained are intended to be used for a multielement approach in disposal management. With
this aim, a sediment highly contaminated with Hg
and moderately contaminated with Cu, Pb and Zn
was collected from a river stream, suspended in
deionised water, and metals concentrations in solution were recorded. This experimental study was
combined with speciation calculations in order to
bring some more quantitative arguments about the
factors affecting the mobility of heavy metals
contained in dredged sediments.
2. Materials and methods
2.1. Sediment characterisation
Fresh sediment was dredged from the Scarpe
canal, at a sampling spot located in the city centre
of Douai, north of France (59 district). The sediment was excavated with a grab and poured in a
truck skip. The sediment was immediately thoroughly homogenised in the skip using the grab,
and was transferred to 60 l tanks for storage. River
water was added to the tanks to keep the surface
of the sediment under water and to maintain anoxic
conditions. The sediment was allowed to rest in
the tanks for 30 days before any experiment, in
order to allow the re-establishment of initial anoxic
conditions and to minimise the perturbation
induced by dredging. After this period of time, the
return to anoxic conditions was assessed from Eh
measurements. The redox potential of the sediment
was measured in situ by inserting a redox combination electrode (Pt and AgyAgCl, Bioblock)
directly in the sediment and waiting for a stable
reading, according to the method of Tack et al.
(1996). The initial pH of the sediment was measured in a 1:5 solidyliquid suspension using a KCl
combination electrode (Bioblock). This suspension
was prepared with fresh sediment and de-aerated
water inside a N2 saturated glove box and the pH
was measured inside the glove box after a 6-h
stirring period.
All other analyses were conducted on material
dried at 25 8C under a laminar hood. The CEC
was determined with the standardised cobaltihex-
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
241
Table 1
Metal content of the certified reference sediments measured
CRM 320
CRM 144
Certified values
Mean values measured
Certified values
Mean values measured
Cu
Hg
Pb
Zn
44.1"1.0
42.9"1.3
–
–
–
–
1.49"0.06
1.42"0.09
42.3"1.6
40.7"1.2
–
–
142"3
138"4
–
–
amine method NFX 31-130 (AFNOR, 1999,
2001). Total carbon and inorganic carbon were
measured with a total organic carbon analyser
(TOC y5000 A Touzard and Matignon) equipped
with a solid sample module (SSM). The organic
matter fraction was estimated from organic carbon
content (1.72=C) according to Allison (1965).
With the exception of mercury, metal concentrations in the sediment were determined by plasma
atomic emission spectrometry (ICP-AES JobinYvon 138 Ultrace) after acid digestion. For this
purpose, 0.5 g of dry sediment was calcinated at
450 8C for 4 h. This step was performed in order
to eliminate the organic matter, owing to the large
amount of organic carbon present in the sediment.
The residue was allowed to react with 10 ml HF
(50%) and 3 ml HClO4 (70%) during 12 h.
Afterwards, the solution was heated at 150 8C
until an almost complete evaporation of the acids.
The remaining ‘cake’ was dissolved in 2 ml HCl
(37%) and 5 ml of water. This solution was filtered
with Whatman 2 V filter (8 mm), and the volume
was adjusted to 100 ml.
Analysis of total Hg was performed directly on
25 8C dried sediment, without the calcination step,
to avoid mercury volatilisation. For this purpose,
5 ml KMnO4 (50 g ly1), 5 ml HNO3 (65 %), and
5 ml H2SO4 (96%) were added to 0.5 g of dried
sediment. After 12 h of stirring, the volume was
adjusted to 50 ml with water, the solution was
centrifuged (5500 rpm for 30 min) and filtered
with Whatman 2 V filter (8 mm).
Mercury concentration in the interstitial water
of the sediment was measured on the solution
collected after centrifugation of fresh anoxic sediment for 30 min at 10 000 rpm and filtration at
0.45 mm on syringe filters.
Before mercury analysis, 200 ml of 15 g ly1
NH2OH, HCl solution were added to all the liquid
samples (acid digestions, interstitial water). Mercury concentration in these samples was analysed
by cold vapour atomic absorption spectrometry
(CVAA) after reduction of Hg2q in HgO by
SnCl2 (Varian SpectrAA 220 spectrometer
equipped with a VGA 77 cold vapour kit).
Both spectrometers (ICP-AES and CVAA) were
calibrated with external standards prepared by
dilutions of 1000 mgØly1 reference solutions (ICP
Multielement Standard IV, Merck and Hg standard
solution, Merck) with the same acid mixture used
for sample dissolution. Moreover, the two digestion methods were tested against two certified
sediments (CRM 320 and CRM 144) supplied by
The Community Bureau of Reference Sample
(BRC). The experimental data obtained were compared with certified values, showing that the mean
measured valued differed from less than 5% from
the certified values (Table 1).
Further analysis were performed to assess the
initial speciation of mercury in the sediment. These
analysis included the determination of elemental
mercury (Hg0), monomethyl mercury (MeHgq)
and dimethyl mercury (Me2Hg). These measurements were performed by the LCABIE (University
of Pau, France), using gas chromatography and
inductively coupled plasma mass spectrometer
(ICPyMS) (Amouroux et al., 1998).
X-Ray Diffraction spectra of sediment samples
(XRD) dried at room temperature were also
recorded at the Douai School of Mines (ENSM)
in order to identify the mineralogical major phases.
Mineral identifications was performed using a
Siemens D501 powder diffractometer with Co Ka
radiation (40 KV and 37.5 mA).
2.2. Batch experiments
The preparation of the batch samples was performed in a N2 saturated glove box to prevent any
242
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
oxidation. Ten grams of the wet anoxic sediment
were mixed with 150 ml of deionised and deaerated water in glass flasks, so that the resulting
sediment concentration was 23.3 g of dry material
per liter. The flasks were covered with a foam cap
allowing gas exchange and avoiding dust contamination. After preparation, the batch samples were
withdrawn from the glove box and placed on a
rotating plate at room atmosphere and temperature
(between 19 and 21 8C), and submitted to a
constant stirring (300 rpm). The stirring speed was
adjusted to create a vortex during agitation and to
ensure that the leaching solution was saturated
with atmospheric oxygen. Deionised water was
added periodically to maintain the same volume
of water during the experiment. Twenty-one flasks
were prepared. After 1, 7, 14, 28, 42, 56 and 77
days, three flasks were collected, the suspensions
were centrifuged for 30 min at 12 300 rpm and Eh
and pH were measured on the supernatant. The
solutions were filtered through 0.45 mm sterile
Analypor MC (cellulose esters) syringe filters, and
metals were immediately analysed.
Prior to Hg analysis, 200 ml of an acid solution
of K2Cr2O7 (5 g of K2Cr2O7 and 500 ml HNO3
adjusted to 1 l) and 20 ml of a solution composed
of 5 % (vyv) KBr, 5% (vyv) KBrO3 and 30%
(vyv) HCl were added to the supernatant (XP T
90-113-2 AFNOR method, 1997). Mercury was
then analysed by CVAA spectrometry. To evaluate
the abundance of Hg bound to sediment particles,
Hg was also quantified in 20 ml of non-filtered
supernatant.
Other trace metals in the supernatant were analysed after acidification to 2% HCl; Al, Ca, Fe,
Mn, Ti and Zn were measured with ICP-AES, and
Cu and Pb were analysed with a graphite furnace
atomic absorption spectrometer (SpectrAA 220
Varian).
Dissolved inorganic carbon (DIC) and dissolved
organic carbon (DOC) were measured using the
TOC-5000A. Before analysis, 200 ml of a NaN3
solution (30 g ly1) were added to 15 ml of the
supernatant to inhibit microbial development and
organic matter degradation. Sulfate, phosphate and
chloride ions were analysed by ion chromatography (Dionex DX 500 equipped with an Ionpac
AG4A-SC Guard pre-column PyN 43175 4 mm,
and an Ionpac AS4A-SC Analytical column PyN
43174 4 mm). Data presented here correspond to
the mean values and standard error of the three
replicates.
2.3. Modelling
The MINTEQA2yPRODEFA2 version 4.02
chemical equilibrium model (Allison and Brown,
1991, 1999) was used to perform speciation calculations. These calculations were solely intended
to estimate ion speciation in the solution of the
batch samples. Hence, computations were restricted to the solution compartment of the system, and
it was assumed that the solution species were at
equilibrium. This is conditions can be achieved if
it is considered that reactions in solution (complexation, protonation, etc.) proceed usually much
faster than solid solution exchanges (dissolution,
precipitation, etc«). Therefore, the parameters
(pH, DOC, DIC, Ca, Mg, Al, Cu, Hg, Pb, Zn, Fe,
y
3y
SO2y
and NO3y concentrations) meas4 , Cl , PO4
ured in the solution of the batch samples were
directly used as input data and oversaturated solids
were not allowed to precipitate. Due to the high
values of Eh measured on the leachates (see
discussion below), the measured elements were
assumed to be in their oxidised state and the redox
potential was not explicitly taken into account in
the calculations.
Due to the fact that non-negligible quantities of
DOC ()6 mgØly1) were measured in the batch
solutions, it was attempted to account for the
presence of soluble organic matter in the system.
Several models have been developed to describe
the ligand behaviour of natural organic matter in
solution (Mantoura and Riley, 1975; Buffle et al.,
1977; Perdue and Lytle, 1983; Cabaniss and Shuman, 1988; Dobbs et al., 1989a,b; Kinniburgh et
al., 1996). Most of these models can be successfully used to describe the behaviour of organic
substances originating from soils, surface waters,
wetlands but applications of such models to organic compounds present in dredged sediments porewater are practically non existent. More generally,
very few studies have focused on the exact nature
and complexing capacities of the dissolved organic
matter (DOM) contained in anoxic river sedi-
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
243
Table 2
Physical and chemical properties of the studied sediment
Sediment main characteristics
Initial pH-H2O (1:5 solid:liquid)
Initial redox potential (mV vs. ENH)
Organic matter (%)
Carbonate content (% CaCO3)
Cation exchange capacity (cmolØkgy1 dry matter)
7.6"0.2
y110"12
10"0.8
17"0.9
14.2"0.6
Particle size fraction
- 2 mm (%)
2–20 mm
20–50 mm
) 50 mm
12.9
54.3
25.2
7.6
Total contents (mgØkgy1 dry matter)
Ca
Fe
Al
Mn
Cu
Zn
Pb
Cd
total Hg
Hg0
MeHgq
Me2Hg
Mineralogical major phases (XRD)
Quartz (SiO2), Calcite (CaCO3), Albite (NaSi3AlO8), Kaolinite (in-2mm fraction)
ments. However, calculations performed for this
study are solely intended to support hypothesis
emerging from experimental data. As a consequence, a simple approximation was made by
assimilating the DOC present in the batch samples
to humic substances, since these compounds have
been reported to constitute the main fraction of
the non-volatile DOM contained in anoxic sediment porewater (Thurman, 1985; Abbt-Braun and
Frimmel, 1996). Complexation by humic substances was computed using the Gaussian distribution model proposed by Dobbs et al. (1989a,b).
This model is already implemented in MINTEQA2, in combination with a database originating from Susetyo et al. (1991) concerning a wide
range of cations (Hq, Al3q, Ba2q, Be2q, Ca2q,
Cd2q, Cr3q, Cu2q, Fe3q, Mg2q, Ni2q, Pb2q and
Zn2q). For Hg2q, the constant proposed by Yin et
al. (1997) was used, since this constant has been
determined with the same model as Susetyo et al.
(1991).
76 000"2500
13 200"1100
34 500"3600
300"27
165"7
860"22
300"18
2.2"0.1
20"0.2
27.10y6"2.10y6
19.10y6"1.10y6
34.10y6"3.10y6
3. Results
3.1. Main characteristics of the sediment
The anoxic sediment presented a neutral pH, a
high buffering capacity, and an important percentage of organic matter (Table 2). It was mainly
composed of silts (78.2%) and the major mineralogical phases were quartz, calcite, albite and
kaolinite. These characteristics are in close agreements with the ones determined by Isaure et al.
(2002) for sediments originating from other sampling spots on the same canal. Mercury was by far
the main contaminant present in our sediment. Its
concentration was 200-fold more elevated than its
local background concentration in sediments
whereas the concentrations of Cu, Pb and Zn
corresponded to 8–15-fold their local background
´
concentrations (Becart
et al., 1997). The sediment
was strongly reduced, as shown by the redox
potential of y110 mV. Due to the anoxic nature
244
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
Fig. 1. Eh (mV) of the solution during the prolonged aeration.
of the sediment, a significant occurrence of sulfide
can be expected. This is confirmed by Isaure et al.
(2001, 2002) who detected the presence of pyrite
(FeS2), sphalerite (ZnS) and galena (PbS) in
sediments from the Scarpe canal.
Concerning mercury speciation, the analysis performed show that this element was mainly present
under the form of Hg2q compounds, the fractions
of Hg0, MeHgq and Me2Hg being negligible
compared to the total Hg content.
3.2. Evolution of the sediment during the batch
experiment
The Eh values showed a rapid increase, from
y110 to q500 mV after 7 days (Fig. 1), then a
slight decrease to q420 mV and remained relatively stable during the rest of the experiment. Due
to the complex nature of the solution and the
presence of different redox couples, the Eh measured is probably a mixed potential rather than a
true equilibrium potential. However, it indicates
that the stirring of the batch samples was vigorous
enough to maintain highly oxidative conditions
during the leaching experiments. The pH increased
Fig. 2. pH of the solution during the prolonged aeration.
Fig. 3. Dissolved organic carbon concentration (DIC), and dissolved organic carbon concentration (DOC) in solution in
mgØly1.
slightly after one day, then decreased at day 7, and
increased again to remain constant at an average
value of 7.5 (Fig. 2). The DIC concentration
decreased from 16 to 4 mg ly1 with large oscillations (Fig. 3), whereas the DOC concentration
showed a smoother decrease during the whole
experiment, starting at 10 mg ly1 and ending at 6
mg ly1.
Concerning the major anions, the concentration
in SO2y
and Ca2q increased rapidly during the
4
first 28 days and decreased slightly afterwards
(Fig. 4), the maximum values being 130 mg
SO2y
ly1 and 100 mg Ca2q ly1. The same trend
4
was observed for the phosphate concentration,
which increased slightly from 1.9 to 2.4 mg ly1
(Fig. 5).
Concerning chloride, the concentration evolution can hardly be interpreted, since the concentrations are near to the detection limit of the method
used (0.5 mgyl) and are associated with large
uncertainties.
The major cations exhibited similar behaviours,
since 200 mg Al ly1 and 150 mg Fe ly1 were
detected in solution after 1 day of leaching (Fig.
6), and these concentrations decreased sharply to
Fig. 4. Sulfate and calcium concentrations in solution in
mgØly1.
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
245
Fig. 5. Chloride and phosphate concentrations in solution in
mg ly1.
Fig. 7. Zn concentration in solution in mgØly1.
became undetectable after 14 days. Quite similarly,
manganese was not detected in solution during the
whole experiment. Zinc showed a greater mobility
since its concentration in solution reached 100 mg
ly1 at day 14, and decreased slightly until the end
of the experiment (Fig. 7). Comparatively, mercury
concentration reached its highest value (0.7 mg
ly1) after 12 h of shaking and was undetectable
after 3 days (Fig. 8). For reference, analysis of
the filtrated interstitial water of the initial anoxic
sediment (ts0) revealed no soluble Hg. Similarly,
the Pb concentration was 4 mg ly1 after 12 h, and
decreased slowly to become undetectable after 14
days. The copper concentration evolved in a completely different manner than that of Hg, Pb or Zn.
It steadily increased during the whole assay to
reach 11 mg ly1 after 76 days.
in Table 3. According to these computations, aluminium and iron were present mainly as hydroxocomplexes and calcium as Ca2q (Table 3).
However, it must be noted that the solutions were
estimated to be largely over-saturated (log of
saturation indice)1) with respect to aluminium
and iron (hydr)oxides (boehmite, gibbsite, ferrihydrite, goethite and hematite) during the whole
experiment. Besides, the calculations indicated that
the system was nearly saturated with respect to
calcite. Concerning the metal speciation, mercury
was estimated to be complexed by hydroxyl and
dissolved organic matter (DOM) after 12 h of
shaking, before getting undetectable. According to
the computations, Cu2q and Pb2q exhibited similar
behaviours, being mainly present as carbonate
soluble complexes and as free ions during the
whole experiment. In contrast, Zn was calculated
to be present mainly as Zn2q. It is to note that for
Zn2q, Cu2q and Pb2q, complexation by organic
matter was calculated to be negligible compared
to carbonate complexation. However, due to the
lack of knowledge concerning the nature of the
3.3. Speciation of main elements estimated by the
MINTEQA2 chemical equilibria model
The distribution of the most relevant cations, as
calculated with the MINTEQA2 code, is presented
Fig. 6. Fe, and Al concentrations in solution in mgØly1.
Fig. 8. Hg, Pb, and Cu concentrations in solution in mgØly1.
246
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
Table 3
Percentage speciation (% of total soluble element) of the main species of Al, Ca, Fe, Hgy , Cu, Pby and Zn, as calculated by
MINTEQA2
Days of shaking
% of total soluble element
12 h
7 day
14 day
28 day
42 day
56 day
76 day
Al(OH)3 aq
Al(OH)y
4
1.0
98.9
7.8
86.9
–
–
–
–
–
–
–
–
–
–
Ca2q
CaHCOq
3
CaSO4 aq
94.5
1.1
2.5
90.2
–
8.8
89.0
–
9.9
88.6
–
10.4
87.9
–
10.6
88.4
–
10.8
89.8
–
9.3
Fe(OH)q
2
Fe(OH)3 aq
56.3
40.8
92.4
7.5
–
–
–
–
–
-
–
–
–
–
Hg(OH)2
HgDOM
56.2
43.2
–
–
–
–
–
-
–
–
–
–
–
–
Cu2q
CuOHq
CuCO3 aq
CuSO4 aq
CuDOM
4.0
7.6
82.0
3.8
3.9
38.8
7.9
38.3
3.8
9.8
23.1
9.4
60.0
2.6
3.5
30.9
12.4
44.9
3.6
7.0
7.7
12.5
74.7
–
1.6
37.7
17.0
30.3
4.6
9.3
37.7
11.7
39.7
3.9
5.7
Pb2q
PbOHq
PbCO3 aq
PbHCOq
3
PbSO4 aq
PbDOM
6.2
9.4
65.5
5.2
7.6
12.1
36.4
5.9
18.4
12.9
7.6
18.4
27.3
8.8
36.1
12.5
6.5
8.3
32.9
10.5
24.5
8.5
8.3
14.8
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
–
Zn2q
ZnOHq
ZnCO3
ZnSO4
ZnDOM
73.0
4.4
14.8
1.9
2.8
88.0
–
–
8.2
–
85.3
1.1
2.2
9.0
–
86.0
1.1
1.2
9.6
–
77.2
4.0
7.3
8.9
–
86.5
1.2
–
10.0
–
87.9
–
–
8.7
–
DOM: Dissolved Organic Matter.
DOM contained in the samples, some uncertainties
remain about this statement.
4. Discussion
In general, as assessed from different chemical
extractions schemes, the heavy metals present in
anoxic sediments are generally found to be associated with phases qualified as ‘residual’ or ‘oxid¨
able’ (Kersten and Forstner,
1986; Kersten, 1989;
Giani et al., 1994; Cauwenberg and Maes, 1997;
Yu et al., 2001). Based on these observations,
these metals are generally considered to be associated with sulfide and organic compounds. More
recently, the use of different spectroscopic methods
has enabled to directly evidence the occurrence of
heavy metal sulfides in anoxic river sediments
(Dood et al., 2000; Large et al., 2001; Isaure et
al., 2002). As a consequence, it is now generally
recognised that oxidation of dredged material can
lead to a dramatic modifications of the heavy
¨
metal mobility (Calmano et al., 1993; Forstner,
1995; Tack et al., 1996; Stephens et al., 2001).
As shown on Fig. 4, the prolonged aeration of
the sediments is characterised by the oxidation of
sulfide phases to sulfate, which are gradually
released in solution. Besides, this reaction is
accompanied by a proton production and to a
¨
acidification of the medium (Forstner,
1995).
However, due to the presence of carbonate minerals (mainly calcite) in our sediment, this acidification is buffered by the dissolution of these solid
phases, leading to a concomitant release of calcium
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
in solution (Fig. 4). This observation is supported
by the fact that the calculations performed with
MINTEQA2 indicate that the solution is always
near saturation with respect to CaCO3. The phosphate release observed in Fig. 5 can be attributed
to the oxidation of vivianite wFe3(PO4 )2 x, a mineral
frequently observed in anoxic sediments (Large et
al., 2001).
As a result of the oxidation of sulfide phases,
the associated heavy metals are rapidly released in
solution during the early stages of the aeration.
However, the solubilisation kinetics differ from
one metal to another, Hg, Pb and Cu being released
more rapidly than Zn (Figs. 7 and 8). This is
¨
consistent with the observations made by Forstner
(1995), who reported a similar lower release kinetic for Zn, compared to Pb and Cu. After the rapid
release stage, a decrease in solubility is observed
for Hg, Pb and Zn. This phenomenon is classically
observed in resuspensionyoxidation experiments
of anoxic sediments and is attributed to translocation (re-adsorption or coprecipitation) of heavy
metals onto more soluble solid phases, such as
carbonates, ferric hydroxides or clay minerals
¨
(Calmano et al., 1993; Forstner,
1995; Stephens et
al., 2001). This process has been confirmed by
Isaure et al. (2002) who studied the speciation of
Zn in a sediment dredged from the Scarpe canal
and submitted to meteorological weathering. These
authors have observed the occurrence of Zn-sorbed
ferrihydrite and Zn-containing phyllosilicate in the
oxidised sediment. The formation of the first Znbearing phase is attributed to the oxidation of
pyrite, leading to the release of ferric ion in
solution, which precipitates at pH values greater
than 6 to form ferrihydrite. This is mechanism is
very likely to occur in our case, as indicated by
the rapid decrease in Fe and Al concentrations in
leachate solutions (Fig. 6). Moreover, the speciation calculations confirm that the leachate solutions
are rapidly over-saturated with respect to ferric and
aluminium (hydr)oxides. Zinc-containing phyllosilicates are rather supposed to form by Zn–Si
precipitation, since Isaure et al. (2002) observed
that the porewater of the oxidised sediment was
oversaturated with respect to phyllosilicate. Since
soluble Si could not be measured during our
experiments (due to the use of glass flasks), this
247
latter phenomenon could not be pointed out in our
case. However, when comparing the leaching
behaviour of Zn, Pb and Hg, one has to note that
the re-adsorption of Zn proceeded more slowly
than for Hg and Pb. Again, this is consistent with
¨
observations made by Forstner
(1995) and is
attributed to the lower affinity of Zn for sediment
solid phases (Alloway, 1995; Lin and Chen, 1998;
Rybicka et al., 1995)
As stated above, the behaviour of copper was
significantly different, since it concentration in
solution increased steadily during the whole resuspension period. As a matter of fact, lower readsorption rates are generally observed for copper
during oxidation of dredged sediments (Maass and
¨
Miehlich, 1988; Forstner,
1995). This is frequently
attributed to the high affinity of copper to dissolved organic matter. In our case, the speciation
calculations rather indicate that the high solubility
of Cu is due to strong complexation by carbonate
ions (Table 3). However, due to the uncertainty
remaining about the behaviour of the DOM present
in our samples, complexation by dissolved organic
matter cannot be ruled out at this stage.
In contrast, although Hg and Pb are frequently
reported to have a high affinity for DOM, it can
be noted that a fast and strong re-adsorption of
these metals was observed. Hence, if complexation
by DOM did occur, as seems to be the case at
least for Hg according to the speciation calculations (Table 3), re-adsorption on solid phases was
the dominant process. This is confirmed by the
fact that in non filtered samples, the maximal Hg
concentration was also reached after 12 h of
shaking (1.2 mg ly1), and decreased as rapidly as
in the filtered samples (data not shown). This
shows that no significant complexation of Hg on
inorganic or organic colloidal particles was
observed, and that mercury was re-adsorbed rapidly on the sediment bulk phases.
In the case of Hg, this can be explained by the
fact that this element is bound efficiently on clay
minerals, organic matter, Al, and Fe oxides at
neutral pH (McNaughton, 1973; Alloway, 1995;
Tiffreau et al., 1995). However, as in the case of
Cu, the relative affinity Hg toward DOM or a
specific solid phase cannot be assessed without
248
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
any additional knowledge about its speciation in
solution.
5. Conclusion
The results presented here have shown that the
release of mercury from a resuspendedyaerated
anoxic sediment was very low. Although some
mercury was released in the early stages of the
aeration, the long-term experiments showed that
Hg was rapidly and durably re-adsorbed on the
sediment. This low mobility was observed albeit
high quantities of dissolved organic matter were
present in leaching solutions, and although mercury is considered to have a very high affinity of
natural organic matter (DOM).
In contrast, copper was durably released in
solution during the experiments. According to the
modelling results, copper was maintained in solution due to a strong complexation by carbonates
released by the dissolution of calcite. However,
complexation by DOM cannot be ruled out in this
case.
In a more general sense, these observations
highlight the fact that qualitative arguments are
not sufficient to explain the release of metallic
pollutants from dredged sediments. More precisely,
knowledge about the speciation in the leaching
solution appears to be a key factor in order to
make quantitative assessments about the potential
pollution emanating from dredged material dump
sites. However, this study showed that such a
modelling is at present particularly hampered by
the lack of information concerning the nature and
behaviour of the organic matter contained in anoxic sediments, although these materials are well
known for their high organic content. Moreover,
successful
characterisation
and
modelling
approaches have been applied to organic compounds emanating from surface waters. As a consequence, future works should focus on the nature
and complexing capabilities of natural compounds
contained in dredged sediments.
Acknowledgments
This work was financially supported by Ademe,
Voies Navigables de France, and the European
FEDER funds. We wish to thank H. Garraud and
O. Donard for the determination of mercury speciation at the LCABIE Laboratory (University of
Pau, France). The Douai School of Mines (ENSM)
is acknowledged for the XRD measurements.
References
Abbt-Braun G, Frimmel FH. Interaction of pesticides with
river sediments and characterisation of organic matter of the
¨
sediment. In: Calmano W, Forstner
W, editors. Sediments
and toxic substances. Berlin: Springer Verlag, 1996. p. 51 –
90.
´
´
AFNOR. Norme NF=31–130. Determination
de la capacite´
´
d’echange
cationique (CEC) et des cations extractibles. In:
Qualite´ des sols. Paris: editions AFNOR, 1999. pp 996.
AFNOR. Norme XPT 90–113 2. Determination du mercure.
´´
´´
In: Qualite´ de l’eau, tome 3: elements
majeurs-autres ele´ mineraux.
´
ments et composes
Paris: editions AFNOR, 2001.
´ ´
´
Alastuey A, Garcıa-Sanchez
A, Lopez
F, Quero X. Evolution
of pyrite mud weathering and mobility of heavy metals in
´
the Guadiamar valley after the Aznalcollar
spill, south-west
Spain. Sci Total Environ 1999;242:41 –55.
Allison LE. Organic Carbon. Methods of soil analysis, part 2.
Chemical and microbiological properties. Agronomy Monograph 9, 2nd edition. Madison, WI, USA: ASA, 1965. p.
1367 –1378.
Allison JD, Brown DS, Novo-Gradac KJ. MINTEQA2yPRODEFA2, a geochemical assessment model for environmental
systems: version 3.0 user’s manual, EPAy600y3-91y021
report, Athens, GE, USA, 1991, 115 p.
Allison JD, Brown DS, Novo-Gradac KJ. MINTEQA2yPRODEFA2, a geochemical assessment model for environmental
systems: user manual supplement for version 4.0, EPA
report, Athens, GE, USA, 1999, 76 p.
Alloway BJ. Heavy metals in soils. 2nd edition. New York:
Chapman and Hall, 1995. pp. 365.
´
Amouroux D, Tessier E, Pecheyran
C, Donard OFX. Sampling
and probing volatile metal(loid) species in natural waters
by in situ purge and cryogenic trapping followed by gas
chromatography and inductively coupled plasma mass spectrometry
(P-CT-GC-ICPyMS).
Anal
Chim
Acta
1998;377:241 –254.
´
Becart
D, Pavageau P, Vivant D. Suivi de la qualite´ des
´
` des analyses effectuees
´ entre
sediments,
bilan de synthese
1994 ET 1997. Voies Navigables de France, Douai, France,
1997, 170 pp.
Biester H, Scholtz C. Determination of mercury binding forms
in contaminated soils: mercury pyrolysis vs. sequential
extractions. Environ Sci Technol 1997;3:223 –239.
Blackburn TH. Release of nitrogen compounds following
resuspension of sediment: model predictions. J Marine Syst
1997;11:343 –352.
Brennan EW, Lindsay WL. The role of pyrite in controlling
metal ion activities in highly reduced soils. Geochim Cosmochim Acta 1996;60:3609 –3618.
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
Buffle J, Greter F, Haerdi W. Measurement of complexation
properties of humic and fulvic acids in natural waters with
lead and coper ion-selective electrodes. Anal Chem
1977;49:216 –222.
Cabaniss SE, Shuman MS. Copper binding by dissolved
organic matter: I. Suwannee River fulvic acid equilibria.
Geochim Cosmochim Acta 1988;52:195 –200.
¨
Calmano WW, Ahlf W, Forstner
U. Sediment quality assessment: Chemical and biological approaches. In: Calmano W,
¨
Forstner
U, editors. Sediment and toxic substances: environmental effects and ecotoxicity. Berlin: Springer-Verlag,
1996. p. 2 –35.
¨
Calmano W, Hong J, Forstner
U. Binding and mobilisation of
heavy metals in contaminated sediments affected by pH and
redox potential. Wat Sci Tech 1993;28:223 –235.
Cauwenberg P, Maes A. Influence of oxidation on sequential
chemical extraction of dredged river sludge. Intern J Environ
Anal Chem 1997;68:47 –57.
´ MC, Cornejo J. Heavy metal adsorption by
Celis R, Hermosın
functionalized clays. Environ Sci Technol 2000;34:4593 –
4599.
Cooper DG, Morse JW. Extractability of metal sulfides minerals in acidic solutions: Application to environmental studies of trace metal contamination within anoxic sediments.
Environ Sci Technol 1998;32:1076 –1078.
De Carvalho PSM, Zanardi E, Buratini SV, Lamparelli MC,
Martins MC. Oxidizing effect on metal remobilization and
Daphnia similis toxicity from a Brazilian reservoir sediment
suspension. Water Res 1998;32:193 –199.
De Rooij NM, Gerrits HJ. The transport of pollutants from a
sludge disposal site. Eur Water Pollut Control 1995;5:43 –
55.
Desauziers V, Castre N, Le Cloirec P. Sorption of methylmercury by clays and mineral oxides. Environ Sci Technol
1997;18:1009 –1018.
Dobbs JC, Susetyo W, Carreira LA, Azarraga LV. Competitive
binding of protons and metal ions in humic substances by
lanthanide ion probe spectrosopy. Anal Chem
1989a;61:1519 –1524.
Dobbs JC, Susetyo W, Knight FE, Castles MA, Carreira LA,
Azarraga LV. Characterization of metal binding in fulvic
acids by lanthanide ion probe spectroscopy. Anal Chem
1989b;61:483 –488.
Dood J, Large DJ, Fortey NJ, Milodowski AE, Kemp S. A
petrographic investigation of two sequential extraction techniques applied to anaerobic canal bed mud. Environ Geochem Health 2000;22:281 –296.
¨
Forstner
U. Non linear release of metals from aquatic sediments. In: Salomons W, Stgliani A, editors. Biogeodynamics
of pollutants in soils and sediments. Berlin: Springer Verlag,
1995. p. 247 –308.
¨
Forstner
U, Kersten M. Assessment of metal mobility in
dredged material and mine waste by pore water chemistry
¨
and solid speciation. In: Salomons W, Forstner
U, editors.
Chemistry and biology of solid waste: Dredged material and
mine tailings. Berlin: Springer-Verlag, 1989. p. 171 –214.
249
Giani M, Gabellini M, Pellegrini D, Costantini S, Beccaloni
E, Giordano R. Concentration and partition of Hg, Cr and
Pb in sediments of dredge and disposal sites of the northern
Adriatic Sea. Sci Total Environ 1994;158:97 –112.
Huerta-Diaz MA, Tessier A, Carignan R. Geochemistry of
trace metals associated with reduced sulfur in freshwater
sediments. Appl Geochem 1998;13:213 –233.
Isaure MP, Laboudigue A, Manceau A, Sarret G, Tiffreau C,
Trocellier P. Characterization of zinc in slags issued from a
contaminated sediment by coupling mPIXE, mRBS, mEXAFS and powder EXAFS scpectroscopy. Nucl Inst Meth
Phys Res 2001;181:598 –602.
Isaure MP, Laboudigue A, Manceau A, Sarret G, Tiffreau C,
Trocellier P, Lamble G, Hazemann JL, Chateigner D. Quantitative Zn speciation in a contaminated dredged sediment
by mPIXE, m-SXRF, EXAFS spectroscopy and principal
component
analysis.
Geochim
Cosmochim
Acta
2002;66:1549 –1567.
Kelley BC, Thuovinen OH. Microbiological oxidation of
¨
minerals in mine tailings. In: Salomons W, Forstner
U,
editors. Chemistry and biology of solid waste: dredged
material and mine tailings. Berlin: Springer-Verlag, 1989. p.
33 –54.
Kersten M. Geochemistry of priority pollutants in anoxic
sludges: cadmium, arsenic, methyl mercury and chlorinated
¨
organics. In: Salomons W, Forstner
U, editors. Chemistry
and biology of solid waste: dredged material and mine
tailings. Berlin: Springer-Verlag, 1989. p. 170 –213.
¨
Kersten M, Forstner
U. Chemical fractionation of heavy metals
in anoxic estuarine and coastal sediments. Wat Sci Technol
1986;18:121 –130.
Kinniburgh DG, Milne CJ, Benedetti MF, Pinheiro JP, Filieus
J, Koopal LK, van Riemsdijk WH. Metal ion bindig by
humic acid: Application of the NICA-Donnan model. Environ Sci Technol 1996;30:1687 –1698.
Large D, Fortey N, Milodowski AE, Christy AG, Dodd J.
Petrographic observations of iron, copper, and zinc sulfides
in freshwater canal sediment. J Sediment Res 2001;71:61 –
69.
Lee PK, Touray JC. Leaching of heavy metals (Zn, Cd, Pb)
from contaminated soils and sediments in the vicinity of
´
motorways: an experimental approach. Hydrogeologie
1997;1:3 –11.
Lin JG, Chen SY. The relationship between adsorption of
heavy metal and organic matter in river sediments. Environ
Int 1998;24:345 –352.
Maass B, Miehlich G. Die Wirkung des Redoxpotentials auf
¨
die Zusammensetzung der Porenlosung
in Hafenschlick¨
spulfeldern.
Mitteilgn Dtsch Bodenkundl Gesellsch
1988;56:289 –294.
Mantoura RFC, Riley JP. The use of gel filtration in the study
of humic acids and related compounds. Anal Chim Acta
1975;78:193 –200.
Marseille F, Tiffreau C, Laboudigue A, Lecomte P. Impact of
vegetation on the mobility and bioavailability of trace
elements in a dredged sediment deposit: A greenhouse study.
Agronomie 2000;20:547 –556.
250
N. Caille et al. / The Science of the Total Environment 301 (2003) 239–250
McNaughton MG. Adsorption of mercury(II) at the solidy
water interface. PhD Thesis, Stanford University, Palo Alto,
CA, 1973: pp 230.
Morse JW, Luther GM III. Chemical influences on trace metalsulfide interactions in anoxic sediments. Geochim Cosmochim Acta 1999;63:3373 –3378.
Perdue EM, Lytle CR. Distribution model for binding of
protons and metal ions by humic substances. Environ Sci
technol 1983;17:654 –660.
Rybicka EH, Calmano W, Breeger A. Heavy metals sorptiony
desorption on competing clay minerals; an experimental
study. Appl Clay Sci 1995;9:369 –381.
Schroeder PR, Aziz NM. Contaminant leaching model for
dredged material disposal facilities. J Environ Eng-asce
1999;9:835 –844.
Singh SP, Ma LQ, Tack FMG, Verloo MG. Trace metal
leachability of land-disposed dredged sediment. J Environ
Qual 2000;29:1124 –1132.
Stephens SR, Alloway BJ, Parker A, Carter JE, Hodson ME.
Changes in the leachability of metals from dredged canal
sediments during drying and oxidation. Environ Pollut
2001;114:407 –413.
Susetyo W, Carreira LA, Azarraga LV, Grimm DM. Fluorescence techniques for metal-humic interactions. Frenesius J
Anal Chem 1991;339:624 –635.
Tack FM, Callewaert OWJJ, Verloo MG. Metal solubility as a
function of pH in a contaminated, dredged sediment affected
by oxidation. Environ Pollut 1996;91:199 –208.
Tack FGM, Singh SP, Verloo MG. Leaching behaviour of Cd,
Cu, Pb and Zn in surface soils derived from dredged
sediments. Environ Pollut 1999;106:107 –114.
Thurman EM. In: Nijhoff M, Junk W, editors. Organic geochemistry of natural waters. The Netherlands: Dordrecht,
1985. p. 7 –66.
¨
Tiffreau C, Lutzenkirchen
J, Behra P. Modeling the adsorption
of mercury(II) on (hydr)oxides I. amorphous iron oxide
and a-quartz. J colloid Interface Sci 1995;172:82 –93.
¨
Wallschlager
D, Desai MVM, Wilken RD. The role of humic
substances in the aqueous mobilization of mercury from
contaminated floodplain soils. Water Air Soil Poll
1996;90:507 –520.
Wen Z, Allen HE. Mobilisation of heavy metals from Le An
River sediment. Sci Total Environ 1999;227:101 –108.
Yin Y, Allen HE, Huang CP, Sanders PF. Interaction of Hg(II)
with soil-derived humic substances. Anal Chim Acta
1997;341:73 –82.
Yu KC, Tsai LJ, Chen SH, Ho ST. Chemical binding of heavy
metals in anoxic river sediments. Water Res 2001;35:4086 –
4094.