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The organic carbon derived from sewage sludge as a key parameter determining the fate of trace metals
Chemosphere 69 (2007) 636–643 www.elsevier.com/locate/chemosphere
The organic carbon derived from sewage sludge as a key parameter determining the fate of trace metals C. Parat a
a,*
, L. Denaix b, J. Le´veˆque a, R. Chaussod c, F. Andreux
a
Microbiologie et Ge´ochimie des Sols, UMR 1229, INRA-Universite´ de Bourgogne, Centre des Sciences de la Terre 6 Bd Gabriel 21000 Dijon, France b INRA Bordeaux—Unite´ d’agronomie, BP 81 33883, Villenave d’Ornon Cedex, France c Microbiologie et Ge´ochimie des Sols, UMR 1229, INRA-Universite´ de Bourgogne 17, rue de Sully 21034 Dijon Cedex, France Received 9 July 2006; received in revised form 23 February 2007; accepted 26 February 2007 Available online 18 April 2007
Abstract In a sandy agricultural soil of south-west of France, continuously cultivated with maize and amended with sewage-sludge over 20 years, the behavior of three trace metals (Cu, Pb, and Zn) was studied during the sludge applications (1974–1993) and after its cessation (1993–1998). Using the d13C analysis, the dynamics of different sources of organic matter were followed in order to elucidate the influence of the sludge-derived organic matter on the fate of trace metals in the soil and its particle size fractions. This study revealed that sludge-derived organic matter contributed to the formation of macroaggregates through the binding of preexisting microaggregates. These macroagreggates were thus responsible for the accumulation of trace metals in the coarsest fraction as well as for the protection of maize-derived organic matter against biodegradation. After sludge application ceased, the disaggregation of macroaggregates occurred simultaneously with high losses in Cu and Pb. On the contrary, Zn appeared less affected by the cessation of sludge application, with only a location change from coarse to fine fractions. Ó 2007 Elsevier Ltd. All rights reserved. Keywords: Sewage-sludge; Trace metals; Organic carbon; d13C; Particle-size fractions
1. Introduction The agricultural use of sewage-sludge has been a common practice in waste disposal in recent decades. Sewagesludge contains plant nutrients and organic matter, and therefore may significantly modified soil organic matter, pore sizes and aggregate stability (Adani and Tambone, 2005; Novoszad et al., 2005). But their positive influence may be limited by their contents of potentially harmful substances such as organic micropollutants and trace met-
*
Corresponding author. Present address: Laboratoire de Chimie Analytique, LCABIE, UMR 5254 IPREM, UFR Sciences, Avenue de l’Universite´, 64000 Pau, France. Tel.: +33 3 559 40 76 70; fax: +33 3 559 40 76 74. E-mail address: [email protected] (C. Parat). 0045-6535/$ - see front matter Ó 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2007.02.069
als (Kidd et al., 2007). In last years, the concentrations of most metals in sludge in many countries have shown a pronounced decrease as a result of improved effluent control and waste minimisation (Alloway, 1995). However, in view of the high amount of sludge applied to land in the past and the very long residence time of trace metals in soil surface (McGrath and Lane, 1989), it is still important to be aware of trace metal amounts added with sludge. Current debates focus on two hypotheses: some advocate the sludge protection hypothesis, which states that trace metals are bound within sludge components in forms that are poorly available, and that the specific adsorption capacity of sludge will persist as long as trace metals of concern persist in soil (Shuman, 1999). Others consider that the slow mineralization of sludge-derived organic matter could release metals as more soluble forms (McBride, 1995), an effect often termed ‘‘the sludge time bomb hypothesis’’ (Chang et al., 1997). There is therefore a need
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to know the kinetic behavior of trace metals following sludge applications, as well as after the termination of sludge applications. In this study, we investigated the behavior of three trace metals (Cu, Pb and Zn) in the long-term field experiment laid out in south-west France by the National Institute of Agronomic Research (INRA, Bordeaux) and that began in 1974. Previous work has shown an improvement in microbial biomass in spite of a higher trace metal contamination (Parat et al., 2005). A kinetic study was therefore undertaken in order to elucidate the relationship between sludge-contained trace metals and sludge-derived organic matter. A strength of this study is that complementary approaches (physical and chemical fractionation) were used on stored soils samples collected prior to the start of the experiment in 1974, but also during sewage sludge applications and after their termination. 2. Materials and methods 2.1. Study site Plots of 18 m2 have been established in 1974 at the INRA experimental site of Couhins, near Bordeaux (France). The experimental site consists of four treatments, laid out with five replicates in a randomized block design, the only difference between plots being the type of amendment. A slope of 3% has induced a contamination between blocks, which led us to work on the block no. 5 located at the top of the site. The soil is a slightly acidic sandy soil developed on colluvial material (Fluvisol, FAO classification), with initial contents of organic carbon and nitrogen of 13 ± 2 and 0.90 ± 0.01 g kg1 of air-dry soil, respectively, a pH of 6.0 and a bulk density of 1.36 ± 0.02 g cm3. Before 1974, this soil was always cropped with C3 species (vineyard). Since 1974, it has been continuously cropped with maize (Zea mays L. cv. INRA 260). Anaerobically digested sewage sludge, strongly contaminated with Cu (215 mg kg1 of air-dry sludge), Pb (752 mg kg1 of airdry sludge) and particularly Zn (5052 mg kg1 of air-dry sludge), was incorporated from 1974 to 1993 at a rate of 100 t ha1 every 2 years (S100 soil). This rate represented an organic carbon input of 1.85 kg C m2 every year. The control, which has never received any sludge treatment, was adjusted annually, prior to planting, to the same level of mineral fertilisation (N, P and K) as sludge-treated soil (Juste and Soldaˆ, 1986). Every year, the maize residues were buried in the soil. Between 1974 and 1998, cumulative maize grain yield was 870 ± 80 t ha1 in the control plots and 930 ± 90 t ha1 in the S100 plots. Since 1993, no treatments have been applied. Six soil-cores in the plough-layer (0–25 cm) were taken from each plot before maize planting and mixed to form a representative sample of each plot. Representative soil samples were collected in March 1974, 1981, 1993 and 1998 from the two treatments, airdried and passed through a 2-mm aperture sieve, and stored in plastic sealed containers.
637
2.2. Particle-size fractionation Particle-size fractionation was carried out by the procedure of Bartoli et al. (1991); on the whole soil (<2 mm soil fraction) and on the sewage-sludge samples. Briefly, for each sample, 20 g of material were placed in a glass flask with 600 ml of distilled water, and shaken mechanically for 16 h. Three fractions were obtained by wet sieving: the coarse sand fraction (200–2000 lm), fine sand fraction (50–200 lm), and the silt and clay fraction (0–50 lm), each then dried at 40 °C for 24 h and ground prior to analysis. 2.3. Trace metal analysis Total metal contents (Cu, Pb and Zn) were determined after acid digestion according to the following procedure: soil samples were first crushed and dried at 105 °C during 12 h. Aliquots of 0.1 g were digested in a Teflon bomb with a solution of concentrated HNO3 (3 ml), HClO4 (2 ml) and HF (3 ml) during 16 h. Then, the mixture was evaporated to dryness, 5 ml of deionized water was added, and evaporated to dryness. Finally, the residue was dissolved in 1 ml of concentrated HNO3 and filtered. The solution was then transferred into a 100 ml flask. Contents of Cu, Zn, and Pb were determined by graphite furnace atomic absorption spectrometry (GFAAS, HGA 600). 2.4. Sequential extraction The sequential extraction procedure of Shuman (1985) was performed on all soil samples and one of sludge samples (Parat et al., 2003) and allowed to isolate the following fractions: MgNIT (exchangeable forms), OMOCl (oxidizable organic/sulfide fraction), NH2OH (manganese oxides), TAMAs (iron oxides) and RESID (residual). The technique involved mixing and shaking 1.0 g of airdried soil/sludge previously screened at 2 mm with 10 ml of reagent, followed by centrifugation and washing with 40 ml of deionized water. The residue was dried at 40 °C before the next step. The obtained extracts were transferred into a 100 ml flask. Copper, Pb and Zn were measured by inductively coupled plasma mass spectrometry (ICP–MS). 2.5. C, N, and carbon isotope analyses Carbon and nitrogen measurements were made on 30 mg of ground sample (of both, whole sample and particle size fractions) by the dry combustion method (Carlo Erba CNS 1500 analyser, Dijon, France). The standard used was sulphanilamide (41.84% of C and 16.27% of N). The natural abundance of stable carbon isotope 13C was measured by using a continuous flow mass spectrometer (VG Isochrom-EA, Dijon, France) on ground samples. The variations of soil isotopic signature depend on the organic inputs (Balesdent et al., 1987). Consequently, when dealing with a mixture of carbon originating from different
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vegetations, their respective proportions are calculated using Eqs. (1) and (2): X total ¼ X native þ X maize þ X sludge 13
ð1Þ
13
13
d Csample ¼ X native d Cnative þ X maize d Cmaize 13
þ X sludge d Csludge
where DC is the standard deviation of the concentration C in g kg1 for carbon or in mg kg1 for trace metals, and Dd is the standard deviation of the soil density d. 3. Results and discussion
ð2Þ 3.1. Trace metal accumulation in control and S100 soil
where Xmaize, Xnative and Xsludge are the proportions of maize C, native C and sludge C, at the sampling date, respectively, and d13Cnative (26.5&), d13Cmaize (12.5&), d13Csludge (25.4&) are the mean isotopic values measured in the respective sources. To solve Eqs. (1) and (2), the control soil was used to calculate the contribution of native organic carbon in sludge-treated soil (Parat et al., 2007). 2.6. Morphological study The morphology of the particle-size fractions of the control and S100 soil was studied using a binocular and a scanning electronic microscope (ASEM, JEOL JSM-6400F) interfaced with an energy dispersive X-ray spectrometer (SEM–EDS). 2.7. Amount and error calculations The measurements were all done in triplicate from composite soil samples. Mean and standard errors are reported in tables. Amount calculations were obtained by using the following equation: S whole soil ¼ C whole soil d t
ð3Þ
where Swhole soil is the amount in the whole soil in kg m2 for carbon or g m2 for trace metals, Cwhole soil is the total concentration in g kg1 for carbon or mg kg1 for trace metals, d is the density (here, 1.36 g cm3) and t is the soil thickness (here, 0.25 m). The errors DSwhole soil on the amount Swhole soil have been determined by using the following equation: DC whole soil Dd DS whole soil ¼ S whole soil þ ð4Þ d C whole soil
3.1.1. Variations of total trace metals Between 1974 and 1998, small variations of trace metal appeared in the control soil contrary to the sludge-treated soil which showed a significant accumulation in the surface soil layer till 1993 (Table 1), due to the high content of trace metals in sewage sludge. Since 1993, when sewage sludge applications were stopped, a significant decrease of Cu (47%) and Pb (56%) and a slighter decreased of Zn (16%) were observed, whereas many authors showed that Cu and Pb are less mobile than Zn (Alloway, 1995). It is therefore necessary to relate such a differential decrease to the variation of other soil and sludge components. 3.1.2. Trace metal distribution in particle-size fractions The particle-size fractionation of soil samples resulted in the recovery of 97–99% of the initial sample mass, with similar size distributions for the different plots. At the start of the experiment, the control soil showed a predominance of Cu and Pb in coarse sand and silt and clay fractions, whereas Zn was equally distributed in the three fractions (Table 2). There was no significant variation of trace metal distribution with time of cultivation in any fractions, except in the case of Cu which strongly decreased in the coarsest fraction since 1993. In 1998, the largest amounts of Cu and Pb were found in the silt and clay fraction, whereas Zn was mainly located in the coarse sand fraction. Surprisingly, the repeated application of sewage sludge induced a significant accumulation of trace metals in the coarse sand fraction (Fig. 1). In 1993, the amount of trace metals in the coarse sand fraction was 2–6 times higher than in the fine sand and silt and clay fractions. This is in contradiction with many studies, which showed that trace metals mainly accumulate in the finest fractions with high concentrations of clay minerals and colloidal organic matter (Essington and Mattigod, 1990; do Valle et al., 2005). Since
Table 1 Variations of trace metals, organic carbon, nitrogen, and d13C between 1974 and 1998, in control and S100 soils Cu (g m2)
1974 1981 1993 1998
Pb (g m2)
Zn (g m2)
Carbon (kg m2)
Nitrogen (kg m2)
d13C (&)
Control
S100
Control
S100
Control
S100
Control
S100
Control
S100
Control
S100
7.3 (0.9) 7.3 (0.5) 5.6 (0.4) 4.8 (0.3)
7.3 (0.9) 22 (1) 40 (3) 21 (1)
10.6 (0.7) 10.7 (0.4) 12.1 (0.4) 7.5 (0.3)
10.6 (0.7) 66 (2) 101 (4) 44 (2)
5.1 (0.4) 3.7 (0.1) 9.2 (0.3) 10.3 (0.4)
5.1 (0.4) 176 (6) 351 (12) 294 (10)
4.4 (0.2) 4.39 (0.08) 4.34 (0.08) 3.74 (0.07)
4.4 (0.2) 8.1 (0.2) 10.4 (0.2) 7.5 (0.1)
0.29 (0.06) 0.30 (0.01) 0.30 (0.01) 0.24 (0.01)
0.29 (0.06) 0.67 (0.02) 0.99 (0.03) 0.58 (0.02)
26.5 (0.1) 24.59 (0.03) 22.67 (0.03) 22.86 (0.03)
26.5 (0.1) 24.63 (0.03) 24.09 (0.03) 23.63 (0.03)
Standard errors (n = 3) are given in parentheses.
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Table 2 Particle-size distribution of trace metals, Cnative and Cmaize in the control soil Cu (g m2)
1974 1981 1993 1998
Pb (g m2)
Zn (g m2)
Cnative (kg m2)
Cmaize (kg m2)
Coarse sand
Fine sand
Silt clay
Coarse sand
Fine sand
Silt clay
Coarse sand
Fine sand
Silt clay
Coarse sand
Fine sand
Silt clay
3.1 (0.9) 3.0 (0.4) 1.6 (0.3) 0.3 (0.0)
0.8 (0.4) 0.8 (0.2) 0.7 (0.2) 0.8 (0.2)
3.5 (0.9) 3.5 (0.5) 3.3 (0.6) 3.7 (0.7)
3.6 (0.6) 2.7 (0.2) 3.9 (0.3) 2.6 (0.2)
1.5 (0.5) 1.2 (0.2) 1.7 (0.3) 1.2 (0.2)
5.5 (0.9) 6.8 (0.6) 6.6 (0.6) 3.7 (0.3)
1.5 (0.2) 1.4 (0.1) 2.9 (0.2) 5.0 (0.4)
1.6 (0.5) 0.7 (0.1) 1.9 (0.3) 1.8 (0.3)
2.0 (0.4) 1.7 (0.2) 4.5 (0.4) 3.5 (0.3)
1.1 (0.3) 0.9 (0.2) 0.79 (0.04) 0.35 (0.04)
0.7 (0.3) 0.5 (0.2) 0.40 (0.06) 0.6 (0.1)
2.5 (0.7) 2.3 (0.7) 2.1 (0.2) 1.7 (0.2)
Coarse sand 0 0.26 (0.03) 0.41 (0.02) 0.15 (0.02)
Fine sand 0 0.13 (0.05) 0.22 (0.03) 0.36 (0.08)
Silt clay 0 0.17 (0.05) 0.44 (0.03) 0.50 (0.07)
Standard errors (n = 3) are given in parentheses.
Fig. 1. Trace metal variations in particle-size fractions of sludge-treated soil.
1993, the lack of recently applied sludge induced a loss of Cu and Pb from the coarse fraction, but only a transfer of Zn from the coarse to the fine fractions. 3.1.3. Trace metal speciation in sludge sample and S100 soil In order to determine if there was also a change in the chemical forms of metal simultaneously to their transfer from the coarse to the fine fractions, sequential extractions were performed on one sample of sewage sludge and on all soil samples. It appeared that between 1974 and 1993, sludge applications induced an accumulation of Cu in the TAMAs and OMOCl fractions, whereas Pb was mainly localized in the RESID fraction and Zn in the TAMAs fraction (Fig. 2). The comparison of the Pb speciation in soil with that in the sludge did not show strong differences, contrary to Cu and Zn. A partial shift was observed from
the TAMAs fraction to the OMOCl fraction in the case of Cu and from the TAMAs fraction to the NH2OH fraction in the case of Zn. When the sewage sludge applications ceased in 1993, immediate losses of Cu and Pb occurred mainly at expense of the RESID fraction for Pb, and of both TAMAs and OMOCl fractions for Cu. This would suggest that organic matter and iron oxides are the main components involved in the trace metal behavior. A Cu distribution change has been also observed in a similar study, with an increase in the more labile metal forms, without concentration increase in the shoot (Mendoza et al., 2006; Kidd et al., 2007). Between 1993 and 1998, the total iron amounts decreased from 3122 to 2497 g m2. The formation of some organo-metallic complexes or some colloidal forms of organic matter or iron oxides could be involved in leaching of Cu and Pb (Kiekens, 1983; Jensen et al., 1999; Denaix
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Trace metal speciation in sludge 100% 90% 80% 70% 60% 50% 40% 30% 20% 10% 0%
40
TAMAs NH2OH
Cu (g m-2)
35 RESID
OMOC
30 25 20 15 10
MgNI
5 0 Cu (%)
Pb (%)
Zn (%))
120
1974
1981
1993
1999
1981
1993
1999
400 350
100
Zn (g m-2)
Pb(g m-2)
300 80 60 40
250 200 150 100
20 0
50 1974
1981
1993
1999
0
1974
Fig. 2. Variations of Cu, Pb and Zn speciation in added sludge and sludge-treated soil.
et al., 2001). This suggests that the addition of organic matter derived from sludge can have a decisive impact on the fate of trace metals over time.
Table 3 Variations of different sources of organic carbon (g m2), between 1974 and 1993, in control and S100 soils Control
3.2. Organic matter accumulation in soils 3.2.1. Variations of total organic carbon, nitrogen and d13C in control and S100 soil Organic carbon and organic nitrogen did not vary significantly in the control soil between 1974 and 1993 (Table 1), whereas they decreased after 1993 although maize grain yield did not change over time. This suggested that after 1993, the maize residues are not sufficient to overcome the natural losses in organic matter. The sewage sludge applications induced a significant accumulation of organic carbon and nitrogen in the surface soil layer. After a last application, a fast decrease in organic carbon and nitrogen took place to reach 7.5 and 0.58 g m2 respectively, in 1998. The d13C values of the whole soil organic matter increased continuously in both the control and the S100 soils (Table 1), which indicates an increasing incorporation of maize C from 1974 to 1993. 3.2.2. Variations of each organic carbon source in control and S100 soil The quantification of each organic carbon source showed strong variations over the 1974–1998 period (Table 3). In the control, the maize cultivation, started in 1974, induced a substitution of native carbon by organic carbon derived from maize, which accounted for 27% of total carbon in 1998. In the S100 soil, the increase in total organic carbon appeared mainly due to the incorporation of sludge
1974 1981 1993 1998
S100
Cnative
Cmaize
Cnative
Cmaize
Csludge
4.4 (0.2) 3.8 (0.2) 3.19 (0.09) 2.77 (0.08)
0.0 0.60 (0.03) 1.20 (0.03) 0.97 (0.03)
4.4 3.8 3.1 2.8
0.0 0.81 (0.07) 1.32 (0.09) 1.26 (0.08)
0.0 3.5 (0.1) 5.9 (0.2) 3.5 (0.1)
(0.2) (0.3) (0.2) (0.2)
Standard errors (n = 3) are given in parenthesis.
C, which reached 5.9 g m2 in 1993. The accumulation of maize C appeared slightly higher in the sludge-treated soil than in the control, whereas it has been shown that microbial populations in the S100 soil were more efficient than those in the control (Parat et al., 2005). When the same calculation was performed on particlesize fractions, it was observed (Table 2) that the depletion of native carbon in the control was mainly localized in the coarse sand fraction (68% in 24 years). The maize carbon accumulated mainly in the finest fraction in agreement with other studies (Balesdent et al., 1987; Bonde et al., 1992; Desjardins et al., 1994). In 1998, about 50% of maize carbon (0.44 kg m2) was present in this fine fraction. In the particle-size fractions of the sludge-treated soil, the incorporation of both maize and sludge carbon predominated in the coarse sand fraction (Fig. 3). Moreover, the cumulated amounts of maize carbon in this fraction between 1974 and 1993 were clearly higher in the sludgetreated soil than in the control, suggesting that the sludge has a role in the accumulation of maize carbon in the coarse fraction. Indeed, both sludge carbon and maize
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Fig. 3. Variations of maize and sludge C in particle-size fractions of sludge-treated soil.
carbon significantly decreased in this coarse sand fraction once sewage sludge applications have ceased: 30% of maize carbon and 73% of sludge carbon in 5 years. Simultaneously, an increase in both, sludge carbon and maize carbon, was observed in the silt and clay fraction, suggesting that microdivision of the sludge material predominated over mineralisation at least the first five years. 3.2.3. Morphological study The morphological studies of coarse sand and fine sand fractions revealed the presence of water-stable macroaggregates (Fig. 4a–c) composed of organic matter fragments, mineral particles and sludge particles (Fig. 4d and e). Many physical effects have been already reported after sludge
applications, such aggregation (Epstein, 1975; Pagliai et al., 1981; Metzger et al., 1987; Lindsay and Logan, 1998) which is considered as an important mechanism for soil carbon stabilization (Adu and Oades, 1978; Oades, 1984; Balesdent et al., 2000; Six et al., 2000; Denef et al., 2001). The introduction of a quite stable particulate organic material through sewage sludge could help forming macroaggregates in the sand fraction, protecting thus the maize derived organic carbon against biodegradation, and allowing its sustained incorporation in the coarse sand fraction. Similarly Puget et al. (1995) showed that in silty cultivated soils, stable macroaggregates containing young crop residues were responsible for carbon contents in coarse sand fractions.
Fig. 4. Morphological study of S100 soil fractions from 1993: (a) coarse sand fraction, (b) fine sand fraction and (c) silt and clay fraction; (d) cross-section of a macroaggregate in natural light and (e) in polarized light (OMF: organic matter fragment, Q: Quartz, SP: sludge particle); morphological study of S100 soil fractions from 1998: (f) coarse sand fraction, (g) fine sand fraction and (h) silt and clay fraction.
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The morphological study of particle-size fractions of the 1998 samples showed that the termination of sludge applications induced changes as water-stable macroaggregates had disappeared from the coarse and fine sand fractions (Fig. 4f and g) and that microaggregates predominated in the finest fraction (Fig. 4h). These morphological changes coincided with the quantitative changes observed in sludge-derived organic carbon, which also shifted from the coarse to the fine fractions. According to Tisdall and Oades (1982), macroaggregates (>250 lm) would result from the binding of microaggregates by transient or temporary organic matter such as roots and hyphae, microaggregates being stabilized by more persistent organic matter. According to this, the addition of sludge would be a much more efficient agent of macroaggregate formation and stability than the maize-derived organic material, probably because of the sandy texture and low organic matter content of this soil at the begin of the experiment. The depletion of maize-derived organic carbon after 1993 in the coarse sand fraction of the sludge-treated soil strengthens the idea that this organic carbon source was protected against biodegradation by microorganisms over the sludge application period. The analysis of these aggregates by SEM–EDS showed that the trace metal distribution was strongly related to macroaggregates (not shown). After stopping the sewage sludge treatments and therefore the input of stable particulate organic material, the previously introduced particulate organic matter was decomposed, the macro-aggregates disintegrated and the trace metals bound in the macro-aggregates were lost. In a similar way, Besnard et al. (2001) showed that organic amendments had a direct effect on the retention of Cu in soil due to particulate organic matter, and Balabane et al. (1999) showed a process of mutual protection of plant debris towards biodegradation under a metallophyte grassland.
4. Conclusion This study showed that repeated applications of sludge strongly influenced soil physical properties and consequently trace metal retention or mobility. Repeated applications of sludge in a sandy soil led to form macroaggregates that incorporated maize-derived organic matter and on which trace metals were present in detectable amounts. After sludge applications ceased, mineralization and humification processes of sludge-derived organic matter continued, resulting in a strong modification of the soil structure with a progressive disappearance of the recently formed macroaggregates. As a consequence, a shift of sludge-derived organic carbon from coarse fractions to the finest fraction occurred together with high losses in Cu and Pb and a redistribution of Zn between fractions. The amount and state of decomposition of sludge-derived organic matter appeared therefore as a key parameter determining the fate of trace metals.
Acknowledgements This work was supported by a Ph.D. Grant from the Agence de l’Environnement et de la Maıˆtrise de l’Energie (ADEME) and the National Institute of Agronomy Research (INRA). We thank the INRA of Bordeaux, especially the agronomy unit for their collaboration. References Adani, F., Tambone, F., 2005. Long-term effect of sewage sludge application on soil humic acids. Chemosphere 60, 1214–1221. Adu, J.K., Oades, J.M., 1978. Physical factors influencing the decomposition of organic materials in soil aggregates. Soil Biol. Biochem. 10, 109–115. Alloway, B.J., 1995. Heavy Metals in Soils. Blackie Academic & Professional, London. Balabane, M., Faivre, D., Van Oort, F., Dahmani-Muller, H., 1999. Mutual effects of soil organic matter dynamics and heavy metals fate in a metallophyte grassland. Environ. Pollut. 105, 45–54. Balesdent, J., Chenu, C., Balabane, M., 2000. Relationship of soil organic matter dynamics to physical protection and tillage. Soil Till. Res. 53, 215–230. Balesdent, J., Mariotti, A., Guillet, B., 1987. Natural 13C abundance as a tracer for studies of soil organic matter dynamics. Soil Biol. Biochem. 19, 25–30. Bartoli, F., Burtin, G., Herbillon, A.J., 1991. Disaggregation and clay dispersion of oxisols: Na resin, a recommended methodology. Geoderma 49, 301–317. Besnard, E., Chenu, C., Robert, M., 2001. Influence of organic amendments on copper distribution among particle-size and density fractions in Champagne vineyard soils. Environ. Pollut. 112, 329–337. Bonde, T.A., Christensen, B.T., Cerri, C.C., 1992. Dynamics of soil organic matter as reflected by natural 13C abundance in particle size fractions of forested and cultivated oxisols. Soil Biol. Biochem. 24, 275–277. Chang, A.C., Hyun, H.N., Page, A.L., 1997. Evaluating ten years of Cd uptake data for Swiss chard grown on composted sewage sludge-treated field plots: Plateau or a time bomb? J. Environ. Qual. 26, 11–19. Denaix, L., Semlali, R.M., Douay, F., 2001. Dissolved and colloidal transport of Cd, Pb and Zn in a silt loam soil affected by atmospheric industrial deposition. Environ. Pollut. 114, 29–38. Denef, K., Six, J., Paustian, K., Merckx, R., 2001. Importance of macroaggregate dynamics in controlling soil carbon stabilization: short-term effects of physical disturbance induced by dry-wet cycles. Soil Biol. Biochem. 33, 2145–2153. Desjardins, T., Andreux, F., Volkoff, B., Cerri, C.C., 1994. Organic carbon and 13C contents in soils and soil size-fractions, and their changes due to deforestation and pasture installation in eastern Amazonia. Geoderma 61, 103–118. do Valle, C.M., Santana, G.P., Augusti, R., Egreja Filho, F.B., Windmoller, C.C., 2005. Speciation and quantification of mercury in Oxisol, Ultisol, and Spodosol from Amazon (Manaus, Brazil). Chemosphere 58, 779–792. Epstein, E., 1975. Effect of sewage sludge on some soil physical properties. J. Environ. Qual. 4, 139–142. Essington, M.E., Mattigod, S.V., 1990. Element partitioning in size and density-fractionated sewage sludge and sludge-amended soil. Soil Sci. Soc. Am. J. 54, 385–394. Jensen, D.L., Ledin, A., Christensen, T.H., 1999. Speciation of heavy metals in landfill-leacheate polluted groundwater. Water Res. 33, 2642–2650. Juste, C., Soldaˆ, P., 1986. Effect on a long term sludge disposal on cadmium and nickel toxicity to a continuous maize crop. In: L’Hermite, P. (Ed.), Processing and use of organic sludge and liquid agricultural wastes. Springer-Verlag, New York, pp. 336–347.
C. Parat et al. / Chemosphere 69 (2007) 636–643 Kidd, P.S., Dominguez-Rodriguez, M.J., Diez, J., Monterroso, C., 2007. Bioavailability and plant accumulation of heavy metals and phosphorus in agricultural soils amended by long-term application of sewage sludge. Chemosphere 66, 1458–1467. Kiekens, L., 1983. Behaviour of heavy metals in soils. In: Berglund, S., Davis, R.D., L’Hermite, P. (Eds.), Utilisation of Sewage Sludge on Land: Rates of Application and Long-Term Effects of Metals. Reidel Publish Co, Boston, MA, pp. 126–134. Lindsay, B.J., Logan, T.J., 1998. Field response of soil physical properties to sewage sludge. J. Environ. Qual. 27, 534–542. McBride, M.B., 1995. Toxic metal accumulation from agricultural use of sludge: are USEPA regulations protective? J. Environ. Qual. 24, 5–18. McGrath, S.P., Lane, P.W., 1989. An explanation for the apparent losses of metals in a long-term field experiment with sewage sludge. Environ. Pollut. 60, 235–256. Mendoza, J., Garrido, T., Martin, G., Castillo, N.S., 2006. Metal availability and uptake by sorghum plants grown in soils amended with sludge from different treatments. Chemosphere 65, 2304–2312. Metzger, L., Levanon, D., Mingelgrin, U., 1987. The effect of sewage sludge on soil structural stability: microbiological aspects. Soil Sci. Soc. Am. J. 51, 346–351. Gerzabek, M., Novoszad, M.H., Haberhauer, G., Jakusch, M., Lischka, H., Tunega, D., Kirchmann, H., 2005. Sorption of naphthalene derivatives to soils from a long-term field experiment. Chemosphere 59, 639–647. Oades, J.M., 1984. Soil organic matter and structural stability: mechanisms and implications for management. Plant Soil. 76, 319–337.
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Pagliai, M., Guidi, G., La Marca, M., Giachetti, M., Lucamante, G., 1981. Effect of sewage sludges and composts on soil porosity and aggregation. J. Environ. Qual. 4, 556–561. Parat, C., Chaussod, R., Le´veˆque, J., Andreux, F., 2005. Long-term effects of metal-containing farmyard manure and sewage sludge on soil organic matter in a fluvisol. Soil Biol. Biochem. 37, 673–679. Parat, C., Le´veˆque, J., Andreux, R., Chaussod, F., 2007. Sludge-derived organic carbon in an agricultural soil estimated by 13C abundance measurements. Eur. J. Soil Sci. 58, 166–173. Parat, C., Le´veˆque, J., Dousset, S., Chaussod, R., Andreux, F., 2003. Comparison of three sequential extraction procedures to study trace metal distribution in an acidic sandy soil. Anal. Bioanal. Chem. 376, 243–247. Puget, P., Chenu, C., Balesdent, J., 1995. Total and young organic matter distributions in aggregates of silty cultivated soils. Eur. J. Soil Sci. 46, 449–459. Shuman, L.M., 1985. Fractionation method for soil microelements. Soil Sci. 140, 11–22. Shuman, L.M., 1999. Organic waste amendments effect on zinc fractions of two soils. J. Environ. Qual. 28, 1442–1447. Six, J., Elliot, E.T., Paustian, K., 2000. Soil macroaggregate turnover and microaggregate formation: a mechanism for C sequestration under no-tillage agriculture. Soil Biol. Biochem. 32, 2099–2103. Tisdall, J.M., Oades, J.M., 1982. Organic matter and water-stable aggregates in soils. J. Soil Sci. 33, 141–163.
The organic carbon derived from sewage sludge as a key parameter determining the fate of trace metals C. Parat a
a,*
, L. Denaix b, J. Le´veˆque a, R. Chaussod c, F. Andreux
a
Microbiologie et Ge´ochimie des Sols, UMR 1229, INRA-Universite´ de Bourgogne, Centre des Sciences de la Terre 6 Bd Gabriel 21000 Dijon, France b INRA Bordeaux—Unite´ d’agronomie, BP 81 33883, Villenave d’Ornon Cedex, France c Microbiologie et Ge´ochimie des Sols, UMR 1229, INRA-Universite´ de Bourgogne 17, rue de Sully 21034 Dijon Cedex, France Received 9 July 2006; received in revised form 23 February 2007; accepted 26 February 2007 Available online 18 April 2007
Abstract In a sandy agricultural soil of south-west of France, continuously cultivated with maize and amended with sewage-sludge over 20 years, the behavior of three trace metals (Cu, Pb, and Zn) was studied during the sludge applications (1974–1993) and after its cessation (1993–1998). Using the d13C analysis, the dynamics of different sources of organic matter were followed in order to elucidate the influence of the sludge-derived organic matter on the fate of trace metals in the soil and its particle size fractions. This study revealed that sludge-derived organic matter contributed to the formation of macroaggregates through the binding of preexisting microaggregates. These macroagreggates were thus responsible for the accumulation of trace metals in the coarsest fraction as well as for the protection of maize-derived organic matter against biodegradation. After sludge application ceased, the disaggregation of macroaggregates occurred simultaneously with high losses in Cu and Pb. On the contrary, Zn appeared less affected by the cessation of sludge application, with only a location change from coarse to fine fractions. Ó 2007 Elsevier Ltd. All rights reserved. Keywords: Sewage-sludge; Trace metals; Organic carbon; d13C; Particle-size fractions
1. Introduction The agricultural use of sewage-sludge has been a common practice in waste disposal in recent decades. Sewagesludge contains plant nutrients and organic matter, and therefore may significantly modified soil organic matter, pore sizes and aggregate stability (Adani and Tambone, 2005; Novoszad et al., 2005). But their positive influence may be limited by their contents of potentially harmful substances such as organic micropollutants and trace met-
*
Corresponding author. Present address: Laboratoire de Chimie Analytique, LCABIE, UMR 5254 IPREM, UFR Sciences, Avenue de l’Universite´, 64000 Pau, France. Tel.: +33 3 559 40 76 70; fax: +33 3 559 40 76 74. E-mail address: [email protected] (C. Parat). 0045-6535/$ - see front matter Ó 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2007.02.069
als (Kidd et al., 2007). In last years, the concentrations of most metals in sludge in many countries have shown a pronounced decrease as a result of improved effluent control and waste minimisation (Alloway, 1995). However, in view of the high amount of sludge applied to land in the past and the very long residence time of trace metals in soil surface (McGrath and Lane, 1989), it is still important to be aware of trace metal amounts added with sludge. Current debates focus on two hypotheses: some advocate the sludge protection hypothesis, which states that trace metals are bound within sludge components in forms that are poorly available, and that the specific adsorption capacity of sludge will persist as long as trace metals of concern persist in soil (Shuman, 1999). Others consider that the slow mineralization of sludge-derived organic matter could release metals as more soluble forms (McBride, 1995), an effect often termed ‘‘the sludge time bomb hypothesis’’ (Chang et al., 1997). There is therefore a need
C. Parat et al. / Chemosphere 69 (2007) 636–643
to know the kinetic behavior of trace metals following sludge applications, as well as after the termination of sludge applications. In this study, we investigated the behavior of three trace metals (Cu, Pb and Zn) in the long-term field experiment laid out in south-west France by the National Institute of Agronomic Research (INRA, Bordeaux) and that began in 1974. Previous work has shown an improvement in microbial biomass in spite of a higher trace metal contamination (Parat et al., 2005). A kinetic study was therefore undertaken in order to elucidate the relationship between sludge-contained trace metals and sludge-derived organic matter. A strength of this study is that complementary approaches (physical and chemical fractionation) were used on stored soils samples collected prior to the start of the experiment in 1974, but also during sewage sludge applications and after their termination. 2. Materials and methods 2.1. Study site Plots of 18 m2 have been established in 1974 at the INRA experimental site of Couhins, near Bordeaux (France). The experimental site consists of four treatments, laid out with five replicates in a randomized block design, the only difference between plots being the type of amendment. A slope of 3% has induced a contamination between blocks, which led us to work on the block no. 5 located at the top of the site. The soil is a slightly acidic sandy soil developed on colluvial material (Fluvisol, FAO classification), with initial contents of organic carbon and nitrogen of 13 ± 2 and 0.90 ± 0.01 g kg1 of air-dry soil, respectively, a pH of 6.0 and a bulk density of 1.36 ± 0.02 g cm3. Before 1974, this soil was always cropped with C3 species (vineyard). Since 1974, it has been continuously cropped with maize (Zea mays L. cv. INRA 260). Anaerobically digested sewage sludge, strongly contaminated with Cu (215 mg kg1 of air-dry sludge), Pb (752 mg kg1 of airdry sludge) and particularly Zn (5052 mg kg1 of air-dry sludge), was incorporated from 1974 to 1993 at a rate of 100 t ha1 every 2 years (S100 soil). This rate represented an organic carbon input of 1.85 kg C m2 every year. The control, which has never received any sludge treatment, was adjusted annually, prior to planting, to the same level of mineral fertilisation (N, P and K) as sludge-treated soil (Juste and Soldaˆ, 1986). Every year, the maize residues were buried in the soil. Between 1974 and 1998, cumulative maize grain yield was 870 ± 80 t ha1 in the control plots and 930 ± 90 t ha1 in the S100 plots. Since 1993, no treatments have been applied. Six soil-cores in the plough-layer (0–25 cm) were taken from each plot before maize planting and mixed to form a representative sample of each plot. Representative soil samples were collected in March 1974, 1981, 1993 and 1998 from the two treatments, airdried and passed through a 2-mm aperture sieve, and stored in plastic sealed containers.
637
2.2. Particle-size fractionation Particle-size fractionation was carried out by the procedure of Bartoli et al. (1991); on the whole soil (<2 mm soil fraction) and on the sewage-sludge samples. Briefly, for each sample, 20 g of material were placed in a glass flask with 600 ml of distilled water, and shaken mechanically for 16 h. Three fractions were obtained by wet sieving: the coarse sand fraction (200–2000 lm), fine sand fraction (50–200 lm), and the silt and clay fraction (0–50 lm), each then dried at 40 °C for 24 h and ground prior to analysis. 2.3. Trace metal analysis Total metal contents (Cu, Pb and Zn) were determined after acid digestion according to the following procedure: soil samples were first crushed and dried at 105 °C during 12 h. Aliquots of 0.1 g were digested in a Teflon bomb with a solution of concentrated HNO3 (3 ml), HClO4 (2 ml) and HF (3 ml) during 16 h. Then, the mixture was evaporated to dryness, 5 ml of deionized water was added, and evaporated to dryness. Finally, the residue was dissolved in 1 ml of concentrated HNO3 and filtered. The solution was then transferred into a 100 ml flask. Contents of Cu, Zn, and Pb were determined by graphite furnace atomic absorption spectrometry (GFAAS, HGA 600). 2.4. Sequential extraction The sequential extraction procedure of Shuman (1985) was performed on all soil samples and one of sludge samples (Parat et al., 2003) and allowed to isolate the following fractions: MgNIT (exchangeable forms), OMOCl (oxidizable organic/sulfide fraction), NH2OH (manganese oxides), TAMAs (iron oxides) and RESID (residual). The technique involved mixing and shaking 1.0 g of airdried soil/sludge previously screened at 2 mm with 10 ml of reagent, followed by centrifugation and washing with 40 ml of deionized water. The residue was dried at 40 °C before the next step. The obtained extracts were transferred into a 100 ml flask. Copper, Pb and Zn were measured by inductively coupled plasma mass spectrometry (ICP–MS). 2.5. C, N, and carbon isotope analyses Carbon and nitrogen measurements were made on 30 mg of ground sample (of both, whole sample and particle size fractions) by the dry combustion method (Carlo Erba CNS 1500 analyser, Dijon, France). The standard used was sulphanilamide (41.84% of C and 16.27% of N). The natural abundance of stable carbon isotope 13C was measured by using a continuous flow mass spectrometer (VG Isochrom-EA, Dijon, France) on ground samples. The variations of soil isotopic signature depend on the organic inputs (Balesdent et al., 1987). Consequently, when dealing with a mixture of carbon originating from different
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vegetations, their respective proportions are calculated using Eqs. (1) and (2): X total ¼ X native þ X maize þ X sludge 13
ð1Þ
13
13
d Csample ¼ X native d Cnative þ X maize d Cmaize 13
þ X sludge d Csludge
where DC is the standard deviation of the concentration C in g kg1 for carbon or in mg kg1 for trace metals, and Dd is the standard deviation of the soil density d. 3. Results and discussion
ð2Þ 3.1. Trace metal accumulation in control and S100 soil
where Xmaize, Xnative and Xsludge are the proportions of maize C, native C and sludge C, at the sampling date, respectively, and d13Cnative (26.5&), d13Cmaize (12.5&), d13Csludge (25.4&) are the mean isotopic values measured in the respective sources. To solve Eqs. (1) and (2), the control soil was used to calculate the contribution of native organic carbon in sludge-treated soil (Parat et al., 2007). 2.6. Morphological study The morphology of the particle-size fractions of the control and S100 soil was studied using a binocular and a scanning electronic microscope (ASEM, JEOL JSM-6400F) interfaced with an energy dispersive X-ray spectrometer (SEM–EDS). 2.7. Amount and error calculations The measurements were all done in triplicate from composite soil samples. Mean and standard errors are reported in tables. Amount calculations were obtained by using the following equation: S whole soil ¼ C whole soil d t
ð3Þ
where Swhole soil is the amount in the whole soil in kg m2 for carbon or g m2 for trace metals, Cwhole soil is the total concentration in g kg1 for carbon or mg kg1 for trace metals, d is the density (here, 1.36 g cm3) and t is the soil thickness (here, 0.25 m). The errors DSwhole soil on the amount Swhole soil have been determined by using the following equation: DC whole soil Dd DS whole soil ¼ S whole soil þ ð4Þ d C whole soil
3.1.1. Variations of total trace metals Between 1974 and 1998, small variations of trace metal appeared in the control soil contrary to the sludge-treated soil which showed a significant accumulation in the surface soil layer till 1993 (Table 1), due to the high content of trace metals in sewage sludge. Since 1993, when sewage sludge applications were stopped, a significant decrease of Cu (47%) and Pb (56%) and a slighter decreased of Zn (16%) were observed, whereas many authors showed that Cu and Pb are less mobile than Zn (Alloway, 1995). It is therefore necessary to relate such a differential decrease to the variation of other soil and sludge components. 3.1.2. Trace metal distribution in particle-size fractions The particle-size fractionation of soil samples resulted in the recovery of 97–99% of the initial sample mass, with similar size distributions for the different plots. At the start of the experiment, the control soil showed a predominance of Cu and Pb in coarse sand and silt and clay fractions, whereas Zn was equally distributed in the three fractions (Table 2). There was no significant variation of trace metal distribution with time of cultivation in any fractions, except in the case of Cu which strongly decreased in the coarsest fraction since 1993. In 1998, the largest amounts of Cu and Pb were found in the silt and clay fraction, whereas Zn was mainly located in the coarse sand fraction. Surprisingly, the repeated application of sewage sludge induced a significant accumulation of trace metals in the coarse sand fraction (Fig. 1). In 1993, the amount of trace metals in the coarse sand fraction was 2–6 times higher than in the fine sand and silt and clay fractions. This is in contradiction with many studies, which showed that trace metals mainly accumulate in the finest fractions with high concentrations of clay minerals and colloidal organic matter (Essington and Mattigod, 1990; do Valle et al., 2005). Since
Table 1 Variations of trace metals, organic carbon, nitrogen, and d13C between 1974 and 1998, in control and S100 soils Cu (g m2)
1974 1981 1993 1998
Pb (g m2)
Zn (g m2)
Carbon (kg m2)
Nitrogen (kg m2)
d13C (&)
Control
S100
Control
S100
Control
S100
Control
S100
Control
S100
Control
S100
7.3 (0.9) 7.3 (0.5) 5.6 (0.4) 4.8 (0.3)
7.3 (0.9) 22 (1) 40 (3) 21 (1)
10.6 (0.7) 10.7 (0.4) 12.1 (0.4) 7.5 (0.3)
10.6 (0.7) 66 (2) 101 (4) 44 (2)
5.1 (0.4) 3.7 (0.1) 9.2 (0.3) 10.3 (0.4)
5.1 (0.4) 176 (6) 351 (12) 294 (10)
4.4 (0.2) 4.39 (0.08) 4.34 (0.08) 3.74 (0.07)
4.4 (0.2) 8.1 (0.2) 10.4 (0.2) 7.5 (0.1)
0.29 (0.06) 0.30 (0.01) 0.30 (0.01) 0.24 (0.01)
0.29 (0.06) 0.67 (0.02) 0.99 (0.03) 0.58 (0.02)
26.5 (0.1) 24.59 (0.03) 22.67 (0.03) 22.86 (0.03)
26.5 (0.1) 24.63 (0.03) 24.09 (0.03) 23.63 (0.03)
Standard errors (n = 3) are given in parentheses.
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639
Table 2 Particle-size distribution of trace metals, Cnative and Cmaize in the control soil Cu (g m2)
1974 1981 1993 1998
Pb (g m2)
Zn (g m2)
Cnative (kg m2)
Cmaize (kg m2)
Coarse sand
Fine sand
Silt clay
Coarse sand
Fine sand
Silt clay
Coarse sand
Fine sand
Silt clay
Coarse sand
Fine sand
Silt clay
3.1 (0.9) 3.0 (0.4) 1.6 (0.3) 0.3 (0.0)
0.8 (0.4) 0.8 (0.2) 0.7 (0.2) 0.8 (0.2)
3.5 (0.9) 3.5 (0.5) 3.3 (0.6) 3.7 (0.7)
3.6 (0.6) 2.7 (0.2) 3.9 (0.3) 2.6 (0.2)
1.5 (0.5) 1.2 (0.2) 1.7 (0.3) 1.2 (0.2)
5.5 (0.9) 6.8 (0.6) 6.6 (0.6) 3.7 (0.3)
1.5 (0.2) 1.4 (0.1) 2.9 (0.2) 5.0 (0.4)
1.6 (0.5) 0.7 (0.1) 1.9 (0.3) 1.8 (0.3)
2.0 (0.4) 1.7 (0.2) 4.5 (0.4) 3.5 (0.3)
1.1 (0.3) 0.9 (0.2) 0.79 (0.04) 0.35 (0.04)
0.7 (0.3) 0.5 (0.2) 0.40 (0.06) 0.6 (0.1)
2.5 (0.7) 2.3 (0.7) 2.1 (0.2) 1.7 (0.2)
Coarse sand 0 0.26 (0.03) 0.41 (0.02) 0.15 (0.02)
Fine sand 0 0.13 (0.05) 0.22 (0.03) 0.36 (0.08)
Silt clay 0 0.17 (0.05) 0.44 (0.03) 0.50 (0.07)
Standard errors (n = 3) are given in parentheses.
Fig. 1. Trace metal variations in particle-size fractions of sludge-treated soil.
1993, the lack of recently applied sludge induced a loss of Cu and Pb from the coarse fraction, but only a transfer of Zn from the coarse to the fine fractions. 3.1.3. Trace metal speciation in sludge sample and S100 soil In order to determine if there was also a change in the chemical forms of metal simultaneously to their transfer from the coarse to the fine fractions, sequential extractions were performed on one sample of sewage sludge and on all soil samples. It appeared that between 1974 and 1993, sludge applications induced an accumulation of Cu in the TAMAs and OMOCl fractions, whereas Pb was mainly localized in the RESID fraction and Zn in the TAMAs fraction (Fig. 2). The comparison of the Pb speciation in soil with that in the sludge did not show strong differences, contrary to Cu and Zn. A partial shift was observed from
the TAMAs fraction to the OMOCl fraction in the case of Cu and from the TAMAs fraction to the NH2OH fraction in the case of Zn. When the sewage sludge applications ceased in 1993, immediate losses of Cu and Pb occurred mainly at expense of the RESID fraction for Pb, and of both TAMAs and OMOCl fractions for Cu. This would suggest that organic matter and iron oxides are the main components involved in the trace metal behavior. A Cu distribution change has been also observed in a similar study, with an increase in the more labile metal forms, without concentration increase in the shoot (Mendoza et al., 2006; Kidd et al., 2007). Between 1993 and 1998, the total iron amounts decreased from 3122 to 2497 g m2. The formation of some organo-metallic complexes or some colloidal forms of organic matter or iron oxides could be involved in leaching of Cu and Pb (Kiekens, 1983; Jensen et al., 1999; Denaix
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C. Parat et al. / Chemosphere 69 (2007) 636–643 45
Trace metal speciation in sludge 100% 90% 80% 70% 60% 50% 40% 30% 20% 10% 0%
40
TAMAs NH2OH
Cu (g m-2)
35 RESID
OMOC
30 25 20 15 10
MgNI
5 0 Cu (%)
Pb (%)
Zn (%))
120
1974
1981
1993
1999
1981
1993
1999
400 350
100
Zn (g m-2)
Pb(g m-2)
300 80 60 40
250 200 150 100
20 0
50 1974
1981
1993
1999
0
1974
Fig. 2. Variations of Cu, Pb and Zn speciation in added sludge and sludge-treated soil.
et al., 2001). This suggests that the addition of organic matter derived from sludge can have a decisive impact on the fate of trace metals over time.
Table 3 Variations of different sources of organic carbon (g m2), between 1974 and 1993, in control and S100 soils Control
3.2. Organic matter accumulation in soils 3.2.1. Variations of total organic carbon, nitrogen and d13C in control and S100 soil Organic carbon and organic nitrogen did not vary significantly in the control soil between 1974 and 1993 (Table 1), whereas they decreased after 1993 although maize grain yield did not change over time. This suggested that after 1993, the maize residues are not sufficient to overcome the natural losses in organic matter. The sewage sludge applications induced a significant accumulation of organic carbon and nitrogen in the surface soil layer. After a last application, a fast decrease in organic carbon and nitrogen took place to reach 7.5 and 0.58 g m2 respectively, in 1998. The d13C values of the whole soil organic matter increased continuously in both the control and the S100 soils (Table 1), which indicates an increasing incorporation of maize C from 1974 to 1993. 3.2.2. Variations of each organic carbon source in control and S100 soil The quantification of each organic carbon source showed strong variations over the 1974–1998 period (Table 3). In the control, the maize cultivation, started in 1974, induced a substitution of native carbon by organic carbon derived from maize, which accounted for 27% of total carbon in 1998. In the S100 soil, the increase in total organic carbon appeared mainly due to the incorporation of sludge
1974 1981 1993 1998
S100
Cnative
Cmaize
Cnative
Cmaize
Csludge
4.4 (0.2) 3.8 (0.2) 3.19 (0.09) 2.77 (0.08)
0.0 0.60 (0.03) 1.20 (0.03) 0.97 (0.03)
4.4 3.8 3.1 2.8
0.0 0.81 (0.07) 1.32 (0.09) 1.26 (0.08)
0.0 3.5 (0.1) 5.9 (0.2) 3.5 (0.1)
(0.2) (0.3) (0.2) (0.2)
Standard errors (n = 3) are given in parenthesis.
C, which reached 5.9 g m2 in 1993. The accumulation of maize C appeared slightly higher in the sludge-treated soil than in the control, whereas it has been shown that microbial populations in the S100 soil were more efficient than those in the control (Parat et al., 2005). When the same calculation was performed on particlesize fractions, it was observed (Table 2) that the depletion of native carbon in the control was mainly localized in the coarse sand fraction (68% in 24 years). The maize carbon accumulated mainly in the finest fraction in agreement with other studies (Balesdent et al., 1987; Bonde et al., 1992; Desjardins et al., 1994). In 1998, about 50% of maize carbon (0.44 kg m2) was present in this fine fraction. In the particle-size fractions of the sludge-treated soil, the incorporation of both maize and sludge carbon predominated in the coarse sand fraction (Fig. 3). Moreover, the cumulated amounts of maize carbon in this fraction between 1974 and 1993 were clearly higher in the sludgetreated soil than in the control, suggesting that the sludge has a role in the accumulation of maize carbon in the coarse fraction. Indeed, both sludge carbon and maize
C. Parat et al. / Chemosphere 69 (2007) 636–643
641
Fig. 3. Variations of maize and sludge C in particle-size fractions of sludge-treated soil.
carbon significantly decreased in this coarse sand fraction once sewage sludge applications have ceased: 30% of maize carbon and 73% of sludge carbon in 5 years. Simultaneously, an increase in both, sludge carbon and maize carbon, was observed in the silt and clay fraction, suggesting that microdivision of the sludge material predominated over mineralisation at least the first five years. 3.2.3. Morphological study The morphological studies of coarse sand and fine sand fractions revealed the presence of water-stable macroaggregates (Fig. 4a–c) composed of organic matter fragments, mineral particles and sludge particles (Fig. 4d and e). Many physical effects have been already reported after sludge
applications, such aggregation (Epstein, 1975; Pagliai et al., 1981; Metzger et al., 1987; Lindsay and Logan, 1998) which is considered as an important mechanism for soil carbon stabilization (Adu and Oades, 1978; Oades, 1984; Balesdent et al., 2000; Six et al., 2000; Denef et al., 2001). The introduction of a quite stable particulate organic material through sewage sludge could help forming macroaggregates in the sand fraction, protecting thus the maize derived organic carbon against biodegradation, and allowing its sustained incorporation in the coarse sand fraction. Similarly Puget et al. (1995) showed that in silty cultivated soils, stable macroaggregates containing young crop residues were responsible for carbon contents in coarse sand fractions.
Fig. 4. Morphological study of S100 soil fractions from 1993: (a) coarse sand fraction, (b) fine sand fraction and (c) silt and clay fraction; (d) cross-section of a macroaggregate in natural light and (e) in polarized light (OMF: organic matter fragment, Q: Quartz, SP: sludge particle); morphological study of S100 soil fractions from 1998: (f) coarse sand fraction, (g) fine sand fraction and (h) silt and clay fraction.
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The morphological study of particle-size fractions of the 1998 samples showed that the termination of sludge applications induced changes as water-stable macroaggregates had disappeared from the coarse and fine sand fractions (Fig. 4f and g) and that microaggregates predominated in the finest fraction (Fig. 4h). These morphological changes coincided with the quantitative changes observed in sludge-derived organic carbon, which also shifted from the coarse to the fine fractions. According to Tisdall and Oades (1982), macroaggregates (>250 lm) would result from the binding of microaggregates by transient or temporary organic matter such as roots and hyphae, microaggregates being stabilized by more persistent organic matter. According to this, the addition of sludge would be a much more efficient agent of macroaggregate formation and stability than the maize-derived organic material, probably because of the sandy texture and low organic matter content of this soil at the begin of the experiment. The depletion of maize-derived organic carbon after 1993 in the coarse sand fraction of the sludge-treated soil strengthens the idea that this organic carbon source was protected against biodegradation by microorganisms over the sludge application period. The analysis of these aggregates by SEM–EDS showed that the trace metal distribution was strongly related to macroaggregates (not shown). After stopping the sewage sludge treatments and therefore the input of stable particulate organic material, the previously introduced particulate organic matter was decomposed, the macro-aggregates disintegrated and the trace metals bound in the macro-aggregates were lost. In a similar way, Besnard et al. (2001) showed that organic amendments had a direct effect on the retention of Cu in soil due to particulate organic matter, and Balabane et al. (1999) showed a process of mutual protection of plant debris towards biodegradation under a metallophyte grassland.
4. Conclusion This study showed that repeated applications of sludge strongly influenced soil physical properties and consequently trace metal retention or mobility. Repeated applications of sludge in a sandy soil led to form macroaggregates that incorporated maize-derived organic matter and on which trace metals were present in detectable amounts. After sludge applications ceased, mineralization and humification processes of sludge-derived organic matter continued, resulting in a strong modification of the soil structure with a progressive disappearance of the recently formed macroaggregates. As a consequence, a shift of sludge-derived organic carbon from coarse fractions to the finest fraction occurred together with high losses in Cu and Pb and a redistribution of Zn between fractions. The amount and state of decomposition of sludge-derived organic matter appeared therefore as a key parameter determining the fate of trace metals.
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