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a Swiss Federal Institute of Technology (EPFL), ENAC-ISTE-LPE, Ecublens, 1015 Lausanne, Switzerland
b Agriculture and Environment Division, IACR-Rothamsted, Harpenden, Herts AL5 2JQ, UK
* Corresponding author (catherine.keller{at}epfl.ch)
Received for publication July 5, 2001.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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Abbreviations: DOC, dissolved organic carbon
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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Considerable uncertainty exists about the long-term fate of polluting trace metals contained in sewage sludge. One possibility is the stabilization and even a reduction of the trace metal availability to plants by progressive trace metal immobilization into less soluble forms such as occlusion in Fe and Al oxides or precipitation as silicates, phosphates, or carbonates (Dowdy et al., 1994; Smith, 1997; Brown et al., 1998). Another possibility might be an increase of trace metal bioavailability and leaching through sewage sludge organic matter mineralization ("time bomb effect") (Zhao et al., 1997). Field studies covering several decades have produced unclear results (Chang et al., 1997; Logan et al., 1997) and led to contradictory conclusions (Chaney and Ryan, 1993; McBride, 1995; McGrath et al., 2000). Bioavailability to plants and phytotoxicity of trace metals obtained from long-term experiments have recently been summarized by Berti and Jacobs (1996), Barbarick et al. (1997), Miner et al. (1997), Sloan et al. (1997), Zhao et al. (1997) and Keller et al. (2000). Conclusions varied with plant species, soil characteristics, and trace metal loadings.
Little is known about trace metal leaching from sewage sludge–treated soil to ground water. Although dilution will usually prevent toxic effects from occurring, locally and/or temporally trace metal concentrations might reach unacceptable levels for drinking water (Richards et al., 1998) in sensitive areas with shallow and small water bodies and large areas treated with sewage sludge. Mass balances calculated for long-term experiments suggest either losses of trace metals from the topsoil not attributable to plant uptake (McBride et al., 1999) or total retention within the sludge incorporation layer after sludge application (Sommers et al., 1979; Williams et al., 1984; Williams et al., 1985; Sloan et al., 1998). In the first case, lateral movements of soil due to cultivation (McGrath and Lane, 1989) or physical mixing with the underlying layer by plowing (Sloan et al., 1998) can be responsible for a significant part of the losses from the original amended soil volume. However, mass balances calculated for sites where little or no tillage was performed have shown less than 100% trace metal recovery (McBride et al., 1999).
Because the transfer through the soil profile was often observed indirectly by measuring total and extractable metal concentrations at different depths, methodological problems due to trace metal extractability from soils have been proposed as an explanation for a recovery of less than 100% (Chang et al., 1982; Sloan et al., 1998). However, Darmody et al. (1983) have shown that Zn and Cu concentrations increased at depth after composted sewage sludge application on soils. After 10 yr of sludge application, Barbarick et al. (1998) measured higher Zn concentrations in NH4HCO3–DTPA at a 100- to 150-cm depth for different soils. Baveye et al. (1999) found higher DTPA-extractable Cu, Ni, and Pb at a 75-cm depth under a silt loam soil treated with digested sewage sludge than under the control soil, although total concentrations were similar throughout the soil profile, and Keller et al. (2000) found also increased NaNO3–extractable Zn concentrations 20 cm below the top layer of a loamy soil mixed with a highly contaminated sewage sludge. From these results the possibility of metal migration to ground water as an explanation for the losses calculated from mass balances cannot be ruled out, especially in soils with coarse textures or with preferential pathways. Indeed, this approach gives reliable results only if the metals are still present and have not been leached out of the profile. Also, owing to the general soil heterogeneity, the difference in soil concentrations between treated and untreated plots has to be large enough to be significant, so only important metal leaching can be detected.
However, few direct measurements in soil solutions or drainage waters have been performed so far: for example, 16 yr after a single sludge application of 244 Mg ha-1 dry wt., McBride et al. (1999) measured a whole range of elements in soil waters collected at a depth of 60 cm under control and treated soils. They found that Cu, Zn, Cd, and Ni had higher concentrations under the treated field than under a nearby control field. But the concentrations measured at that time could not explain the relative loss of metal in the first 25 cm of the soil (compared with the immobile metal Cr) ranging from 20 (Zn) to 70% (Sr). This was probably due to the fact that major losses occurred immediately after sludge application when no soil solution was sampled, as suggested by Richards et al. (1998). Indeed, the period most at risk for metal leaching is probably just after sludge application when soluble organic matter is available and/or when preferential flows may occur (Darmody et al., 1983; Dowdy et al., 1991; Camobreco et al., 1996; McBride et al., 1999; Maeda and Bergstrom, 2000).
Transfer through the soil profile can occur in solution or as suspended matter. In unpolluted soils, suspended clay-sized particles are well known to migrate from the subsurface horizon to deeper horizons (Chadwick and Graham, 2000). In polluted soils, however, few attempts have been made to identify and quantify particle movement (Thompson and Scharf, 1994), although particles may participate in the transfer of organic and inorganic pollutants through the soil (Vinten et al., 1983; Dunnivant et al., 1992; Gasser et al., 1994; Keller and Mavrocordatos, 1997).
The purpose of this study was thus to follow the effect of average and heavy rates of sewage sludge inputs on a loamy soil to assess food chain and ground water contamination during the first 37 mo, assuming that most of the losses would occur shortly after sludge application. Trace metal phytotoxicity and plant uptake were monitored and trace as well as some major element losses were quantified in drainage water of a brown soil–meadow system just after repeated sewage sludge applications with undisturbed monolith lysimeters that allow mass balance calculations and prevent lateral losses.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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Sewage Sludge
We used fairly contaminated digested urban–industrial sewage sludges (Davyhulme, Manchester, and Minniworth Works, Birmingham, UK) with different trace metal concentrations (Table 2). The first sludge received only a primary sedimentation treatment. This was a liquid sludge (3% dry matter) with a pH of 7.2, 7.2% N, 2.4% P, and 57% w/w organic matter in dry weight (Wang and Jones, 1994). Its composition is well documented and was fairly stable over time regarding trace metals such as Zn, Ni, Cd, Pb, Cu, and Cr, and also organic compounds (Alcock and Jones, 1993; Wang and Jones, 1994; Wilson et al., 1994). Trace metal concentrations were above the average concentrations measured in sludge applied to land in the UK (Gendebien et al., 1999). However, trace metal concentrations measured in the first sludge were below the maximum permitted by the Commission of the European Communities (1986) and the USEPA regulations (USEPA, 1993). The second sludge had greater trace metal concentrations and was chosen to simulate a worse case. Cadmium, Cu, and Pb concentrations were above the limits set by the European Communities for sludge use in agriculture. If compared with the USEPA limits, only Ni and Pb were clearly too high whereas Zn was at the limit. Because of the Zn concentration, this sludge would not be allowed for use in agriculture in many European Union states, but the total metal loading complied with the U.S. regulations on maximum annual and cumulative loadings for sludge-treated soils (see below).
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The sewage sludge was applied manually on the top of the lysimeters at time zero, and after 11 and 17 mo. The surface was not mixed with the topsoil because the vegetation was already growing on it, but it was slightly scraped to prevent the formation of a hydrophobic layer. Also because of this, the highest treatment was applied in two doses over 2 d. No fertilizer was applied, but the three applications of sludge over 17 mo resulted in total inputs of 648 and 293 kg ha-1 of N and P for the 3 x 100 m3 ha-1 application and of 2592 and 1174 kg ha-1 of N and P for the 3 x 400 m3 ha-1 load. These values are above the 250 kg N ha-1 yr-1 recommended in the UK code (Ministry of Agriculture, Fisheries and Food, 1993). Regarding trace metals (Table 2), and for the first sludge application, both rates were below the annual metal loading limits set for sludge-treated soils by the UK Department of the Environment and the USEPA (1993). For the second sludge application, both loads were also below the USEPA limits but above the European Union limits for Zn in the 100 m3 ha-1 treatment and for Cu, Zn, and Ni in the 400 m3 ha-1 treatment.
Due to the exceptionally dry weather during the first spring, the lysimeters were watered between Months 1 and 4, which resulted in a 220-mm addition of equivalent rain. The water content inside the lysimeters was not monitored. Drainage occurred only once during this period. The average rainfall at Woburn is 641 mm (1961–1990). The rainfall was 645 mm (150 mm in September) plus 220 mm added during spring in 1995, 469 mm in 1996, 503 mm in 1997, and 236 mm in the first four months of 1998. Although the annual rainfall in 1995 was as high as the following years, 1995 was significantly warmer and monthly rainfall between April and September were well below the average. The original meadow herbage was harvested in Months 2 and 27. Water samples were collected after every rain event between Month -3 (Day -80; i.e., 3 mo before sludge application) and Month 37 (Day 1125). Mass balances were calculated between Month 0 (Day 0) and Month 37 (Day 1125).
Soil, Sludge, and Plant Analysis
Soil and dry sludge samples were analyzed for total element concentrations after digestion in aqua regia (McGrath and Cunliffe, 1985). The herbage was weighed immediately after harvest, dried at 50°C, and weighed again to obtain percent dry matter. For each lysimeter, herbage was sorted by hand to remove shoots spoiled with sewage sludge (visual evaluation based on the fact that the sludge had left black residues on shoots). The samples were then ground for N analysis (LECO [St. Joseph, MI] CNS 2000). Plants were digested in a mixture of 15% perchloric acid and 85% nitric acid (Zhao et al., 1994). Major and trace elements were measured by inductively coupled plasma–atomic emission spectroscopy (ICP–AES) (ARL; Fisons, Accuris, Loughborough, UK). Accuracy of analysis was checked against an internationally certified reference material (ryegrass [Lolium perenne L.], CRM 281).
Water Sampling and Analysis
All samples were collected in metal-free containers and all apparatus that came into contact with samples and filters was washed in 5% HNO3 (v/v) and rinsed with purified water prior to use. Purified water (RO Still; Whatman, Maidstone, UK) was used throughout. After the drainage volume was measured, 2 L were taken to the laboratory in Teflon-coated polyethylene bottles. Total organic carbon (TOC) and pH were measured on unfiltered solutions. One liter was filtered through a 0.22-µm Whatman nitrate cellulose membrane on a Teflon-coated support, and aliquots were taken for dissolved organic carbon (DOC), anions (Cl, NO3), and major and trace elements analysis. The membrane type was chosen after a series of tests performed on different membranes to check for trace metal adsorption and release. For samples collected between Months 0 and 6, the membranes were digested in a mixture of 70% HNO3 and 35% HCl to quantify trace metals present in the particulate fraction (>0.22 µm). After Month 6 the solutions were filtered on a 0.45-µm membrane because no difference in composition was detected between 0.22 and 0.45 µm–filtered solutions, and no further particulate matter determination was performed.
Trace and major elements in solutions were analyzed by inductively coupled plasma–atomic emission spectroscopy (ARL; Fisons), except Cd, which was analyzed by graphite furnace atomic absorption spectrometry (GFAAS) (PerkinElmer [Norwalk, CT] 4100ZL). Total organic carbon and DOC were analyzed by UV–persulfate oxidation (Dohrman [Cincinnati, OH] DC-80). Chloride and NO3–N were measured by continuous colorimetric flow analysis (Skalar SAN PLUS system; Skalar Analytical BV, Breda, the Netherlands).
X-ray diffraction was performed on the particulate fraction of two samples from L9 (in a sample from Month -1) and L7 (Month 2). Samples were prepared as follows: 30 mL of 1 M MgCl2 was added to 3 L of soil solution and left to precipitate overnight. The sample was then centrifuged at 1300 x g for 30 min. The clear supernatant was discarded. The samples were shaken twice with 40 mL H2O and centrifuged again for 30 min at 1300 x g. The supernatant was discarded and the remaining material was dried with ether in a crucible and weighed. All X-ray diffraction work was performed with Fe-filtered CoK
radiation (35kV, 40 mA) and a Philips PW1050 vertical goniometer (FEI Company, Eindhoven, the Netherlands) controlled by a PW1710 controller, connected to a microcomputer. A xenon-filled proportional counter was used, with pulse-height discrimination in the counting chain.
Statistical Analysis
Comparison between sludge treatments was performed with unpaired Student t tests on mass balance data and Pearson correlations on herbage data. Unpaired Student t tests were performed on restricted periods after sludge applications (after 500 and 900 d) to assess the short-term effect of sludge application on trace and major element concentrations in drainage waters.
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RESULTS AND DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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Calcium, NO3–N, and Cl concentrations showed the same pattern (Fig. 1 and 2 ) with significant correlations of rCa-NO3 = 0.982*** and rCa-Cl = 0.889*** for the highest sludge treatment. They presented a slight increase after the first application and a larger increase after the second and third applications. The highest load resulted in a larger leaching. In the case of the first sludge application there was a delay of approximately 4 mo after the sludge application and before an effect could be observed. This was due to the exceptionally dry season with no drainage, which could not be compensated by watering. Calcium, Cl, and NO3–N concentrations measured after the first sludge application were less than half of those measured after the second and third treatments. The Ca concentrations were higher in the second sludge and could explain the difference. However, nitrogen content was the same and the amount leached after the first sludge application should have been half of the amount leached the second time. However, due to a lack of drainage, nitrogen was probably either adsorbed on the soil matrix or taken up by the plants from the capillary soil solution. In this case, the possible amount to be leached with the following rain would be reduced. Nitrogen concentrations measured in herbage harvested after the first application were indeed higher than those measured in herbage harvested after the later applications. Calcium, NO3–N, and Cl concentrations in the highest treatment returned to control levels within 3 mo after a peak was observed for the first sludge application and within 6 mo (after Day 900) after the common peak of the second and third applications. Richards et al. (2000) found, in smaller columns (35-cm length) of a coarse-textured soil amended with dewatered digested sludge and watering optimization, that Na and K concentrations returned to background levels after approximately 3 and 6 mo, respectively. Despite the different experimental conditions, one can consider their results as fairly similar to our findings.
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Table 4 gives the results of the t tests performed to compare means of the control lysimeters with the means of the treated lysimeters for five rain events collected immediately after sludge application (Day 590 to Day 714) and for the last seven rain events (Day 955 to Day 1125). From these results it can be concluded that the risk of element leaching started shortly after sludge application and that rainfall partly determined the speed and duration of the leaching. Thus, concern about sludge applied to soil should focus on the first months and up to 2 yr after application, at least for soils with similar characteristics (loamy sand with possible preferential flow pathways). In our case, measurements performed several months after treatment would have missed most of the leaching. In studies where losses in mass balances were calculated on the basis of average concentrations measured under the topsoil several years after sludge application (McBride et al., 1999), a similar phenomenon may have occurred.
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Another explanation might be organic matter solubilization, especially after 955 d. As already mentioned, the four replicates receiving sewage sludge were divided into two replicates with high DOC and two with low DOC. This situation remained throughout the study and resulted in large standard deviations of the mean DOC values (Fig. 1). In contrast, the two control lysimeters had low initial DOC concentrations. The DOC mean value of the control lysimeters was thus always lower than the treated ones. As a consequence we could only expect an increase in the difference between the controls and the treated lysimeters after the sludge was applied. However, there was no difference between the two sludge treatments during the whole experiment except during the last winter when DOC concentrations increased in the 400 m3 ha-1 sludge treatment slightly after 590 d and more after 955 d (Table 4), following the same pattern as Ni and Cu. The sludge organic matter might have started to decompose, and the pH range of the solutions was favorable to organic matter solubilization. The increase observed in Cu and Ni concentrations (especially after Day 900) could thus result from mobilization by the soluble organic matter. Indeed, McBride et al. (1999) observed that the mobile forms of Cu, Zn, and Cd in percolates with a pH of 7 to 7.5 collected under sludge-treated soils were largely complexed forms, probably with dissolved organic matter. Garnett et al. (1987) also found that Cu and Ni formed soluble complexes with organic ligands. Dunnivant et al. (1992) observed increased Cd mobility through a column of sand in the presence of DOC, but we could not find a similar trend in our study. Although this point needs to be verified over a longer period, the possibility of trace metal mobilization by the soluble organic matter is supported by the fact that average Cu concentrations appeared to be higher under lysimeters with higher DOC concentrations (t test significant with a probability of 97.5% for Cu concentrations after Day 900).
Particulate Transfer through the Soil
During the first part of the experiment (before and after the first sludge application), particulate transfer was quantified for some elements. The results varied between lysimeters but the treatments had no effect on the amount of particulate matter collected. However, some lysimeters gave systematically higher amounts of suspended matter (L7, 8, 9, 10). These lysimeters were also those that had systematically higher DOC concentrations. Calculated over six sampling dates between Months 2 and 6, the average particulate matter concentration was 4.6 mg L-1 for the four lysimeters mentioned above and 0.6 mg L-1 for the others. However, total organic C and DOC concentrations were identical (i.e., organic carbon was not preferentially associated with the particulate matter). Calculated over a 7-mo period (Fig. 3) , this particulate matter accounted for a maximum 20% of the total trace metal concentration (soluble + particulate). Higher percentages calculated for Cr and Co were due to very low concentrations and high measurement uncertainty and thus not reliable. Even though these results are subject to high variability, the quantity and the nature of the particulate matter may favor trace element transfer down the soil profile. Indeed, X-ray diffraction performed on the particulate fraction of two samples showed the presence of goethite and of some clay minerals with a swelling component (including mica-smectite and interstratified clays) as well as some kaolinite, similar to those found in the clay fraction of all the soil horizons (Catt et al., 1975). Clay minerals, especially smectites, have large cation exchange capacity whereas hydrous Fe can coprecipitate or adsorb a broad range of trace metals (Alloway, 1997). For example, Gasser et al. (1994) found that in serpentine soils, lysimeter waters generally contained more than 50% of colloidal (calculated as being of a size between 10 000 Da and 0.45 µm) Cr and Ni (average of three rain events).
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Total Exports of Trace Metals and Selected Major Elements in Solution
Table 5 shows mass balance calculations for the study period for trace elements as well as for NO3–N and Ca. Chloride balance was not calculated. The input includes the three sewage sludge applications, the herbage offtake includes two harvests, and the soluble output represents the sum of all filtered drainage waters. The initial soil stocks were calculated from the total element analysis done every 20 cm down to a 100-cm depth, with an average bulk density of 1.5 (Catt et al., 1975). All trace metals were originally present in small quantities in the soil and thus account for a very small part of their total amount present in the lysimeters at the beginning of the experiment. Unfortunately, no direct metal measurement within the soil profile after sludge application could be performed because the lysimeters remained in use after the experiment.
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In total, the sludge treatments resulted mostly in an increase in soluble output. However, the effect on herbage offtake was also significant on several metals but after the first sludge application only. From the percentages calculated, it is clear that the major mechanism operating within the soil column was retention. This resulted in an increase in metal concentrations in soil because most of the sludge-applied trace metals remained in the soil column.
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CONCLUSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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In the case presented here, the main long-term risk is thus the rapid increase in trace metal concentrations in the topsoil as well as the increase up to toxic levels in herbage. In the worst case (1200 m3 ha-1 in 3 yr of the same sludge as in the experiment), and assuming negligible leaching, the Zn limit concentrations (200 mg kg-1) according to the UK code (Ministry of Agriculture, Fisheries and Food, 1993) would be reached within a time span of 10 yr if metals did not migrate much below the plow layer (20 cm).
Additional results suggest caution when trying to quantify outputs and risks for the food chain and ground water. The increased leaching of DOC and particulate matter found in some of the lysimeters might be due to preferential pathways within the soil columns. This situation will probably increase the risk of subsoil and ground water contamination as also seen by Camobreco et al. (1996). Indeed, we found that over a 7-mo period, particulate matter could amount to up to 20% of the total trace metal leaching. The phenomenon has already been mentioned as an increasing risk factor for trace metals and nitrogen leaching (Maeda and Bergstrom, 2000).
Finally, after 30 mo we detected an increase in DOC under the treated lysimeters. Before that, no effect of the sludge on the DOC concentrations had been observed. Solubilization of the sludge organic matter seemed to occur, resulting in a possible increased Cu and Ni mobilization down the profile and to the ground water. This suggests that the sludge was subject to degradation and the system had not reached a steady state yet. This effect was delayed by at least one year and it has to be concluded that, despite the observed rapid response of the lysimeters to the sludge application, further effects from sludge on leachate quality may occur.
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ACKNOWLEDGMENTS |
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REFERENCES |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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