Ecotoxicology and Environmental Safety 54 (2003) 65–73
Mobility and toxicity of metals in sandy sediments deposited on land
Z. Prokop,a,* M.L. Vangheluwe,b P.A. Van Sprang,b C.R. Janssen,c and I. Holoubeka
a
Department of Environmental Chemistry and Exotoxicology, Masaryk University, Kotlarska 2, 602 00 Brno, Czech Republic
b
EURAS Grote Baan, 199,9000 Ghent, Belgium
c
Laboratory of Environmental Toxicology and Aquatic Ecology, Ghent University, J. Plateaustraat 22, 9000 Ghent, Belgium
Received 24 April 2001; received in revised form 19 March 2002; accepted 30 May 2002
Abstract
A times series of laboratory experiments were conducted to investigate the effect of land deposition of contaminated sediments on
the bioavailability and mobility of metals. Four sandy sediments were sampled at sites expected to have elevated levels of cadmium
and zinc. The physical and chemical characteristics and ecotoxicity of sediments, pore waters, and leachates were evaluated after
periods ranging from 1 to 45 days of land deposition. Cd and Zn retardation and leaching potential were calculated and this
simulation gave good predictions of subsequently observed Cd and Zn mobility. The mobility and leaching of Cd and Zn in the
sediments increased with decreasing pH and with decreasing content of organic matter. During the deposition an increase in
sediment toxicity to plants and an increase in eluate toxicity to invertebrates were observed. A high rate of water flow through the
sediment resulted in a lower toxicity enhancement of the sediments and a higher toxicity enhancement of the eluates. This result
suggests that water flow through the sediment reduces the actual toxicity of the upper layer of deposited sediment but at the same
time intensifies the risk of groundwater contamination.
r 2002 Elsevier Science (USA). All rights reserved.
Keywords: Bioavailability; Deposition; Dredged sediments; Metals; Mobility
1. Introduction
Human activity promotes the accumulation of contaminated sediments in water courses. Increased maintenance dredging results in large amounts of
contaminated solids for which safe disposal sites have
to be found. Frequently, dredged sediments are deposited on land immediately next to dredged waterways.
This practice alters the physicochemical characteristics
of the sediments and may result in the release and
transport of codeposited contaminants into soil and
groundwater. Hence, the fate and effects of sediment
metals after land disposal are of particular concern.
The mobility and bioavailability of metals bound to
sediments depend on multiple factors, with sediment
characteristics and the physical–chemical form of the
metal being the key factors. Generally, free metal ions
are the most mobile and the most bioavailable form
(Janssen et al., 1997a, b; Sauve! et al., 1998). Transfor*Corresponding author. Fax: 54-112-9506.
E-mail address: [email protected] (Z. Prokop).
mation of the different metal forms and alteration of the
sediment will be induced by a change in the environmental conditions of sediments following deposition on
land. For example, a decrease in pH (e.g., an effect of
acid rain) may cause a release of metals from complexes
and from solid matter surface by increased competition
for sorption sites by the H+ ion. A second example is
the mobilization of metals, originally bound to sulfides
under anoxic conditions, that occurs after the introduction of oxygen into the land-deposited sediment.
Oxidation of organic matter also increases under aerobic
conditions (Fu et al., 1992). Oxidation of organic matter
is considered one of the most important mechanisms
inducing mobilization of metals (Benninger-Truax and
Taylor, 1993). The oxidation of sulfides and organic
matter may also generate, if the buffer capacity of the
receiving environment is not sufficient, acidic conditions, which may provoke increased mobility of some
metals. However, the mobile fraction does not necessarily correspond to the bioavailable fraction. Ion pairs,
complex ions, polymers, and microparticulates, as well
as sorption on solid surfaces and biological surfaces,
0147-6513/03/$ - see front matter r 2002 Elsevier Science (USA). All rights reserved.
doi:10.1016/S0147-6513(02)00022-2
66
Z. Prokop et al. / Ecotoxicology and Environmental Safety 54 (2003) 65–73
reduce the activity of the free ionic form of the metals
and, hence, the potential for exerting toxicity (Roy and
Campbell, 1997).
In this study, a time series of laboratory experiments
were conducted to investigated the effects of land
deposition of metal-contaminated sediments. The main
study objectives were (1) the identification and assessment of the effect of land deposition on the mobility and
leaching of metals from dredged sediments, and (2) the
evaluation of changes in the ecotoxicity and availability
of codeposited metals.
2. Materials and methods
2.1. Sample collection and handling
Surface sediments were collected from the brook
Scheppelijke Nete situated in the north of Belgium. The
collection sites were selected to span a gradient of
contamination characterized by elevated levels of
metals, especially Cd and Zn. In total, four stations
located in depositional areas were sampled: SN 1 and
SN 2 in the southern arm of the Scheppelijke Nete
suspected of being the most heavily contaminated; SN 3
in the northern arm of the Scheppelijke Nete; and SN 4,
1 km downstream from the conjunction of the northern
and southern arms. Sediment samples at each location
were collected with a Van Veen grab (sampling depth
was approximately 10 cm) and were stored at 4 C
until use.
Exposure of sediments to terrestrial conditions was
simulated in the laboratory using leaching chambers
consisting of two polyethylene trays. The upper tray
(2-L volume) contained the sediment (1.75 L and 10 cm
thick). The bottom of this tray was perforated to allow
the eluate to drain. A 1-cm sand layer prevented leakage
of the sediment particles into the elaute. The bottom
tray was used to collect the eluate. The surface of the
sediments was sprinkled with distilled water twice a day
to simulate rainfall. The amount of ‘‘artificial rain’’
corresponded to average Belgian conditions (700 mm of
rainfall per year). The leaching chambers were not
covered and were exposed to the air under controlled
temperature conditions (2072 C). A set of eight
leaching chambers were set up for each sediment sample,
except for sediment sample 4. Each chamber corresponded to one exposure period (1, 2, 3, 5, 7, 14, 28, and
45 days). Sediment sample 4 was investigated only after
1 and 45 days since this sample was suspected to be the
least contaminated (based on previous measurements).
2.2. Physical and chemical analyses
The sediments were characterized for general parameters such as clay content, pH, acid volatile sulfide,
and total organic carbon (TOC). pH (KCl) was
measured (Consort pH meter) at a 1:2.5 soil:liquid ratio
with 1 M KCl according to ISO (1994). Dry bulk and
particle density of sediments were measured according
to Culley (1993) and total porosity according to Carter
and Ball (1993). Water holding capacity (WHC) was
determined by measuring the water content of the soil
after inundating it for 3 h in water and subsequently
draining it for 2 h (ISO, 1996). The clay fraction
(o2 mm) was determined by measuring the sedimenta.
tion velocity by the pipet method of Robinson–Kohn,
and organic matter content was estimated from the
carbon content (multiplying by a factor of 1.7), which
was obtained by oxidation with potassium dichromate
in sulfuric acid. Acid volatile sulfides (AVS) were
measured according to a spectrophotometry method as
outlined by Allen et al. (1993).
Zn and Cd concentrations in eluates, pore waters, and
sediments were measured by atomic adsorption spectrometry using a Varian AA-100 spectrophometer (Varian
Instruments, Canada) according to ASTM (1973).
Samples for determination of the dissolved metal
concentrations were filtered at 0.45 mm. The eluates
and pore waters were acidified to pH 2. Total sediment
metal concentrations were obtained after microwave
digestion of the dry sediment in a mixture of concentrated HNO3, HCl, and deionized water (4:1:1, v/v/v).
Chloride was determined using a Spectroquant 14755
Kit, pore water ammonium concentrations were measured using a Spectroquant 14752 Kit, and sulfate was
analyzed using a Spectroquant 14791 Kit (Merck
KGaA, Germany). The hardness of eluates and/or pore
waters was measured using an Aquamerck 1.11104.0001
Kit (Merck KGaA, Germany). TOC in the pore water
and elaute was measured using the TOC Test LCK
383-4 (Dr. Lange, Germany).
2.3. Ecotoxicity tests
Toxicity tests with the eluates were performed using
the freshwater crustacean Thamnocephalus platyurus.
This cyst-based toxicity test was performed following
the procedure described by Centeno et al. (1995).
Toxicity tests were conducted in 24-well polystyrene
test plates at 2371 C and a photoperiod (L:D) of 16:8.
Each concentration consisted of three replicates, with 10
juveniles in 1 mL test solution. Mortality was scored
after the 24-h exposure period. The plant growth
inhibition test was performed according to OECD
Guideline 208 with Lolium perene (ryegrass, cat. 1) and
Raphanus sativum (radish, cat. 2). The plants were
maintained at 20 C (72 C) and 12:12 L:D cycle at
6000 lx. For each sediment four replicas were used
consisting of 50 g (wet wt) soil and 5 seeds per replicate.
Tests were terminated 14 days after 50% of the control
seedlings had emerged. The number of emerged plants
Z. Prokop et al. / Ecotoxicology and Environmental Safety 54 (2003) 65–73
was recorded and the average dry weight of the
harvested plants was measured. The 14-day mortality
test with Enchytraeus albidus was used to assess
the toxicity of the treated sediments to invertebrates
.
(Rombke
et al., 1998). Ten adult enchytraeids were
exposed in 20 g of soil in covered glass vessels (three
replicates). During exposure, vessels were kept at
2071 C and a 16:8 L:D cycle. Soil moisture content
was adjusted twice a week by replenishing weight loss
with the appropriate amount of deionized water. After
the 14-day exposure period the surviving animals were
counted.
2.4. Data analysis
The percentage effect (e.g., mortality or inhibition)
was calculated when undiluted samples were tested.
Significant differences (Po0:05) of mean survival/
growth were tested using a one-way analysis of variance
(ANOVA) in combination with Duncan’s multiple
range test. Data for percentage survival were arcsine
square root transformed prior to analysis and tested for
normality and homogeneity of variances using the
Kolmogorov–Smirnov and Barlett tests. Arcsine square
root transformed data fulfilled the assumptions of
parametric statistics. LC50 values and corresponding
95 confidence intervals were calculated with the moving
average method.
The chemical speciation of Cd and Zn in the
sediments was calculated using the Windermere humic
aqueous model (WHAM-SOIL) (Tipping, 1994). Zinc
and cadmium partitioning coefficients (Kd ) were determined as the ratio of metal concentration in bulk
sediment (Cs ) to the concentration of metal dissolved in
pore water (Cw ):
Kd ¼ Cs =Cw :
ð1Þ
The theoretical rate at which a sorbing chemical can
move through the sediment (vmþ ) is equal to the seepage
velocity (v) divided by the retardation factor (R). The
retardation factor, describing behavior of chemical
sorption and possibility of transport, is defined by the
67
equation (Hemond and Fechner, 1994)
ð2Þ
R ¼ 1 þ Db =St Kd ;
were Db is bulk density and St is the total porosity of the
sediment. The potential amount of Cd and Zn released
from sediment by the flow of water was calculated as
Mtot ¼ tvmþ As Cs Db ;
ð3Þ
where t is time, As is the area of a cross section of the
sediment column, Cs is the metal concentration in the
sediment, and Mtot is the potential amount of Cd and
Zn that can be released from sediment.
The rate of metal leaching from the sediments was
calculated as the ratio between its concentration in the
eluate and its concentration in the bulk sediment,
ML ¼ CL =CS ;
ð4Þ
where CL is the metal concentration in the leachate at
the end of the leaching test and ML is the leaching rate.
3. Results
3.1. Sediment characteristics
The physicochemical characteristics and total metal
content of the sediments prior to land deposition (Day
0) are summarized in Table 1. The low pH, ranging from
4.45 to 6, and the low content of organic matter (less
than 1.3% dry wt), as well as the low content of clay
(under 4% dry wt) and AVS, were expected to result in
high bioavailability and mobility of metals in the
sediment samples.
Metal analyses focused on Cd and Zn as these were
the main metals expected to be present in the samples.
Total metal concentrations in the sediments ranged from
70 to 285 mg/kg dry wt for Zn and from 2 to 15 mg/kg
dry wt for Cd. For Cd these values are well above the
background concentration (0.38 mg Cd/kg dry wt)
measured in uncontaminated sediments in Flanders.
The average background level of zinc in Flemish
sediments is 70 mg/kg dry wt (De Cooman et al., 1999).
Table 1
Physical and chemical characterization of the sediments
Sample
pH
Organic matter (%)
1
2
3
4
4.45
4.43
6.05
5.15
0.9
0.4
1.3
0.3
1
2
3
4
Porosity (m%)
0.40
0.38
0.45
0.41
Bulk density (g/cm3)
1.49
1.54
1.33
1.53
Clay (o2 mm) (%)
4.00
1.80
1.70
2.20
Water flow (mL/day)
3.21
7.68
13.04
17.78
SEM (mmol)
87
154
99
120
WHC (%)
32
32
39
32
AVS (mmol)
0.368
0.059
0.412
0.315
SO4 (mg/L)
101
173
108
84
SEM/AVS
235
2594
241
380
Cl (mg/L)
53
46
49
38
Z. Prokop et al. / Ecotoxicology and Environmental Safety 54 (2003) 65–73
68
3.2. Transformation experiments
concentrations in bulk sediments were observed during
the leaching test, while metal concentrations in leachates
increased with an increase in leaching test duration.
No significant changes in chloride concentration and
sediment and eluate pH were observed. In the leachates,
Total and dissolved Cd and Zn in sediments and
eluates were determined at Days 0, 1, 2, 3, 5, 7, 14, 28
and 45 (Table 2). No significant changes in metal
Table 2
Concentrations of Cd and Zn in sediments, pore waters, and eluates at Day 0 and during exposure of sediments to terrestrial conditions
Sample
1
2
3
4
Parameter
Time (days)
0
1
2
3
5
7
14
28
45
Total Cd in sediment (mg kg )
Cd in pore water (mg L1)
Total Cd in eluate (mg L1)
Dissolved Cd in eluate (mg L1)
Calculated Cd released (mg)
Measured Cd released (mg)
2167
7
—
—
—
—
2300
15
13
8
0.90
0.39
2233
17
22
15
1.43
0.77
2300
15
19
17
2.25
1.52
2500
14
23
11
2.49
0.92
2267
13
44
27
2.68
1.76
2067
6
45
36
2.86
3.15
2033
—
89
62
3.00
8.01
2067
—
—
—
—
—
Total Zn in sediment (mg kg1)
Zn in pore water (mg L1)
Total Zn in eluate (mg L1)
Dissolved Zn in eluate (mg L1)
Calculated Zn released (mg)
Measured Zn released (mg)
58,333
451
—
—
—
—
60,667
741
377
254
44
11
63,333
734
699
415
70
24
60,333
397
692
606
110
55
60,667
800
1076
765
122
43
65,000
826
2000
1410
131
80
56,333
1582
2040
1790
140
143
57,667
—
4128
—
147
372
83,800
—
—
—
—
—
Total Cd in sediment (mg kg1)
Cd in pore water (mg L1)
Total Cd in eluate (mg L1)
Dissolved Cd in eluate (mg L1)
Calculated Cd released (mg)
Measured Cd released (mg)
6733
204
—
—
—
—
7000
316
339
213
17.91
10.17
7333
303
525
253
28.15
18.38
9200
210
373
321
42.21
26.11
6400
268
520
370
49.40
31.20
6333
237
543
445
57.74
48.87
7633
439
718
528
63.83
93.34
5900
—
865
551
70.02
185.98
6967
—
1053
619
—
—
Total Zn in sediment (mg kg1)
Zn in pore water (mg L1)
Total Zn in eluate (mg L1)
Dissolved Zn in eluate (mg L1)
Calculated Zn released (mg)
Measured Zn released (mg)
110,667
1101
—
—
—
—
109,000
10,470
11,600
9,380
550
348
115,333
10,220
13,070
9590
865
457
106,333
8210
11,580
9810
1297
811
121,667
10,160
12,958
10,894
1518
777
107,667
9860
12.524
11,399
1775
1127
112,000
10,970
14,077
12,136
1962
1830
100,667
—
21,075
15,980
2152
4531
103,333
—
25,475
19,695
—
—
Total Cd in sediment (mg kg1)
Cd in pore water (mg L1)
Total Cd in eluate (mg L1)
Dissolved Cd in eluate (mg L1)
Calculated Cd released (mg)
Measured Cd released (mg)
3467
4
—
—
—
—
4233
12
13
11
1.19
0.65
4200
18
14
13
1.96
0.91
4100
15
13
13
2.84
1.43
4167
10
10
7
3.50
1.40
4033
11
6
8
4.30
1.40
4133
18
5
4
4.71
1.10
4033
—
17
13
5.06
6.21
5233
—
23
3
—
—
Total Zn in sediment (mg kg1)
Zn in pore water (mg L1)
Total Zn in eluate (mg L1)
Dissolved Zn in eluate (mg L1)
Calculated Zn released (mg)
Measured Zn released (mg)
321,333
247
—
—
—
—
370,667
294
753
98
27
38
366,000
418
333
120
44
22
373,333
169
123
47
64
14
359,667
239
323
81
79
45
370,667
321
193
125
97
45
375,000
—
201
56
107
44
351,000
—
544
90
115
199
448,033
—
1037
54
—
—
Total Cd in sediment (mg kg1)
Cd in pore water (mg L1)
Total Cd in eluate (mg L1)
Dissolved Cd in eluate (mg L1)
16,400
2
—
—
17,667
11
48
9
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
21,577
—
493
346
Total Zn in sediment (mg kg1)
Zn in pore water (mg L1)
Total Zn in eluate (mg L1)
Dissolved Zn in eluate (mg L1)
257,667
307
—
—
288,667
1660
709
151
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
334,667
—
15,145
12,030
1
Note.—, Not measured.
Z. Prokop et al. / Ecotoxicology and Environmental Safety 54 (2003) 65–73
69
Table 3
Physical and chemical characterization of sediments during the exposition of sediments to terrestrial conditions
Parameter
TOC (mg L
Sample
1
of leachate)
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
DOC (mg L1 of leachate)
pH sediment
pH leachate
Chloride (mg L1 of leachate)
Time (days)
1
2
3
5
7
14
28
45
12
17.2
24.8
14.8
5.6
16.5
24.8
8.3
4.7
4.7
6.7
5.8
5.6
5.5
7.7
7.2
49
36
36
23
10.6
8.54
15.8
—
8.0
8.5
15.8
—
4.9
4.9
6.7
—
6.0
5.6
8.3
—
55
40
45
—
20.3
37.3
25.1
—
10.8
8.2
22.3
—
4.7
4.7
6.6
—
5.5
5.4
8.3
—
50
39
45
—
35.4
50.6
26
—
9.1
9.0
19.5
—
4.7
4.8
6.6
—
5.0
5.2
8.0
—
54
41
44
—
46.4
61.6
26.2
—
11.7
5.8
16.6
—
4.7
4.8
6.6
—
5.1
5.6
8.1
—
69
38
40
—
39.8
125
19.5
—
10.8
6.5
18.5
—
4.6
4.7
6.5
—
5.4
5.3
8.2
—
49
39
44
—
110
152
22.4
—
7.0
10.4
11.1
—
4.6
4.7
6.1
—
5.3
5.3
7.8
—
57
37
35
—
—
68.8
78.4
113.35
—
9.4
24.8
29.2
4.2
4.7
6.5
5.5
—
5.2
7.7
6.4
—
—
27
15
Note. TOC, total organic carbon; DOC, dissolved organic carbon; —, not measured.
Table 4
Distribution coefficients, retardation factors, and rates of Cd and Zn
leaching in the sediment samples
Sample
1
2
3
4
Cadmium
Zinc
Kd
Rf
ML
Kd
Rf
ML
145
25
334
2854
10.4
7.3
9.7
15.7
0.044
0.147
0.004
0.023
80
13
1317
299
8.9
9.3
10.2
12.0
0.072
0.247
0.002
0.045
Note. Kd ; distribution coefficient; Rf ; retardation factor; ML ; rate of
leaching.
the concentration of TOC increased significantly (3–10
times), while the dissolved organic carbon concentration
of the eluate remained unchanged in sediments 1 and 3
and slightly decreased in sediment 2. A significant
increase in DOC was observed only in sediment 4
(Table 3).
The distribution coefficient (Kd ) of Cd and Zn
varied from 13 to 2900 among the different sediments
(Table 4). Lee et al. (1996) studied metal partitioning in
soil and concluded that pH and organic matter content
were the most important factors affecting Cd partitioning. The results show that Kd significantly increased
when pH increased. Soil pH seemed to be an important
factor controlling mobility of Cd and Zn in the
sediments. Leachability in sediments 1 and 2 was
expected to be higher compared with sediment 3 or 4.
Indeed, the highest concentrations of Cd and Zn were
observed in leachate from sediment 2 which had a pH of
Fig. 1. Relationship between Cd and Zn leachability and pH of the
sediments.
4.43 and Kd values of 25 and 13 for Cd and Zn,
respectively. The lowest Cd and Zn concentrations were
observed in the leachate obtained from sediment 4,
which had a pH 5.15 and Kd values of 2850 and 300 for
Cd and Zn, respectively.
The rate of leaching (Eq. (4)) ranged from 0.004 to
0.15 for Cd and from 0.002 to 0.25 for Zn (Table 4).
Fig. 1 illustrates the relationship between pH and rate of
Cd and Zn leaching and shows that the amount of metal
leaching from the dredged sediments after deposition
was larger in sediments with a low pH.
The calculation of retardation and potential leaching
of metal in sediment can predict the risk of soil or
70
Z. Prokop et al. / Ecotoxicology and Environmental Safety 54 (2003) 65–73
Table 5
Fractional distribution and speciation of Cd and Zn in the sediment
samples
Fig. 2. Comparison of measured and calculated amounts of Cd
released from sediments 1–3.
Sample
Fraction
bound on
clay (%)
Cd
1
2
3
4
93.3
58.8
23.9
51.3
Zn
1
2
3
4
93.6
60.4
23.6
51.8
Fraction
bound on
solid
organic
carbon
(%)
Fraction
bound to
dissolved
organic
carbon
(%)
Free ion
form (%)
0.2
0.5
5.5
2.0
4.3
10.7
41
13
2.0
23.3
26.6
28.0
0.3
1.1
10.5
4.0
4.2
10.7
39.5
12.8
1.9
24.0
26.8
28.7
were calculated as bound to clays, respectively. The
results of WHAM calculation also showed that no more
than 10% of total metal concentration was adsorbed to
solid organic matter. On the other hand, 4–40% of total
Cd and Zn was bound to dissolved organic carbon in the
sediment samples.
3.4. Metal ecotoxicology
Fig. 3. Comparison of measured and calculated amounts of Zn
released from sediments 1–3.
groundwater contamination before deposition of the
sediment on land. The retardation and potential
leaching of Cd and Zn were calculated based on the
metal partitioning between sediment and water on Day
0 (Table 4). Comparison of calculated and experimentally determined concentrations of Cd and Zn in the
eluate during the leaching test is illustrated in Figs. 2
and 3.
3.3. Chemical speciation
The equilibrium partitioning concept is a common
approach to determination of metal distribution in the
solid and pore water phases of soils or sediments
(Janssen et al., 1997a). However, total dissolved metal
concentrations do not necessarily correspond to the
bioavailable fraction. Ion pairs, complex ions, polymers,
or microparticulates can reduce fee ion species of metal
in solution (Green et al., 1993). According to WHAM
calculations (Tipping, 1994) Cd and Zn exhibited a
similar clay sorption behavior (Table 5). In sediments 1,
2, 3, and 4, a 93.3%, 58.8%, 23.9%, and 51.3% of total
Cd and 93.6%, 60.4%, 23.6%, and 51.8% of total Zn
The results of the ecotoxicity tests on Day 0
(evaluated immediately after sampling) and the consecutive days of the leaching test are summarized in
Table 6. Toxicity was observed at the beginning of the
experiment (Day 0) because the sediment as such was
already toxic. The toxicity of whole sediments was
evaluated using two terrestrial plant tests and one
invertebrate test with E. albidus. No toxicity was
observed for E. albidus at the start of the leaching
experiment. Slight phytotoxic effects were observed at
Day 0 for sediments 1 and 2 for the radish R. sativum
with, respectively, 25% and 20% seed germination
inhibition. For the same endpoint, no toxicity was
observed with the rye grass L. perenne. However, dry
weight analyses revealed a reduction in L. perenne
biomass (19–50%) for all sediments. For both the rye
grass L. perenne and the radish R. sativum, no significant
(Po0:05) change in toxicity (dry weight) was observed
during the exposure of sediments to terrestrial conditions simulating land deposition (see Section 2). However, a significant decrease in seed germination
(Po0:05) was observed for both plant species exposed
to sediments 1 and 2. None of the sediments were found
to be toxic to E. albidus, even after the 45 days of
exposure to terrestrial conditions (Table 6).
The T. platyurus bioassay was used to evaluate the
toxicity of the eluate. A significant (Po0:05) increase in
toxicity of the eluates as a function of transformation
Z. Prokop et al. / Ecotoxicology and Environmental Safety 54 (2003) 65–73
71
Table 6
Toxicity of the sediments from Day 0 to 45 days after the exposure to the terrestrial conditions (95% confidence intervals).
Sample
T. platyurus 24-h toxicity (toxic units)
a
L. perene seed germination (% inhibition)
L. perene 14-day growth (% inhibition)
R. sativum seed germination (% inhibition)
R. sativum 14-day growth (% inhibition)
E. albidus 14-day toxicity (% mortality)
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
1
2
3
4
Time (days)
0
3
7
14
28
45
—
—
—
—
375
0
375
375
36717
31714
19712
5375
25721
20712
375
875
1374
1377
5712
373
NT
NT
NT
NT
NT
3.3
NT
—
0
375
0
—
3778
26716
673
—
375
875
375
—
1278
2078
13722
—
NT
NT
NT
—
NT
5.5
NT
—
375
8710
13715
—
33710
27713
1377
—
375
8710
375
—
1975
2372
1674
—
NT
NT
NT
—
NT
3.1
NT
—
576
375
5710
—
33714
217
1579
—
8713
575
575
—
1273
1773
773
—
NT
NT
NT
—
0.9
2.6
NT
—
375
875
1375
—
45711
4679
3178
—
576
576
5710
—
1376
1875
873
—
NT
NT
NT
—
—
10.0
NT
416
48713
25719
8710
375
44715
41711
1678
2178
65719
33721
8710
10714
879
476
277
1712
—
—
—
—
Note.—, not measured; NT, nontoxic.
a
Toxicity of eluate.
time was observed. The toxicity of the eluates increased
slightly in sediment 1 from 0 to 0.9 toxic unit (TU), more
significantly in sediment 2 from 3.3 to 10 TU, and in
sediment 4 form 0 to more than 16 TU. The eluates of
sediment 3 were not toxic.
The change in toxicity was calculated as the difference
between TU at Days 0 and 45. The rate of increase in
toxicity differed between the sediments. Fig. 4 shows the
relationship between the rate of toxicity enhancement
and the rate of water flow through the sediments.
4. Discussion
The major factors controlling the release of metals are
pH, organic matter content, major element chemistry,
and biological activity. The rate of leaching of Cd and
Zn from sediments 1 and 2 was higher compared with
that from sediments 3 and 4. This higher mobility was
related to the low pH of sediments 1 and 2. Stronger
competition of metal ions with H+ ions for sorption
sites could be the reason for the more intensive release of
Cd and Zn from sediments of lower pH (Janssen et al.,
1997a). According to Kiekens and Cottenie (1985), Cd
and Zn tend to pass into solution at pH values lower
than 4. The pH of both sediments with high leaching
rates was lower than 4.5, while the pH of sediment 3,
Fig. 4. Relationship between water flow rate and the sediment and
eluate toxicity. Toxicity increases were calculated as a difference
between toxicity at Day 0 and toxicity at Day 45 (in TU).
which exhibited the lowest rate of leaching, was 6.
Warwick et al. (1998) investigated Zn and Cd mobility
in porous media and obtained similar results: in all cases
Zn and Cd were found to be more mobile at pH 4 than
at pH 6.5. A break in the relationship between pH and
the rate of Cd and Zn leaching was noted at pH 4.5 in
this study. This suggests that the rate of leaching
from such sediments probably strongly increases at
pHo4.5.
72
Z. Prokop et al. / Ecotoxicology and Environmental Safety 54 (2003) 65–73
As land-deposited dredged sediments are subjected to
transformation to terrestrial conditions, changes in the
chemical forms of metals may affect their mobility and
bioavilability (Tack et al., 1999). WHAM speciation
calculations predicted that Cd and Zn would remain
mainly bound on the clay fraction. However, in our
leaching tests, no relationship was observed between
clay content and metal behavior (distribution and
leachability). On the other hand, a correspondence
between TOC and leachability was observed: a higher
Cd and Zn leaching rate was noted in sediments of lower
organic carbon content. It should, however, be mentioned that the sediments of low TOC content also had a
low pH. The absence of clay as an important adsorption
phase, which can be attributed to the relatively weak
metal binding on clay surfaces compared with binding
to organic matter, was previously described by Janssen
et al. (1997b). In each sediment an increase in Cd and Zn
concentrations in the leachates as a function of time was
observed which may be explained by increased TOC
leaching. As metals are usually bound to the small-size
fraction of sediment particles or dissolved organic
matter in pore water, faster leaching through facilitated
transport of these particles without desorption may be
expected.
Based on the metal partitioning, the potential leachable amounts of Cd and Zn were calculated. The results
of these theoretical calculations are, in general, in
accordance with the observed experimental leaching
results. At Days 14 and 28, the observed Cd and Zn
concentrations in the eluates increased more rapidly
compared with the calculated values, which may due to
a possible underestimation of water flow through the
sediments. Indeed, as the flow through the sediments,
involved in the leaching calculation, was determined
form the amounts of eluate and eluate volumes were
affected by evaporation, the exact volumes of water that
passed through the sediments could not be exactly
assessed.
Reported ranges of NOEC values for zinc are between
265 and 600 mg/kg dry wt for soil invertebrates and
between 100 and 500 mg/kg dry wt for plants. NOEC
data for cadmium range between 5 and 320 mg/kg dry
wt for soil invertebrates and between 1.8 and 160 mg/kg
dry wt for plants (Van Gestel and Van Diepen, 1997).
The large variability in NOEC values is partly the result
of differences in the sensitivity of the species used and/or
the endpoints used, but also reflects the variability of
bioavailability-modifying factors in soils.
Since all investigated sediments were low in organic
carbon content and pH values ranged between 4.4 and 6,
the bioavailability of metal was expected to be high.
During transformation experiments, changes in toxicity
were observed. This was most obvious for sediments 1
and 2, for which an increase in phytotoxicity was
observed, despite the lower metal contamination. Eluate
toxicity, as measured with the invertebrate T. platyurus,
also exhibited a pronounced increase as a function of
transformation time. The ecotoxicity results corresponded well with the chemical observations described
above. Metal concentrations in the eluate were significantly higher at the end of the transformation
experiments than at the beginning, which is reflected in
the increased toxicity. The results indicated a relationship between water flow and changes in the toxicity of
the sediments and eluates. The greater water flow
through the sediments was connected with the lower
increase in sediment toxicity and greater increase in
eluate toxicity. The seeping water probably rinsed a
toxic fraction of metals out of the sediments into the
eluates. The rate of water flow through the layer of
sediment plays an important role in contaminant
leaching and threatening of groundwater.
5. Conclusions
In general it can be concluded that the retardation
and leaching potential calculations predicted well the
mobility of Cd and Zn in land-deposited sediments. The
mobility and leaching of Cd and Zn in the sediments
depended mainly on sediment pH and organic matter
content. During the transformation experiments an
increase in sediment toxicity to plants and increase in
eluate toxicity to invertebrates were observed. These
toxicity results corresponded with the chemical observations. Additionally, a correspondence between the rate
of toxicity enhancement and the rate of water flow
through the sediment was noted. A high rate of water
flow through the sediment resulted in a lower toxicity
enhancement of the sediments and higher toxicity
enhancement of the eluates. This suggests that water
flow through the sediment can reduce the actual toxicity of the upper layer of deposited sediment but at
the same time can intensify the risk of groundwater
contamination.
Acknowledgments
The authors thank Dr. Willie J.G.M. Peijnenburg and
Dr. Ji$rı́ Damborsk!y for editing the manuscript.
References
Allen, H.E., Fu, G., Deng, B., 1993. Analysis of acid-volatile sulfide
(AVS) and simultaneously extracted metals (SEM) for the
estimation of potential toxicity in aquatic sediments. Environ.
Toxicol. Chem. 12, 1441–1453.
American Society for Testing and Materials (ASTM), 1973. Standard
method of Test for metals in water and waste water by atomic
absorption spectrophotometry. ASTM Method D 2576-70. Annual
Z. Prokop et al. / Ecotoxicology and Environmental Safety 54 (2003) 65–73
Book of ASTM Standards, Water; Atmospheric Analysis. ASTM,
Philadelphia.
Benninger-Truax, M., Taylor, D.H., 1993. Municipal sludge metal
contamination of old-field ecosystems: do liming and tilling affect
remediaiton? Environ. Toxicol. Chem. 12, 1931–1943.
Carter, M.R., Ball, B.C., 1993. Soil porosity. In: Carter, M.R. (Ed.),
Soil Sampling and Methods of Analysis. CRC Press, Boca Raton,
FL, USA.
Centeno, M.D., Persoone, G., Goyvaerts, M., 1995. Cyst-based
toxicity tests: IX. The potential of Thamnocephalus platyurus as
test species in comparison with Streptocephalus proboscideus
(Crustacea, Branchiophoda, Anostraca). Environ. Toxicol. Water
Qual. 10, 25–34.
Culley, J.L.B., 1993. Density and compressibility. In: Carter, M.R.
(Ed.), Soil Sampling and Methods of Analysis. CRC Press, Boca
Raton, FL, USA.
De Cooman, W., Florus, M., Vangheluwe, M., Janssen, C., Heylen, S.,
De Pauw, N., Rillaerts, E., Meire, P., Verheyen, R., 1999. Sediment
characterisation of rivers in Flanders. The triad approach.
Proceedings: CATS 4, Characterisation and Treatment of Sediments, Antwerp, 15–17 September, pp. 351–367.
Fu, G., Allen, H.E., Cao, Y., 1992. The importance of humic acids to
proton and cadmium binding in sediments. Environ. Toxicol.
Chem. 11, 1363–1372.
Green, A.S., Chandler, G.T., Blood, E.R., 1993. Aqueous-, porewater-, and sediment-phase cadmium: toxicity relationships for a
meiobenthic copepod. Environ. Toxicol. Chem. 12, 1497–1506.
Hemond, H.F., Fechner, E.J., 1994. Chemical Fate and Transport in
the Environment. Academic Press, London.
International Organisation for Standardisation (ISO), 1994. Soil
Quality—Determination of pH, ISO No. 10390. ISO, Geneva.
International Organisation for Standardisation (ISO), 1996. Soil
Quality—Effects of Pollutants on Earthworms (Eisenia fetida),
Part 2: Determination of Effect on Reproduction, ISO No. 11268-2.
ISO, Geneva.
Janssen, R.P.T., Peijnenburg, W.J.G.M., Posthuma, L., Van Den
Hoop, M.A.G.T., 1997a. Equilibrium partitioning of heavy metals
in Dutch field soils. I. Relationship between metal partition
coefficients and soil characteristics. Environ. Toxicol. Chem. 16,
2470–2478.
73
Janssen, R.P.T., Posthuma, L., Baerselman, R., Hollander, H.A.D.,
Van Veen, R.P.M., Peijnenburg, W.J.G.M., 1997b. Equilibrium
partitioning of heavy metals in Dutch field soils. II. Prediction of
metal accumulation in earthworms. Environ. Toxicol. Chem. 16,
2479–2488.
Kiekens, L., Cottenie, A., 1985. Principles of investigations on the
mobility and plant uptake of heavy metals. In: Leschber, R., Davis,
R.D., L’Hermite, P. (Eds.), Chemical Methods for Assessing Bioavailable Metals in Sludges and Soils. Elsevier Applied Science,
New York, USA, pp. 32–41.
Lee, S.Z., Allen, H.E., Huang, C.P., Sparks, D.L., Sanders, P.F.,
Peijnenbugh, W.J.G.M., 1996. Predicting soil–water partitioning coefficients for cadmium. Environ. Sci. Technol. 30,
3418–3424.
.
Rombke,
J., Moser, T., Pessina, G., 1998. The Enchytraeid Reproduction Test. Protocol of the Final Assessment Workshop of the
International Ringtest, Ispra, Italy.
Roy, R.L., Campbell, P.G.C., 1997. Decreased toxicity of Al to
juvenile Atlantic salmon (Salmo salar) in acidic soft water
containing natural organic matter: a test of the free-ion model.
Environ. Toxicol. Chem. 16, 1962–1969.
Sauv!e, S., Dumestre, A., McBride, M., Hendershot, W., 1998.
Derivation of soil quality criteria using predicted chemical
speciation of Pb2+ and Cu2+. Environ. Toxicol. Chem. 17,
1481–1489.
Tack, F.M.G., Singh, S.P., Verloo, M.G., 1999. Leaching behaviour of
Cd, Cu, Pb and Zn in surface soils derived from dredged sediments.
Environ. Pollut. 106, 107–114.
Tipping, E., 1994. Wham-A chemical equilibrium model and computer
code for waters, sediments, and soils incorporating a discrete site/
electrostatic model of ion-binding by humic substances. Comp.
Geosci. 6, 973–1023.
Van Gestel, C.A.M., van Diepen, M.F., 1997. The influence of soil
moisture content on the bioavailability and toxicity of cadmium for
Folsomia candida willem (Collembola: Isotomidae). Ectoxicol.
Environ. Saf. 36, 123–132.
Warwick, P., Hall, A., Pashley, V., Van der Lee, J., Maes, A., 1998.
Zinc and cadmium mobility in sand: effects of pH, speciation,
cation exchange capacity (CEC), humic acid and metal ions.
Chemosphere 36, 2283–2290.
3B2v7:51c
GML4:3:1
Prod:Type:
pp:121ðcol:fig::NILÞ
YEESA : 2409
ED:Gayathri
PAGN: Mamatha SCAN: Nil
ARTICLE IN PRESS
1
3
Ecotoxicology and Environmental Safety ] (]]]]) ]]]–]]]
5
Erratum
7
Erratum to ‘‘Mobility and toxicity of metals in sandy sediments
deposited on land’’ [Ecotoxicol. Environ. Saf. 54 (2003) 65–73] $
9
11
Z. Prokop,a, M.L. Vangheluwe,b P.A. Van Sprang,b C.R. Janssen,c and I. Holoubeka
a
13
Department of Environmental Chemistry and Exotoxicology, Masaryk University, Kotlarska 2, 602 00 Brno, Czech Republic
b
EURAS Grote Baan, 199,9000 Ghent, Belgium
c
Laboratory of Environmental Toxicology and Aquatic Ecology, Ghent University, J. Plateaustraat 22, 9000 Ghent, Belgium
15
Received 17 February 2003; received in revised form 17 March 2003; accepted 14 April 2003
57
O
21
On p. 67, Table 1 contains incorrect values in the SEM concentration column. These mistakes do not alter the
discussion nor the conclusions (the text of article) because the SEM/AVS ratios are still greater than 1. For the reader’s
convenience, the corrected Table 1 appears here.
O
19
F
17
Table 1
Physical and chemical characterization of the sediments
27
Sample
pH
Organic matter (%)
1
2
3
4
4.45
4.43
6.05
5.15
0.9
0.4
1.3
0.3
Porosity (%)
Bulk density (g/cm3)
0.40
0.38
0.45
0.41
1.49
1.54
1.33
1.53
33
37
Water flow (mL/day)
3.21
7.68
13.04
17.78
61
63
65
SEM (mmol)
AVS (mmol)
SEM/AVS
67
1.21
2.96
7.08
2.29
0.368
0.059
0.412
0.315
3.3
50.1
17.2
7.3
69
WHC (%)
SO4 (mg/L)
Cl (mg/L)
73
32
32
39
32
101
173
108
84
53
46
49
38
R
35
1
2
3
4
TE
31
4.00
1.80
1.70
2.20
EC
29
Clay (o2 mm) (%)
D
25
PR
23
59
71
75
77
Note: SEM, simultaneous extracted metal; AVS, acid volatile sulfide; WHC, water-holding capacity.
R
39
O
41
C
43
N
45
U
47
49
81
83
85
87
89
91
51
53
79
$
doi of original article 10.1016/S0147-6513(02)00022-2.
Corresponding author. Department of Environmental chemistry and exotoxicology, Masaryk University, Veslarska 230B, Brno 637 00, Czech
Republic. Fax: +54-112-9506.
E-mail address: [email protected] (Z. Prokop).
93
55
0147-6513/03/$ - see front matter r 2003 . Published by Elsevier Science (USA). All rights reserved.
doi:10.1016/S0147-6513(03)00078-2
95