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Mobility and toxicity of metals in sandy sediments deposited on land
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
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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
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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
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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.
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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
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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
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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.
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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$rı´ 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.
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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.
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
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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
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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
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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
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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
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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.
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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$rı´ 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.