Catena 122 (2014) 216–228
Contents lists available at ScienceDirect
Catena
journal homepage: www.elsevier.com/locate/catena
Application of chromium stable isotopes to the evaluation of Cr(VI)
contamination in groundwater and rock leachates from central Euboea
and the Assopos basin (Greece)
Maria Economou-Eliopoulos a,⁎, Robert Frei b,c, Cathy Atsarou a
a
b
c
Department of Geology and Geoenvironment, University of Athens, Athens 15784, Greece
Department of Geoscience and Natural Resource Management, University of Copenhagen, NordCEE, University of Copenhagen, Denmark
Department of Geoscience and Natural Resource Management, Nordic Center for Earth Evolution, NordCEE, University of Copenhagen, Denmark
a r t i c l e
i n f o
Article history:
Received 5 December 2013
Received in revised form 13 May 2014
Accepted 24 June 2014
Available online xxxx
Keywords:
Chromium isotopes
Groundwater
Leachates
Cr(VI)
Euboea
Assopos
a b s t r a c t
Major and trace elements (a) in groundwater, ultramafic rocks from natural outcrops and soil samples from cultivated sites of Central Euboea and Assopos basin, and (b) in experimentally produced laboratory water leachates
of rocks and soils were investigated by SEM/EDS, XRD and ICP/MS. In addition, stable chromium isotopes
(expressed as δ53Cr values) were measured in groundwater and leachates in order to identify potential sources
for Cr-contamination.
The higher Cr(VI) concentrations in soil leachates compared to those in the rock pulp leachates potentially can be
explained by the presence of larger amounts of Fe (Fe(II)) and Mn (Mn-oxides acting as oxidizing catalysts).
Assuming that redox processes produce significant Cr isotope fractionation (groundwater δ53Cr values range between 0.8 and 1.98‰), the compilation of the obtained analytical data suggests that the dominant cause of Cr isotope fractionation is post-mobilization reduction of Cr(VI). However, the lack of a very good negative relationship
between Cr(VI) concentrations and δ53Cr values may reflect that sorption, precipitation and biological processes
(fractionation during uptake by plants) complicate the interpretation of the Cr isotope signatures.
The variation in δ53Cr values (0.84 to 1.98‰ in groundwater from Euboea, and from 0.98 to 1.03‰ in samples
from the Assopos basin) imply initial oxidative mobilization of Cr(VI) from the ultramafic host rocks, followed
by reductive processes that lead to immobilization of portions of Cr(III). Using a Rayleigh distillation model
and different fractionation factors of Cr(VI) reduction valid for aqueous Fe(II) and Fe(II)-bearing minerals, we calculate that more than ~53%, but maximum ~94%, of the originally mobile Cr(VI) pool was reduced to immobile
Cr(III) in the waters investigated. This indicates that efficient processes in the aquifers may facilitate natural attenuation of the toxic Cr(VI) to less harmful Cr(III).
© 2014 Elsevier B.V. All rights reserved.
1. Introduction
Chromium contamination of soil and groundwater is a significant
problem worldwide and is becoming a serious threat to our environments. In nature, chromium occurs as trivalent [Cr(III)] and hexavalent
[Cr(VI)] species, with respective compounds (Hem, 1970). Health problems, such as lung cancer and dermatitis are caused by the highly toxic
and very soluble oxidized Cr(VI), in chromate oxyanions such as CrO2−
4 ,
2−
(ATSDR, 2000; Losi et al., 1994). In contrast, the reHCrO−
4 and Cr2O7
duced Cr(III) is an essential nutrient, required for normal glucose and
lipid metabolism in human bodies, adsorbs strongly on solid surfaces
and co-precipitates with Fe(III) hydroxides (Kotas and Stasicka, 2000).
Due to the toxicity of Cr(VI), most countries of the European Union
⁎ Corresponding author. Tel./fax: +30 210 7274214.
E-mail address: [email protected] (M. Economou-Eliopoulos).
http://dx.doi.org/10.1016/j.catena.2014.06.013
0341-8162/© 2014 Elsevier B.V. All rights reserved.
have currently regulated the limit to 50 μg·L− 1 for total chromium
[EC, 1998 Council Directive (98/83/EC)].
A number of studies have shown that chromium stable isotopes are
effective in monitoring Cr in natural conditions to determine natural
and/or anthropogenic sources (Basu and Johnson, 2012; Berna et al.,
2010; Døssing et al., 2011; Ellis et al., 2002, 2004; Han et al., 2012;
Izbicki et al., 2008; Jamieson-Hanes et al., 2012; Johnson, 2011; Kitchen
et al., 2012; Novak et al., 2014). Recently, Cr stable isotopes have seen
growing use in environmental applications that range form monitoring
Cr at a contaminated site to paleo-environmental applications that
examine oxygenation of our environment in the Precambrian (BIFs,
paleosols) etc. (Crowe et al., 2013; Frei and Polat, 2013; Frei et al.,
2009). The monitoring of Cr-contaminated groundwater using Cr stable
isotope tracing has been demonstrated by Berna et al. (2010), Ellis et al.
(2002), Halicz et al. (2008); Izbicki et al. (2008), Novak et al. (2014);
Schoenberg et al. (2008), Sikora et al. (2008), Zink et al. (2010) and
others.
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
Chromium has four stable isotopes; 54Cr, 53Cr, 52Cr, and 50Cr, with natural abundances of 2.37%, 9.5%, 83.8%, and 4.35%, respectively (Moynier et
al., 2011; Rotaru et al., 1992). Redox processes have been shown to produce significant Cr isotope fractionation during the transition from
Cr(VI) to Cr(III) (Schauble et al., 2004). During reduction, the lighter isotopes are preferentially reduced, resulting in an enrichment of 53Cr relative to 52Cr values in the remaining Cr(VI) pools. This enrichment is
measured as the change in the ratio of 53Cr/52Cr, and is expressed as
δ53Cr values in units per mil (‰) relative to a standard (Ellis et al.,
2002). The enrichment or depletion of 53Cr relative to 52Cr can be quantified by measuring the 53Cr/52Cr values in aqueous solutions (Basu and
Johnson, 2012; Berger and Frei, 2013; Berna et al., 2010; Farkas et al.,
2013; Halicz et al., 2008; Han et al., 2012; Izbicki et al., 2008;
Jamieson-Hanes et al., 2012; Kitchen et al., 2012; Schoenberg et al.,
2008; Sikora et al., 2008; Zink et al., 2010; Wanner and Sonnenthal,
2013). The reduction of Cr(VI) species to Cr(III) species in aqueous systems, by abiotic (e.g. Fe(II)-minerals) and/or biological (microorganisms,
organic acids) reduced species is accompanied by an isotope fractionation
preferring the light isotopes in the reductant (Basu and Johnson, 2012;
Sander and Koschinsky, 2000). The reduction of toxic and mobile Cr(VI)
to Cr(III) is a remediation technology commonly proposed, and can naturally be enhanced by organic matter, Fe(II)-minerals and reduced species
of sulfur (Kozuh et al., 1994). Thus, measurement of the 53Cr/52Cr values
in groundwater has been proposed as a method to track Cr(VI) migration
processes and evaluate the performance of remediation activities
(Blowes, 2002; Ellis et al., 2002; Jamieson-Hanes et al., 2012).
Since the fractionation of Cr isotopes is considered to be little affected by dilution or adsorption processes (Ellis et al., 2004), local anomalies
in the isotope signature of natural water can be used as a tracer for the
reduction of Cr(VI). Thus, they can potentially give clues to the efficiency of natural attenuation processes transforming the dissolved and toxic
hexavalent Cr(VI) to less harmful Cr(III) at specific sites (Berna et al.,
2010; Ellis et al., 2002; Izbicki et al., 2008; Johnson, 2011; Raddatz
et al., 2011).
The research interest has focused on the Assopos basin because it is
an industrial zone (hundreds of industrial plants, such as using chromium plating, leather tanning, and applying wood staining) and the running Assopos river was proclaimed as a “processed industrial waste
receiver” since 1969. Besides, untreated or poorly treated industrial
waste may have been dumped illegally in now covered fills. The Assopos
(Avlona) and Central Euboea basins (Messapia), dominated geologically
by the widespread occurrence of ophiolites, were selected for the present
study, because there is no clear-cut answer to the question regarding the
influence of industry versus natural processes to the soil and groundwater contamination (Economou-Eliopoulos et al., 2011, 2012). The herein
presented integrated approach is based on a compilation of geochemical/hydrochemical data of (a) ultramafic rocks, soils and groundwater
samples, (b) experimentally produced laboratory water leachates of
these rocks and soils and (c) stable chromium isotope data (expressed
as δ53Cr values) of selected natural and experimentally produced leachate water samples, originally aimed at identifying potential sources for
Cr-contamination in these basins.
2. Geological and hydrological outline
2.1. Central Euboea
The area of central Euboea is covered by alluvial and Neogene sediments. It is characterized by strong geomorphological contrast and is
built up mainly of Pleistocene to Holocene sediments hosting the most
productive aquifers in this area (Fig. 1, sampling area). In addition,
two different types of aquifers are hosted by strongly tectonized ultramafic rocks, which are widespread in central Euboea, and by the deeper
karstified Triassic–Jurassic limestones.
The ophiolitic masses consist mainly of serpentinised peridotites
(harzburgites and lherzolites) with some minor mafic rocks. The
217
ophiolitic rocks are overthrusted onto Upper Cretaceous limestones
and flysch sediments. The main aquifer which is probed by the wells
is hosted by ophiolitic rocks and is categorized as a fissured rock aquifer.
Alluvial deposits are the host rocks to the aquifer which is probed by
many shallow wells for agricultural activities. These wells reach depths
between 11 and 180 m (Megremi, 2010).
2.2. Assopos basin (Avlona)
The Neogene Assopos basin (Fig. 1, sampling area 2) is mainly composed by Tertiary and Quaternary sediments of more than 400 m thickness, and expands over approximately 700 km2. Alternations of marls
and marly limestones occur in the lowest parts of the basin sequences,
and continental sediments consisting of conglomerates with small intercalations of marls, marly limestones, schists, sandstones, clays and
flysch are dominant in the upper parts. A sharp tectonic contact between the sediment types, due to the intense neotectonic deformation,
is a characteristic feature of the entire area (Chatoupis and Fountoulis,
2004). Peridotites and a Ni-laterite occurrence, overthrusted on the
Triassic–Jurassic carbonates, have been described from the Aynola
area by (Valeton et al., 1987). The morphotectonic structure and evolution of this basin are the result of E–W to WNW–ESE trending fault
systems (Chatoupis and Fountoulis, 2004; Papanikolaou et al., 1988).
Quaternary sediments cover large parts of the Assopos valley and host
two types of aquifers: a) aquifers within Neogene conglomerates, sandstones and marly limestone to a depth approximately 150 m, and
b) karst type aquifers within the Triassic–Jurassic limestones at deeper
levels of the basin fill (Giannoulopoulos, 2008).
3. Samples and methods of investigation
For the purpose of the present study, 10 groundwater samples, 15
soil samples and 21 rock samples were collected from the extended
area of the municipality of Messapia in central Euboea and from the
neighboring area of Avlona located in the Assopos basin (Fig. 1).
Soil and rock samples were collected from cultivated sites and from
natural outcrops of ultramafic rocks on Central Euboea. Soils were air
dried, crumbled mechanically and those containing large stones or
clods were first sieved through a 10 mm mesh and then through a
5 mm mesh. Subsequently, after passing the samples through a
2 mm mesh, the fraction b 2 mm was pulverized and used for analysis.
Rock samples were crushed by jaw crusher, then pulverized using a triturator and an agate mortar and pestle, and subsequently sieved through
a b2 mm mesh. This fraction was used for the leaching experiments.
Major and trace elements were analyzed by inductively coupled plasma
mass spectroscopy (ICP-MS) after multi-acid digestion (HNO3–HClO4–
HF–HCl) at the ACME Analytical Laboratories in Canada. The detection
limits for those elements are presented along with the analytical results
in Table 1.
Groundwater samples were collected from domestic and irrigation
wells spread over the study area in October 2012. Physical and chemical
parameters (pH, redox, total dissolved solids, conductivity and total dissolved solids) of the water samples were measured in the field using a
portable Consort 561 Multiparameter Analyzer. The collected samples
were divided into two aliquots and each one was stored in polyethylene
containers at 4 °C in a portable refrigerator. One of the sample aliquots
was acidified by addition of concentrated HNO3 and stored at 4 °C as
well. Because acidification potentially can affect the solubility of Cr(VI)
and because biotic activity could change the valence state of chromium
in the samples, concentrations of total Cr and Cr(VI) were determined in
the non-acidified aliquot of the water samples, within 24 h after
collection. The analyses of total chromium were performed by GFAAS
(Perkin Elmer 1100B system), with an estimated detection limit of
~ 2 μg/L. The chemical analyses for Cr(VI) were performed by the 1,5diphenylcarbohydrazide colorimetric method, using a HACH DR/4000
spectrophotometer. The estimated detection limit of the method was
218
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
Fig. 1. Location map showing the localities of sampling. 1. Central Euboea, 2. Assopos Basin.
determined at ~ 4 μg·L−1. All the above described analyses were performed at the Laboratory of Economic Geology and Geochemistry,
Faculty of Geology and Geoenvironment, University of Athens. Other
elements were analyzed in the acidified portion of the samples by
Inductively Coupled Plasma Mass Spectroscopy (ICP/MS) at ACME Analytical Laboratories in Canada. The detection limits for those elements
are presented along with the analytical results in Table 2.
The mineralogical composition of soil and rocks was investigated by
optical microscopy, X-ray diffraction and mineral phase analysis. XRD
data were obtained using a Siemens Model 5005 X-ray diffractometer,
applying Cu Ka radiation at 40 kV and 40 nA, in 0.020° steps at 1.0 s
step intervals. The XRD patterns were interpreted using the EVA 2.2 program included in the D5005 software package. Polished sections prepared from soil and rocks, after carbon coating, were examined by
reflected light microscopy and with a scanning electron microscope
(SEM) and its energy dispersive spectroscopy (EDS) tool. Microprobe
analyses and SEM imaging were carried out at the Department of Geology and Geoenvironment, University of Athens, using a JEOL JSM 5600
scanning electron microscope, equipped with automated energy dispersive analysis system ISIS 300 OXFORD, with the following operating
conditions: accelerating voltage of 20 kV, beam current of 0.5 nA, time
of measurement of 50 s and beam diameter of 1–2 μm.
A series of batch leaching experiments were carried out in order to
study the long-term leaching responses of Cr under atmospheric conditions. For these experiments, 10 g of a crushed soil and/or rock sample
was suspended in 100 mL of deionized water in a 200 mL Erlenmeyer
flask at room temperature. The reaction flask was shaken at approximately 120 rpm by a reciprocal shaker for seven days. After the period
of shaking, the slurries were filtered through a 0.45 μm polyamide membrane filter. The filtered leachates were first analyzed for Cr(VI) concentrations, using a HACH DR/4000 spectrophotometer (estimated
detection limit ~ 4 μg·L−1), and then for total Cr by GFAAS (Perkin
Elmer 1100B system) (estimated detection limit of the method ~2 μg/L).
Water samples in the amount which would yield about 1 μg of total
chromium were pipetted into 25 mL Erlenmeyer flasks together with an
amount of a 50Cr–54Cr double spike so that a sample to spike ratio of
~3:1 (total chromium concentrations) was achieved. The addition of a
50
Cr–54Cr double spike of a known isotope composition to a sample before chemical purification allows accurate correction of both the chemical and the instrumental shifts in Cr isotope abundances (Ellis et al.,
2002; Schoenberg et al., 2008). The mixture was totally evaporated
and 3 mL of concentrated aqua regia was subsequently added. After 3
h during which the sample was exposed to aqua regia on a hotplate at
100 °C, the sample was again dried down. Finally, the sample was
then taken up in 20 mL of Milli Q water and 1 mL of 1 N HCl, to which
0.5 mL of a 1 M ammonium peroxydisulfate solution (puratronic® quality) was added. The samples were then boiled for 30 min in a sand bath,
during which an hour glass prevented evaporation of the sample in the
Erlenmeyer flask. This enabled the total oxidation of the chromium to
Cr(VI). After cooling to room temperature, the solution was then passed
over an extraction column (BioRad) charged with 2 mL of intensively
pre-cleaned 200–400 mesh AG1 × 8 (BioRAD) anion resin. Cr(VI) is
retained by the resin while cations such as Ca2+, Na+, and K+ are efficiently washed out. After rinsing with 5 mL of 0.1 N HCl, Cr(VI) was reduced, during 30 min on the columns, with 1 mL of 2 N HNO3 to which
three drops of hydrogen peroxide were added. Cr(III) was then extracted with another 5 mL of the same 2 N HNO3–hydrogen peroxide mixture into a 17 mL Savillex™ beaker and subsequently dried down. This
extraction procedure usually has a chromium yield of N90%. The so produced chromium fraction was then purified by passing the sample in 0.5
N HCl over a miniaturized disposable pipette-tip extraction column
(fitted with a bottom and a top disposable PVC frit) charged with
300 μL of a 200–400 mesh cation resin (AGW-X12; BioRad) employing
a slightly modified extraction recipe published by Trinquier et al.
(2009) and Bonnand et al. (2011). The yield of this mini-column extraction and purification step is usually ~70%.
Samples were loaded onto Re filaments with a mixture of 3 μL silica
gel, 0.5 μL 0.5 mol L−1 of H3BO3 and 0.5 μL 0.5 mol L−1 of H3PO4. The
samples were statically measured on a IsotopX “Phoenix” multicollector
thermal ionization mass spectrometer (TIMS) at the Department of
Table 1
Major and trace elemnt contents in ultramafic rock and soils from central Euboea (Messapia).
mg·kg−1
wt.%
Cr
Ni
Co
Mn
M1R1
M1R2A
M1R2B
M1R3
M1R4
M1R5
M1R6
M1R7
M1R8
M1R9
M1R10
M2R1
M2R2
M2R3
M2R4
M2R5
M2R6
M3R1
M3R2
M3R3
M3R4
1180
840
2100
1340
1620
1660
1660
1050
1020
1320
1250
160
860
520
320
1170
1000
1630
1140
1720
1100
940
940
1340
1730
1510
1730
2000
1650
1520
2120
1770
1810
2300
2030
1640
2140
2090
2200
1490
1850
1630
82
44
110
85
74
87
90
84
77
100
80
66
93
95
76
100
100
100
74
95
70
1900
1180
1120
1040
900
740
680
610
640
790
680
340
540
600
520
980
650
530
530
660
520
24
30
28
16
17
14
29
7.4
21
8.7
17
2.5
2.7
2.8
2.9
2.8
5.2
10
8.3
9.7
6.8
2.8
7.7
1.9
0.5
0.1
0.2
1.7
0.3
0.2
0.3
0.1
b0.1
0.1
0.2
0.4
b0.1
0.1
b0.1
1.0
0.2
b0.1
Soils
MSS1
MSS2
MSS3
MSS4
MSS5
MSS6
MSS7
MSS8
MSS9
MSS10
MSS11
MKR1S
MKR2S
PS1
PS2
Det. limit
670
930
2100
1050
880
2100
1340
1760
1790
3350
2200
2000
1380
2750
2300
1
610
930
2260
1190
780
2200
1460
2220
2300
3050
2220
2560
1830
2280
2100
0.1
42
760
60
1050
120
1180
72
1080
52
780
120
1200
80
790
120
1130
110
690
160
1360
140
1630
130
1130
97
990
100
1070
97
980
0.2
1
28
36
24
37
30
30
28
25
22
27
57
19
23
34
22
0.1
200
811
879
183
2
144
304
312
131
0.2
49
586
579
45.0
0.7
Reference materials
STD OREAS24P
STD OREAS45C
STD OREAS45C
STD OREAS24P
BLK
43
95
95
41
b0.2
1061
1086
1005
949
3
Cu
Pb
Zn
Cd
Sb
As
39
48
39
28
43
39
50
25
29
53
38
17
23
22
18
35
30
43
26
37
22
b0.1
0.2
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
0.4
0.2
0.4
0.2
0.1
0.1
b0.1
b0.1
b0.1
0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
4
3
6
b1
b1
1
b1
1
1
2
b1
b1
b1
b1
1
b1
b1
b1
3
b1
b1
15
18
15
17
17
16
16
12
12
12
22
9.1
15
12.4
9
0.1
61
83
83
99
73
96
79
84
71
96
110
70
69
99
70
1
0.2
0.2
0.2
0.3
0.2
0.4
0.4
0.2
0.3
b0.1
0.5
0.1
0.3
0.4
0.2
0.1
6.8
0.7
0.4
0.7
0.7
0.4
0.4
0.3
0.2
0.3
0.7
0.3
0.4
0.3
0.2
0.1
11
14
7
12
14
6
6
6
5
6
14
2
8
3
3
1
3.2
25.3
21.8
2.8
0.3
111
75
73
100
b1
b0.1
0.2
0.2
b0.1
b0.1
b0.1
0.8
1.7
b0.1
b0.1
2
11
10
1
b0.01
Zr
Y
Sr
6.4
31
4.2
0.4
0.5
0.5
0.2
0.4
0.5
0.3
0.2
0.1
0.1
0.1
0.2
b0.1
0.1
b0.1
1.7
0.2
0.2
3.9
8.8
3.5
1.1
1.6
1.0
0.9
0.7
0.9
1.0
0.5
b0.1
b0.1
0.1
b0.1
0.3
b0.1
b0.1
0.9
0.3
0.2
13
19
16
17
59
13
4
5
105
5
6
67
65
21
70
23
7
2
20
6
23
32
39
22
36
30
31
35
34
19
23
50
12
33
20
19
0.1
13
15
8
14
13
10
11
9.3
6.8
7.1
16
3.6
9.7
5.7
5.2
0.1
128
157
157
123
0.1
22.7
12.7
12.6
20.8
b0.1
Mg
Ca
P
Ti
11
8
12
9
14
11
6
5
7
11
3
3
3
4
4
6
4
2
11
4
7
45
53
57
22
37
41
57
53
45
41
38
2
4
9
7
36
30
43
25
33
13
3.7
11.9
2.3
0.6
0.1
0.1
b0.1
b0.1
b0.1
b0.1
0.1
0.2
b0.1
b0.1
0.1
b0.1
b0.1
b0.1
0.7
b0.1
0.1
9
27
6
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
1
b1
b1
0.8
3.9
0.5
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
0.3
b0.1
b0.1
1.2
5.9
0.7
0.2
0.2
0.3
0.1
0.2
b0.1
0.1
0.2
0.1
0.1
0.2
0.2
0.2
0.1
0.1
0.4
0.2
0.1
0.1
0.7
0.1
b0.1
b0.1
0.1
0.1
0.4
0.2
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
0.9
1.1
0.7
0.4
15
55
20
30
19
4.5
3.3
1.8
3.0
2.2
1.4
0.4
0.7
0.5
0.5
2.9
0.8
3.9
6.3
1.5
1.8
3.0
3.4
0.8
9.7
2.5
1.2
4.6
2.8
3.0
6.6
1.4
5.6
4.8
3.5
4.9
4.4
5.1
13
1.7
1.6
5.4
4.0
3.8
6.2
4.4
4.0
5.0
5.8
5.2
4.3
5.5
5.5
2.7
3.9
4.2
3.3
6.0
5.7
5.8
3.7
5.2
3.5
1.3
3.1
1.3
0.6
0.8
0.7
0.8
0.5
0.9
0.7
0.5
0.0
0.0
0.1
0.1
0.6
0.3
0.5
0.5
0.4
0.3
11.3
9.8
6.7
3.8
11.4
18.2
21.2
19.1
18.7
22.2
21.3
12.6
17.8
19.3
12.3
18.9
20.2
21.9
9.8
17.0
10.1
10.6
10.4
12.2
20.9
12.2
4.6
0.7
2.6
3.6
0.2
1.6
11.7
5.7
5.4
15.3
1.3
1.7
0.3
15.4
7.0
18.5
0.004
0.027
0.003
b0.001
b0.001
0.001
b0.001
b0.001
b0.001
b0.001
b0.001
0.002
0.002
0.001
0.004
0.001
0.003
b0.001
0.003
0.001
0.002
0.052
0.190
0.056
0.022
0.028
0.023
0.024
0.015
0.021
0.021
0.014
b0.001
b0.001
0.002
0.001
0.006
0.003
0.005
0.010
0.005
0.002
0.047
0.017
0.016
0.005
0.007
0.003
0.004
0.003
0.007
0.002
0.002
0.004
0.004
0.004
0.005
0.004
0.003
0.003
0.008
0.003
0.008
0.09
0.04
0.05
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
b0.01
0.03
b0.01
b0.01
80 180
64 220
20 120
58 190
76 200
40 140
49 140
28 120
21
80
22
98
43 210
19
47
28 120
23
91
28
73
1
1
61
77
55
76
67
61
56
59
46
74
100
44
74
54
51
1
20
23
10
20
21
14
16
14
9
10
23
6.1
14
8.4
7.8
0.1
40
47
22
41
42
30
32
28
17
22
48
11
27
17
17
1
6.3
7.1
3.5
6.1
6.5
4.5
5.4
4.7
2.8
3.3
7.9
1.4
4.9
2.6
2.6
0.1
6.6
7.5
4.0
7.0
6.2
5.2
6.5
5.8
3.5
4.0
8.8
2.0
5.7
3.2
3
0.1
1.2
1.4
0.8
1.4
1.3
1.1
1.0
1.2
0.7
0.8
1.4
1.0
0.8
0.9
0.6
0.1
27
31
12
29
30
12
16
12
14
15
27
6.7
24
7.2
9.4
0.1
2.6
1.1
1.7
5.3
b0.5
2.9
5.1
2.9
4.7
2.2
b0.5
3.2
2.7
3.5
1.1
0.5
3.7
5.3
7.7
5.9
4.3
8.5
6.0
8.6
8.1
11.6
10.0
8.2
6.8
7.5
7.3
0.01
3.9
4.7
2.3
4.1
4.1
2.7
3.2
2.9
2.0
2.4
4.9
1.3
3.6
1.9
1.8
0.01
2.8
4.3
13.8
6.2
3.7
10.8
7.8
10.8
14.8
10.8
3.9
14.8
6.2
14.4
13.4
0.01
6.2
3.2
0.6
3.1
6.1
1.8
3.2
0.9
0.6
0.8
3.1
0.3
7.9
0.8
1.9
0.01
0.06
0.09
0.05
0.14
0.05
0.09
0.10
0.04
0.04
0.04
0.07
0.05
0.04
0.09
0.04
0.001
0.20
0.24
0.11
0.23
0.21
0.14
0.17
0.13
0.09
0.11
0.24
0.05
0.16
0.09
0.08
0.001
0.44
0.52
0.15
0.45
0.49
0.23
0.28
0.16
0.11
0.11
0.23
0.05
0.10
0.12
0.1
0.001
0.96
1.19
0.56
1.11
1.05
0.75
0.81
0.68
0.40
0.54
1.01
0.29
0.66
0.54
0.49
0.01
155
251
228
136
b1
19.8
27.7
24.0
16.5
b0.1
36
47
47
32
b1
3.0
11.1
10.1
2.7
b0.1
19.1
21.5
22.7
17.0
b0.1
0.7
2.4
2.1
0.7
b0.1
8.1
15.6
13.0
7.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
7.37
16.49
19.26
7.60
b0.01
7.67
6.90
6.71
6.79
b0.01
3.97
0.25
0.21
3.72
b0.01
5.65
0.47
0.45
5.11
b0.01
0.137
0.052
0.048
0.123
b0.001
1.035
1.069
1.044
0.914
b0.001
2.252
0.098
0.086
2.084
b0.001
0.65
0.32
0.32
0.60
0.01
380
35
32
321
b1
Ba
267
259
261
258
b1
V
La
Ce
Th
Nb
U
Li
Te
Fe
Al
Na
K
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
Rocks
219
220
Table 2
Trace element concentrations in groundwater and water after leaching peridotites and soils from C. Euboea (Messapia) and the Assopos basin (Avlona).
μg·L−1
mg·kg−1
Wells
Crtotal
Cr(VI)
E2
E5
E7
E10
E13
E18
MSW8
M3W8
AVLO13W
AVLO14W
77
102
230
48
65
41
65
45
50
93
77
102
230
48
62
40
63
43
48
85
14
18
35
64
30
3.0
3.2
2.0
2.3
2.2
3.9
3.1
4.8
11
3.4
5.0
6.4
5.5
3.6
8.8
15
12
16
35
63
29
b4
b4
b4
b4
b4
b4
b4
4.5
8
b4
4.8
5.0
4.0
b4
7.0
13
Soil leachates
MSS1
MSS2
MSS3B
MSS4B
MSS5B
MSS6B
MSS7B
MSS8B
MSS9
MSS11
MSS10
MKR1S
MKR2S
PS1
PS2
Detection limit
21
27
94
38
37
76
35
64
57
58
84
70
22
54
58
0.5
17
21
87
34
28
68
35
43
43
76
81
48
18
51
55
Reference materials
STD TMDA-70
STD TMDA-70
BLK
404.2
413.5
b0.5
As
B
Ba
3
2
1
2
3
1
3
1
2
1
b0.5
0.9
b0.5
0.8
0.7
b0.5
1.5
b0.5
1.0
3.5
105
88
57
46
136
29
54
35
35
37
42
74
27
18
21
17
4
16
37
64
b1
4
b1
b1
1
2
2
1
1
2
2
1
1
b1
b1
1
1
1
1
1
2
0.5
1.6
1.7
b0.5
0.8
0.6
0.5
b0.5
0.6
0.7
0.6
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
268
154
124
98
78
46
23
103
204
106
55
272
46
40
148
221
143
58
229
48
292
3
7
3
4
7
2
2
6
3
4
4
2
5
4
4
1
4.9
11
5.2
10
4.6
5.6
12
3.4
4.6
2.7
2.9
1.6
2.1
5.7
2.3
0.5
506
543
b1
41.9
43.0
b0.5
Br
Co
Cu
Li
Mn
Ni
P
Pb
S
Sb
Sc
Se
Sr
Ti
U
V
Zn
Ca
Mg
Na
K
69
94
124
72
66
49
114
48
74
70
47.6
40.0
32.2
35.0
39.5
22.3
25.9
22.7
25.9
29.8
1.6
5.4
1.9
2.0
2.3
1.0
1.9
1.7
0.9
0.8
18
25
38
24
25
26
49
25
22
29
37
51
57
46
47
30
46
30
40
97
13
12
12
18
13
15
17
18
23
20
19
26
33
23
23
22
22
19
17
17
20
b1
1
b1
b1
b1
b1
b1
3
b1
b1
4
b1
b1
b1
b1
b1
5
4
b1
2
b1
17
20
22
14
13
17
14
21
21
17
12
18
11
21
18
40.00
b1
b1
b1
b1
b1
2
1
b1
b1
b1
b1
3
b1
1
1
1
120
210
290
150
140
160
240
130
150
300
0.1
0.2
0.3
0.1
0.1
0.1
0.1
0.0
0.3
b0.02
4.0
2.7
2.6
2.4
1.8
1.7
2.6
0.9
1.1
0.6
4.6
7.5
8.9
4.8
5.5
3.1
8.4
3.3
5.9
12
0.5
0.2
1.2
5.6
0.4
b0.05
0.3
b0.05
0.7
0.1
2.3
10
6.5
5.6
6.8
5.0
5.1
15
3.0
0.8
b10
b10
b10
b10
b10
b10
10
b10
b10
b10
0.7
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
20
43
30
16
15
18
15
11
8
3
b0.05
b0.05
b0.05
b0.05
b0.05
b0.05
b0.05
b0.05
b0.05
b0.05
4
5
7
5
5
5
10
5
4
6
0.7
1.8
2.2
1.2
0.6
0.9
1.7
0.6
0.9
2.2
291
332
260
155
174
154
61
161
151
203
b10
b10
b10
b10
b10
b10
13
b10
b10
b10
1.19
3.28
0.20
0.11
b0.02
2.74
0.23
1.84
0.15
0.34
1.6
2.1
2.1
1.5
2.0
1.0
3.7
1.2
2.8
9.6
102
1.3
1.2
6.5
b0.5
b0.5
1.6
b0.5
1.5
b0.5
90
102
92
76
79
83
38
81
80
47
11
3.3
2.4
1.3
5.1
1.6
1.5
1.6
7.0
3.6
9.0
1.7
1.9
3.2
3.3
1.2
1.0
1.8
1.7
1.7
1.7
7
13
b5
b5
b5
b5
b5
12
b5
b5
13
5
11
8
b5
5
13
18
b5
b5
27
b0.02
0.04
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
0.03
0.06
0.05
0.04
0.05
0.03
b0.02
0.04
b0.02
0.08
0.04
0.05
0.9
1.5
0.3
0.5
0.6
0.5
0.4
0.3
0.3
0.4
0.7
0.3
0.3
0.4
0.2
0.4
b0.1
0.2
0.3
0.3
0.4
1.4
2.7
3.2
0.2
14.8
3.4
1.6
1.0
1.4
0.8
0.5
0.5
0.5
0.5
0.4
0.4
0.6
0.9
1.6
0.5
0.9
0.2
0.1
0.2
b0.05
0.5
0.1
0.7
0.1
b0.05
0.2
b0.05
0.2
0.5
0.1
b0.05
0.1
b0.05
b0.05
b0.05
b0.05
0.1
7.0
4.2
2.3
0.9
0.7
0.7
1.1
0.7
0.9
0.8
0.7
0.2
0.4
b0.2
b0.2
0.7
0.7
b0.2
0.6
0.4
0.7
b10
11
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b0.1
b0.1
b0.1
b0.1
b0.1
0.3
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b0.1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
b1
0.37
0.19
0.11
b0.05
0.36
0.15
0.24
0.27
0.13
0.16
0.56
b0.05
b0.05
b0.05
b0.05
b0.05
b0.05
0.11
0.10
0.07
b0.05
3
2
3
b1
3
3
4
4
5
1
b1
1
1
b1
b1
b1
b1
b1
b1
b1
b1
0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
b0.5
20
20
19
21
15
13
12
15
13
27
56
17
22
22
11
16
11
21
17
16
16
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
0.04
0.03
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
b0.02
1.2
3.2
4.2
1.4
2.1
2.2
2.0
4.3
5.8
2.9
2.1
1.3
b0.2
b0.2
0.2
0.9
0.7
1.0
1.9
1.1
0.5
9.1
3.5
5.5
1.3
6.1
6.9
6.0
3.5
3.7
8.2
2.2
4.7
29
2.3
b0.5
1.9
1.8
0.5
1.9
1.6
5.8
27
26
26
26
25
18
18
12
13
13
15
15
25
26
12
17
16
20
18
21
20
2.1
1.9
1.9
6.1
2.9
4.7
6.8
11.4
11.8
8.6
20.0
11.6
8.1
7.6
13.8
7.0
11.3
13.2
6.3
8.1
6.8
3.2
4.2
3.4
2.8
2.4
2.1
1.9
3.3
1.9
1.8
1.6
1.7
1.5
1.6
1.6
1.7
2.3
2.1
2.1
1.7
4.0
1.8
1.1
0.7
0.6
0.6
0.6
0.9
0.7
0.8
0.6
0.6
0.6
0.8
1.6
0.6
0.7
0.6
0.4
1.0
0.8
1.5
2811
5417
930
1233
1412
767
254
253
248
620
1019
286
511
620
740
5
8.5
9.6
8.5
10
11
12
11
11
6.8
10
10
7.2
6.0
9.1
6.7
0.05
52
54
24
36
27
31
45
26
43
49
35
24
34
18
18
5
0.2
0.4
0.4
0.4
0.3
0.9
0.5
0.8
0.1
0.2
0.3
0.63
0.46
0.8
0.4
0.0
6.0
6.7
3.4
6.7
3.8
3.7
6.5
7.4
3.5
5.2
7.3
4.5
2.8
12.0
2.0
0.1
5.5
7.3
2.9
4.4
4.5
2.3
3.1
2.7
4.0
3.0
2.7
3.7
1.8
2.2
2.4
0.1
0.2
0.2
0.2
0.2
0.2
0.3
0.1
0.5
0.1
0.1
0.1
0.4
0.2
0.2
0.1
0.1
10
15
32
21
9
28
25
39
22
13
17
54
24
51
13
0
38
1040
874
2124
120
1236
747
746
972
214
463
756
38
1910
360
10
b0.1
0.4
0.2
0.5
b0.1
0.2
b0.1
0.2
b0.1
0.1
0.2
b0.1
0.2
0.2
b0.1
0.1
2
1
b1
1
1
3
2
3
1
b1
1
2
1
2
b1
1
0.40
0.67
0.24
0.41
0.55
0.38
0.46
0.13
0.22
0.12
0.16
0.27
0.13
0.24
0.28
0.05
b1
b1
b1
b1
b1
b1
b1
b1
1
b1
b1
b1
b1
1
1
1
1.1
1.4
b0.5
1.1
0.9
0.8
1.1
b0.5
1.1
b0.5
0.5
b0.5
1.1
0.8
0.5
0.5
45.8
50.4
29.9
38.6
45.3
49.9
47.7
41.0
32.4
26.2
32.5
46
22
36
31
0.01
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
b10
10
0.47
0.50
0.20
0.31
0.56
0.42
0.35
0.46
0.13
0.19
0.50
0.07
0.29
0.42
0.20
0.02
4.9
8.8
4.0
7.4
4.1
5.9
10.7
4.4
6.9
4.1
3.7
3.1
2.4
4.3
3.4
0.2
7.8
5.0
9.8
1.3
3.8
4.1
4.2
43.5
1.8
5.5
7.6
19
9.6
33.0
b0.5
0.5
30
26
13
20
34
25
29
26
20
26
32
16
44
17
29
0
11.7
15.9
17.4
18.7
6.8
12.8
12.1
21.4
15.7
8.4
6.1
14.2
4.6
16.0
14.0
0
7.5
10.1
4.1
5.9
4.1
4.4
4.7
3.6
4.6
3.9
3.4
3.9
2.3
3.3
3.4
0.1
7.8
9.5
12.6
16.3
8.6
19.8
8.7
14.4
6.4
6.2
5.7
18
11
31
19
0.1
17
18
b5
332.7
348.2
b0.05
28
25
b5
287.3
292.8
b0.02
399.4
415.2
b0.1
22.6
21.1
b0.1
327
340
b0.05
330
324
b0.2
b10
b10
b10
468
456
b0.1
7
b1
b1
22.30
23.40
b0.05
b1
b1
b1
26.4
26.2
b0.5
452.07
460.98
b0.01
b10
b10
b10
59.30
61.95
b0.02
332.6
322.8
b0.2
497.3
504.0
b0.5
22.41
23.70
b0.05
5.78
6.05
b0.05
8.8
8.7
b0.05
1.00
1.00
b0.05
Si
428
469
b40
Cl
13
13
b1
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
Rock leachates
M1R1
M1R2AB
M1R2BA
MIR2BB
M1R3
M1R4
M1R5
M1R6B
M1R7
M1R8
M1R9B
M1R10B
M2R1
M2R2
M2R3
M2R4
M2R6
M3R1
M3R2
M3R3
M3R4
Al
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
Geoscience and Natural Resource Management, University of Copenhagen, at temperatures between 980 and 1100 °C, aiming for beam intensity at atomic mass unit (AMU) 52.9407 of 30–60 mV. Every load was
analyzed two to four times (see Table 3). Titanium, vanadium and
iron interferences with Cr isotopes were corrected by comparing with
49
Ti/50Ti, 50V/51V and 54Fe/56Fe ratios. The final isotope composition of
a sample was determined as the average of the repeated analyses and
reported relative to the certified SRM 979 standard as
53
δ Crð‰Þ ¼
h
53
221
ultramafic rocks (Table 1), probably reflecting the contamination by applied synthetic fertilizers (Kabata-Pendlas, 2000).
4.2. Cr-host minerals
The study by XRD and SEM/EDS revealed that the dominant minerals in soil are quartz, calcite, silicates (serpentine, olivine and chlorite),
chromite, ferrian-chromite, hematite, magnetite and Cr-bearing goethite while montmorillonite, Fe-sulfides and zircon occur in lesser
amounts. A portion of the chromium in soils from both central Euboea
and Avlona is hosted in chromite grains or fragments, in Cr-bearing goethite and in silicates (Fig. 2) transported as detrital components originally derived from the weathering of the ophiolitic parent rocks and
Ni-laterite deposits.
i
52
53
52
Cr= Crsample = Cr= CrSRM979 −1 1000:
Repeated long-term analysis of 0.5 μg loads of unprocessed double
spiked NIST SRM 979 standard yield an average δ53Cr value of 0.08 ±
0.05‰ (n = 245; 2σ; 52Cr signal intensity at 0.4 V) on the “Phoenix”
TIMS which we consider as a minimum external reproducibility for a
sample reproducibility, including separation procedure, double spike
correction error, and respective internal analytical errors.
4.3. Trace element concentrations in groundwater
The majority of groundwater samples from domestic and irrigation
wells throughout Central Euboea (Messapia) and Assopos basin (Avlona)
exhibit concentrations exceeding the maximum acceptable level for
Crtotal in drinking water (50 μg/L, according to the EU Directive; (EC,
1998)). At the Avlona area (Assopos), the total chromium concentrations
range from 50 μg·L−1 to 93 μg·L−1 and for waters from Messapia from
41 to 230 μg·L−1 (Table 2). With the exception of total Cr all the other
elements in the groundwater samples were found to have concentrations below the maximum permissible limits for human usage (EC,
1998). However, there is a significant variation in the concentrations of
several elements, such as As, ranging from b0.5 to 3.5 μg/L, U from 0.02
to 3.3 μg/L, Ni from 0.2 to 15 μg/L, Mn from b 0.05 to 5.6 μg/L, Cu from
0.6 to 4 μg/L and in the Ca, Mg, Na, Si and B ones (Table 2). Since Ca,
Mg, Na, Si and B are common components of water, rocks and seawater, the plot of Mg/Si ratio versus Cr(VI) is given to discriminate potential sources (Fig. 3a). Values are compared to those from the entire
central Euboea as well as from Mg-rich waters from Italy with a similar
geological setting where they are affected by the interaction with ultramafic rocks (Fig. 3b; Fantoni et al., 2002; Megremi et al., 2013). This
was done to evaluate the importance of water–rock interaction vs. seawater contribution into the aquifer. In addition, the plot of Mg/Si ratios
4. Geochemistry of rocks, soils and groundwater
4.1. Major and trace element contents in rocks and soils
The rock samples were collected from outcrops along the national
road from Makrymallis toward Kondodespoti, and from outcrops east
of the Psachna area (Fig. 1). The peridotite samples exhibit varying degree of serpentinization. Although they are generally highly tectonized,
some less altered parts still remain. Major and trace element concentrations are presented in Table 1. As expected, the Cr, Ni, Co, Mn and Fe
contents are elevated in the ultramafic rocks while the relatively high
Zr, Y, Li, K and Ca, and low Mg contents in some samples reflect the
strong serpentinization and alteration of the peridotites. The soil samples, collected from the neighboring cultivated area of Messapia are
characterized by significant Cr, Ni, Mn, Fe and Co contents, reflecting
the contribution of host ultramafic ophiolitic rocks (Table 1) and are
comparable to those given for the area of Avlona (Atsarou and
Economou-Eliopoulos, 2012). Also, a salient feature is the relatively
high B, P, K, Ba, U, Th, Nb, Li, Zr, and Y contents in soils compared to
Table 3
Field paramenters, total chromium, Cr(VI) and δ53Cr values for groundwater from wells and leachates (R.L. and S.L.).
Sample
Χ
xΥ
Depth (m)
Crtotal
−1
d53Cr (per mil)
2SE (abs)
n
pH
Eh (mV)
μg·L
35
20
22
11
20
18
27
30
230
41
77
102
48
65
65
45
230
40
77
102
48
62
63
43
249
42
73
117
47
64
63
40
0.98
1.42
0.84
1.22
1.98
1.41
1.76
1.34
0.08
0.05
0.01
0.07
0.05
0.04
0.08
0.06
2
3
3
4
5
3
5
3
7.3
7.47
7.2
7.26
7.72
7.48
7.3
7.3
−28
−35
−23
−26
−52
−38
−19
−19
23°41′10″E
23°42′7″E
120
80
50
93
48
85
53
86
0.98
1.03
0.03
0.01
3
3
7.36
7.54
−26
−34
38°37′47″N
38°37′47″N
38°37′47″N
23°39′0.5″E
23°390.5″E
23°39′0.5″E
R. L.
R. L.
R. L.
35
64
30
35
63
29
38
66
39
0.56
0.86
0.96
0.04
0.06
0.06
3
4
4
8.07
8.1
8.3
−92
−87
−102
38°34′52″N
38°34′54″N
38°34′52″N
23°39′42″E
23°40′19″E
23°40′49″E
S. L.
S. L.
S. L.
38
76
35
34
68
35
40
76
40
0.59
0.51
0.33
0.05
0.08
0.06
4
5
4
7.96
8.14
7.86
−65
−72
−68
38°34′13″N
38° 36′28″N
38°34′13″N
38°34′11″N
38°34′13″N
38°33′49″N
38°35′21″N
38°35′21″N
23° 40′09″E
23°40′03″E
23°37′48″E
23°37′35″E
23°37′32″E
23°37′02″E
23°41′15″E
23°41′15″E
Avlona
AVLO13W
AVLO14W
38°16′12″N
38°16′01″N
S.L.
MSS4B
MSS6B
MSS7B
Crtotal D DS
−1
μg·L
Wells
Euboea
E7
E18
E2
E5
E10
Ε13
MSW8
Μ3W8
Leachates
R.L.
M1R2BA
M1R2BB
M1R3
Cr(VI)
Euboea
Symbols: R.L. = ultramafic rock leachates; S.L. = soil leachates.
222
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
a
b
c
d
Fe-chr
goeth
chr
chr
Fig. 2. Selected backscattered electron (BSE) images from central Euboea (Messapia) soils, showing Cr-hosts: fragments of chromite (panels a, d), serpentine (panels a, b) and goethite
(panel d). Abbreviations chr = chromite; serpentine = srp; goeth = goethite.
vs. Na concentration (Fig. 3c) shows that elevated Mg/Si ratios are accompanied by an increase of Na concentration. The range of pH (from
7.2 to 7.5) and Eh (from − 52 to − 19 mV) values measured in the
groundwater (Table 3) indicate slightly alkaline and almost neutral
redox conditions.
4.4. Trace element concentrations in rock and soil leachates
The results of leaching experiments of variously serpentinized peridotite samples and soils under atmospheric conditions are contained in
Table 2. In general, the Cr(VI) concentration in leachates from peridotites is much lower than those in soils (Table 2). Also, the highest concentrations of Cr in rock leachates were measured in those samples
which exhibit a strong degree of serpentinization. The Cr(VI) concentrations in soil leachates, ranging from 17 to 87 μg·L−1, show a positive
correlation with concentrations of the total Cr, Fe, Mn and Co in the corresponding soils (Fig. 4). The Mg/Si ratio in leachates from less altered
peridotites is higher than those from highly serpentinized ones while
the Cr(VI) concentrations are much lower in the former than in the latter (Fig. 3a). In addition, the soil leachates are characterized by both relatively high Mg/Si ratio and Cr(VI) concentrations, but the Mg/Si ratios
are lower than those in groundwater.
4.5. Chromium stable isotope values in groundwater
δ53Cr values in groundwater from central Euboea, which is an area
dominated by Cr-bearing peridotites and Fe–Ni laterite deposits (potential sources for chromium contamination by natural processes), and
from the Assopos basin (Avlona area) with a strong industrial impact
and natural contamination as well (Economou-Eliopoulos et al., 2013)
are listed in Table 3. There are significant variations in δ53Cr values. Isotopic signatures range from 0.84 to 1.98‰ in groundwater samples from
Euboea, and from 0.98 to 1.03‰ in samples from the area of Avlona
(Assopos basin). The highest Cr(VI) concentration in groundwater
(230 μg·L−1), at the same time characterized by a relatively low δ53Cr
value (0.98‰), was measured in a sample from a shallow well in the
Psachna area, which is dominated by alluvial sediments. The five samples with the lowest Cr(VI) concentrations (average of 51.2 ±
10.1 μg·L− 1) are characterized by elevated δ53Cr values (average
1.58 ± 0.28‰) instead.
4.6. Chromium stable isotope values in leachates
Highly serpentinized peridotites and soil samples which yielded
significant Cr(VI) concentrations in their leachates were also analyzed for their Cr isotope composition. The measured δ53Cr values
for the highly serpentinized peridotite leachates range from 0.56 to
0.96‰ and those for the soil leachates range from 0.51 to 0.59‰
(Table 3). Although there is no clear-cut relationship between
Cr(VI) concentrations and δ53Cr values, there is a tendency, however
that rock leachates yield at average higher δ 53 Cr values (mean of
~ 0.8‰, at lower 4.6 wt.% Fe) than soil leachates (mean ~ 0.5‰, at average 7.3 wt.% Fe), while mean Cr(VI) concentrations in the leachates
remain similar (42 and 45 μg·L− 1 , respectively). Such a negative
trend is also suggested by the plot of δ53Cr values in the rock and
soil leachates versus Fe content in the corresponding rock and soil
samples (Fig. 6).
5. Discussion
The occurrence of Cr(VI) in ground and surface waters has been previously reported (Fantoni et al., 2002; Megremi, 2009, Oze et al., 2004,
2007; Raddatz et al., 2011; Villalobos-Aragón et al., 2012; Wanner
et al., 2012). The ratio of Cr(VI) to Cr(total) ranges from 0.9 to 1 and
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
223
10.00
Groundwater
L rocks Cr(VI)< 5 μg/L
L rocks Cr(VI) 5 - 63 μg/L
Mg/Si ratio
equal to 2.2
Mg/Si
L soil Cr(VI) 17 - 87 μg/L
1.00
a
0.10
1
10
100
1000
Cr(VI) (μg/L)
100,00
Selective data from Fantoni et al., 2002
Central Euboea Cr<5μg/L
Central Euboea Cr from 5 to 50μg/L
Central Euboea Cr>50μm/L
Mg/Si
10,00
Mg/Si ratio
equal to 2,3
1,00
b
0,10
0,1
1
10
100
1000
Cr(total) (μg/L)
Fig. 3. Plot of the Mg/Si ratio versus Cr showing trends and variations of the chemical components of the groundwater and rock, soil leachates from central Euboea, in comparison with
published data (the dashed line corresponds to the value of Mg/Si ratio equal to 2.3–2.2). (Data: Table 3 (panel a) and Megremi et al., 2013; Fantoni et al., 2002; panel b).
the very good correlation (r2 = 0.99) between Cr(total) and Cr(VI) implies that Cr(VI) is the predominant Cr species in the waters from the
areas studied (Megremi, 2009). The issue of contamination by heavy
metals, including Cr(VI), is a complex and politically delicate one,
because there is often no clear-cut answer to the question regarding
the ultimate sources responsible for a contamination. Strongly positively fractionated Cr(VI) is indicative of mass-transfer processes involving
reductive processes, and therefore stable Cr isotopes (δ53Cr values)
have been proposed as a tool for tracking Cr(VI) migration in groundwater (Berna et al., 2010; Blowes, 2002; Ellis et al., 2002; Halicz et al., 2008;
Izbicki et al., 2008; Jamieson-Hanes et al., 2012; Schoenberg et al., 2008;
Sikora et al., 2008; Zink et al., 2010). Sources of Cr used for industrial
purposes have δ53Cr values close to 0‰ relative to NIST 979 (Ellis
et al., 2002; Schoenberg et al., 2008), while naturally occurring Cr
in groundwater displays a range of δ53Cr: values from +1.0 to +5.8‰
(Ellis et al., 2002; Izbicki et al., 2008; Novak et al., 2014). Such strongly
positively fractionated values reflect reduction of Cr(VI) (after initial oxidative mobilization) during transportation in the aquifer.
The chromium isotope tracing technique has been applied in
Hinkley California (USA), at the Pacific Gas & Electric (PG&E) Compressor Facility, where a groundwater was contaminated by anthropogenic
chromium. The δ53Cr values identified in groundwater samples from a
pilot study carried out at Hinkley (CH2MHill 2007) have been used to
assess the Cr contamination source and to delineate redox processes
(Izbicki et al., 2008) within the aquifer. On the basis of preliminary laboratory experiments these authors determined the variability of kinetic
isotope effects and Cr(VI) reduction, and concluded that industrial
Cr(VI) supplies probably have Cr isotope compositions close to those
of the Earth's mantle. However, δ53Cr values of industrially contaminated waters in the Czech Republic and Poland are positively fractionated
relative to the pollution source, as a result of Cr(VI) reduction in the
water sheds (Novak et al., 2014).
224
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
100
L rocks
L soils
90
b
90
80
70
Cr(VI) (μg/L)
Cr(VI) (μg/L)
80
100
a
60
50
40
30
70
60
50
40
30
20
20
10
10
0
0
0
1000
2000
3000
4000
0
500
Cr (mg/kg)
c
1500
2000
d
100
90
90
80
80
Cr(VI) (μg/L)
Cr(VI) (μg/L)
100
1000
Mn (mg/Kg)
70
60
50
40
70
60
50
40
30
30
20
20
10
10
0
0
0.0
5.0
10.0
15.0
0
50
Fe (wt%)
100
150
200
Co (mg/Kg)
Fig. 4. Plots of Cr(VI) in rock (L rocks) and soil (L soils) leachates versus Cr, Mn, Fe and Co contents in ultramafic rocks and soils, respectively (data from Tables 1 and 2).
5.1. Use of δ53Cr values as a tracer for the reduction of Cr(VI) in natural
waters
In order to evaluate the efficiency of natural attenuation of the dissolved and toxic hexavalent Cr(VI) to less harmful Cr(III) in the groundwater of the studied basin on Euboea area, in our calculation we used a
Rayleigh distillation model, assuming the geogenic background composition to be represented by the average δ53Cr value of 0.64‰ defined by
the soil and rock powder leachates presented herein. This value is
higher than the δ53Cr value of 0.37‰ reported by Ellis et al. (2002) for
Cr(VI) solutions from plating baths (industrial contaminant) and potentially implies that, if the Cr(VI) signatures were purely produced by mixtures of geogenic and industrial Cr(VI), the anthropogenic Cr(VI)
contaminant would have to have δ53Cr values drastically exceeding
those reported by Ellis et al. (2002). We therefore prefer a scenario by
which the measured δ53Cr values of aquifers studied herein reflect a residual, partially reduced geogenic Cr(VI) pool.
Under this assumption, we computed the expected changes in the Cr
isotope composition of dissolved Cr(VI) species in the affected waters,
as a function of the progressive reduction of Cr(VI) to Cr(III). In our calculation, we used the average δ53Cr signature of the soil and rock pulp
leachates as the local geogenic Cr composition. The δ53Cr signature of
the dissolved Cr(VI) species that remain in waters (the reactant pool)
at any given time as back-reduction to Cr(III) proceeds was calculated
based on the following Rayleigh relation (cf. Ellis et al., 2002; Johnson,
2011):
53
δ Cr ¼
h
i
53
ðα−1Þ
−1000
δ Cr0 þ 1000 f
where δ53Cr and δ53Cr0 represent the isotope compositions of the
unreacted dissolved Cr(VI) in the run-off at the site of the Cr source at
the given time (i.e., sampling time) and at the initial stage when the reaction started (t = 0), respectively. The parameter f is the fraction (in %)
of the unreacted Cr(VI) remaining in the groundwaters, and α represents an isotope fractionation factor associated with the Cr(VI) reduction, defined as:
α ¼ RPROD =RREACT
where RPROD and RREACT are the 53Cr/52Cr isotope ratios of the reaction
product, Cr(III), and the reactant (the Cr(VI)), respectively. The relative
isotope difference Δ53/52Cr between these two, i.e. oxidized and reduced, water soluble chromium pools can be calculated according to
the equation:
53
53
53
Δ CrðPROD–REACTÞ ¼ δ CrPROD −δ CrREACT
and/or approximated through the isotope fractionation factor α using
the following relation:
53
1000 ln α Δ CrðPROD−REACTÞ :
We used a range of isotope fractionation values (α) associated with
the abiotic Cr(VI) reduction by magnetite (α = 0.9965; Ellis et al., 2002;
Zink et al., 2010), other Fe(II)-bearing phases (α = 0.9979–0.9961;
Basu and Johnson, 2012) and aqueous Fe (II) (α = 0.9970–0.9958:
Døssing et al., 2011), Kitchen et al., 2012), which corresponds to about
2.1 to 4.2‰ lighter 53Cr/52Cr ratio in the reaction product, i.e. Cr(III),
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
compared to that of the reactant Cr(VI) pool. This is a simplistic assumption since it neither does take potentially biotic (microbial) reduction
mechanisms into consideration, nor does account for the potential presence of a heterogeneous aquiver with multiple reductants. However,
under basic pH such as characterizing the rock leachates (average 8.2)
and aquifers (7.37 ± 0.16) studied herein (Table 3), the reduction of
Cr(VI) by organic reductants is considered minimal, and the reduction
of Cr(VI) by Fe(II) most likely was the predominant reduction mechanisms. For our calculation, we also assume that adsorption/desorption
of Cr(VI) was insignificant (Ellis et al., 2004), and that, consequently,
the dissolved Cr(VI) we measured was representative of the total dissolved Cr(VI) pool, in terms of both isotopic composition and extent of
reduction.
The results of our numerical modeling, assuming the Cr isotope composition of the local geogenic Cr source to be represented by the average
of δ53Cr = 0.64 +/0.47 (2σ; Table 3) (i.e., reflected in δ53Cr of an initial
unreacted Cr(VI) pool, i.e. f = 1), suggest that more than ~53%, but maximum 96%, of the original Cr(VI) pool was reduced to Cr(III) in the waters investigated (Fig. 5 a and b). This implies that there is an ongoing
and relatively efficient process in the basin aquifers studied that
225
facilitates natural attenuation of the dissolved and toxic hexavalent
Cr(VI) to less harmful Cr(III).
5.2. The use of δ53Cr values in rock and soil leachates for identifying
reduction of Cr(VI)
The combination of trace element data with δ53Cr values of rock and
soil leachates from central Euboea (Tables 1–3) and their comparison to
δ53Cr values for geogenic and anthropogenic waters from central
Europe recently published by Novak et al. (2014) may contribute to
the identification of contaminant Cr sources. The measured higher
Cr(VI) concentrations in rock leachates from rocks with higher Mn
contents compared to less altered peridotites, and the higher Cr(VI)
concentrations in soil leachates compared to rock leachates (Tables 1
and 2) in general seem to be consistent with the common occurrence
of ferrian chromite (FeCr2O4) and manganese oxides in the soils
(Figs. 2b and 3b). These phases have a catalytic control over Cr(III) oxidation. In addition, the oxidation of Cr(III) to Cr(VI) in the cultivated
soils of central Euboea may be facilitated by atmospheric oxygen
(which in turn oxidize the Mn2 + to produce Mn4 + catalysts), as a
3.5
α = 0.9979
3.0
53
δ Cr (permil)
2.5
2.0
1.5
1.0
0.5
53%
a
91%
0.0
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1.0
f
7.0
6.0
α = 0.9956
53
δ Cr (permil)
5.0
4.0
3.0
2.0
1.0
74%
b
96%
0.0
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1.0
f
Fig. 5. Results of a theoretical Rayleigh modeling for the quantitative estimates of the amount of hexavalent chromium Cr(VI) reduction to trivalent Cr(III) in groundwaters from Euboea,
assuming the geogenic background composition to be represented by the average 53Cr value of 0.64‰ defined by the soil and rock powder leachates presented herein. Dashed vertical lines
with the percentage numbers illustrate the minimum calculated amounts of Cr(VI) that was reduced, and suggest that more than ~53%, but maximum 96%, of the original Cr(VI) pool was
reduced to Cr(III) in the waters investigated (panels a, b). We used a range of isotope fractionation values (δ) associated with the abiotic Cr(VI) reduction by aqueous Fe(II) and Fe(II)bearing minerals (Ellis et al., 2004).
226
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
1.2
1
53
δ Cr (per mil)
point to the fact that oxidation of chromium has been facilitated considerably by Fe. In addition, in a plot of δ53Cr values vs Cr(VI) concentrations for studied waters (Table 3), supplemented with average values
for geogenic waters from central Europe (Novak et al., 2014), it is apparent that geogenically contaminated waters define separate data arrays
which are different from that defined by anthropogenically contaminated waters by exhibiting generally lower Cr(VI) and somewhat less positively fractionated δ53Cr values, while still showing the significant
negative trend to be expected by reduction processes (Fig. 7).
L rocks
L soils
0.8
mean values
0.6
0.4
0.2
0
0
2
4
6
8
10
6. Conclusions
Fe (mg/Kg)
Fig. 6. Diagram of δ53Cr values in rock (L rocks) and soil (L soils) leachates versus Fe contents in ultramafic rocks and soils, respectively (data from Tables 1–3).
consequence of often applied soil mixing and soil turnovers during agricultural plowing, in contrast to in situ outcrops of peridotites. Lastly,
Cr(VI) can be readily reduced in situ again to Cr(III) by aqueous Fe(II)
or Fe(II)-bearing minerals (Ellis et al., 2002) and/or by bacteria (Sikora
et al., 2008).
Since oxidation and reduction of chromium occur simultaneously in
nature, it has been suggested that a potential limitation to the use of natural attenuation of Cr(VI) depends on the oxidation capacity of the soils
(Stanin, 2005). Also, due to the high redox potential of the Cr(VI)/Cr(III)
couple, the only oxidants present in natural systems that are capable of
oxidizing Cr(III) to Cr(VI) are considered to be manganese oxides
[Mn(IV/III)] and dissolved oxygen (Eary and Rai, 1987, 1988, 1989;
Oze et al., 2007). However, Palmer and Wittbrodt (1994), based on experimental work, concluded that under slightly acidic to basic conditions, due to kinetic reactions, sorption and/or precipitation of Cr(III)
is much faster than oxidation, and hence the oxidation of Cr(III) by dissolved oxygen is an unlikely process. With respect to the redox reactions between chromium and iron species, which involve very
complex processes (Palmer and Puls, 1994), it has been emphasized
that the reduction of Cr(VI) by Fe(II) is 100 times faster than the reduction rate by organic matter (Wielinga et al., 2001). In addition, a significant amount of Cr(VI) can potentially be reduced, due to the rapid
cycling of Fe(II) back to Fe(III) (Stanin, 2005). The negative trend between δ53Cr values in the rock and soil leachates, and respective Fe concentrations in peridotite and soil samples (Tables 1 and 2; Fig. 6), may
The compilation of trace element data on groundwater, ultramafic
rocks and soil samples from Central Euboea, and δ53Cr values in representative groundwater, rock and soil water leachates, led to the following conclusions:
• The higher Cr(VI) concentrations in soil water leachates compared to
those of rock powder leachates can be explained by increased oxidation capacities in the presence of Fe(II) hydroxides and Mn oxides.
• Although the dominant cause for Cr isotope fractionation (δ53Cr
values ranging from 0.56 to 0.96‰ in rock leachates and from 0.51
to 0.59‰ in the soil leachates) is reduction, processes other than reduction, such as sorption, precipitation and uptake by plants may
complicate the interpretation of the observed δ53Cr values.
• There is a significant variation in δ53Cr values, ranging from 0.84 to
1.98‰ in groundwater samples from Euboea and from 0.98 to
1.03‰ in samples from the area of Avlona (Assopos basin). Assuming
the geogenic background composition to be represented by our experimental leachates of soils and rock powders, the elevated δ53Cr
values potentially imply reductive processes during transport of the
mobilized Cr(VI) in the different aquifers investigated.
• Using a range of different fractionation factors valid for aqueous Fe(II)
and Fe(II)-bearing mineral reduction, and a geogenic δ53Cr value of
~ 0.64‰ for an initial geogenic aquifer composition deduced from
the leaching experiments, we calculate, using a Rayleigh distillation
model, that is between 53% and 96% of the original Cr(VI) pool was reduced to Cr(III) in the waters investigated.
• This implies that there is an ongoing and relatively efficient process in
the groundwater studied that facilitates natural attenuation of the dissolved and toxic Cr(VI).
Water-Gr
4.5
L rocks-Gr
4
L soil-Gr
Geo-water-C.E.
Antro-water-C.E.
3
2.5
2
53
δ Cr (per mil)
3.5
1.5
1
Reduction
0.5
0
1
10
100
1000
10000
100000
Cr(VI) (μg/L)
Fig. 7. Diagram of δ53Cr values in water, rock (L rocks) and soil (L soils) leachates from Greece, (Gr), and water contaminated by geogenic (geo-water)/anthropogenic (anthro-water) in
Central Europe (C. E.) versus Cr(VI) concentrations. Data from Table 3 and from Novak et al. (2014).
M. Economou-Eliopoulos et al. / Catena 122 (2014) 216–228
Acknowledgments
The Mayor and the Municipality of Messapia–Dirfis is acknowledged
for the financial support of this work (A.K. 70/3/11730). Mr. E.
Michaelidis, University of Athens, is thanked for his assistance with
the SEM/electron probe analyses. We are thankful for the help of Toni
Larsen with ion chromatographic separations and thank Toby Leeper
for always maintaining the mass spectrometers in perfect running
conditions. Financial support through The Danish Agency for Science,
Technology and Innovation grant no. 11-103378 to RF and through
the Danish National Research Foundation center of excellence NordCEE
(DNRF grant number DNRF53) is highly appreciated. Constructive comments by Christopher Oze and an anonymous reviewer helped to improve the initially submitted manuscript.
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