Published online 9 January 2007
Published in J Environ Qual 36:44-52 (2007)
DOI: 10.2134/jeq2006.0039
© 2007 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
677 S. Segoe Rd., Madison, WI 53711 USA
TECHNICAL REPORTS
Heavy Metals in the Environment
Plant Availability of Zinc and Copper in Soil after Contamination with Brass Foundry Filter Dust
Effect of Four Years of Aging
Isabel Hilber,
Andreas Voegelin*,
Kurt Barmettler and
Ruben Kretzschmar
Soil Chemistry, Institute of Biogeochemistry and Pollutant Dynamics, ETH Zurich, ETH Zentrum CHN, CH-8092 Zurich, Switzerland
* Corresponding author (voegelin{at}env.ethz.ch)
Received for publication January 26, 2006.
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ABSTRACT
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We investigated the effect of 4 yr of aging of a noncalcareous soil contaminated with filter dust from a brass foundry (80% w/w ZnO, 15% w/w Cu0.6Zn0.4) on the chemical extractability of Zn and Cu and their uptake by barley (Hordeum vulgare L.), pea (Pisum sativum L.), and sunflower (Helianthus annus L.). Pot experiments were conducted with the freshly contaminated soil (2250 mg kg–1 Zn; 503 mg kg–1 Cu), with the contaminated soil aged for 4 yr in the field (1811 mg kg–1 Zn; 385 mg kg–1 Cu), and with the uncontaminated control soil (136 mg kg–1 Zn; 32 mg kg–1 Cu). In comparison with the uncontaminated soil, the growth of barley and pea was clearly reduced in both contaminated soils, while toxicity symptoms did not systematically vary from the freshly contaminated to the 4 yr aged soil. The sunflower did not grow in the contaminated soils. The slow oxidative dissolution of the brass platelets led to an increase in the solubility and the plant uptake of Cu from the freshly contaminated to the 4 yr aged soil. In an earlier study, we found that the fine-grained ZnO dissolved in the field soil within 9 mo and that about half of the released Zn was incorporated into a layered double hydroxide phase and about half was adsorbed to the soil matrix. These changes in Zn speciation did not lead to a reduction of the Zn contents in the shoots and roots of barley and pea grown in the aged soil as compared with the freshly contaminated soil.
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INTRODUCTION
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ELEVATED concentrations of Zn and Cu in contaminated soils may adversely affect soil microbial activity, soil fertility, and crop yield, and in severe cases lead to soil degradation and erosion (Adriano, 2001). Initially, the speciation of Zn and Cu and their bioavailability in a contaminated soil depend on the source and type of soil contamination. Examples for the contamination of soils with Zn and Cu in readily available dissolved form include the application of Bordeau mixture (CuSO4+Ca(OH)2) in winegrowing (Chaignon et al., 2002) and the contamination of soils from the runoff of zinced steel constructions (Zhang et al., 2004; Bertling et al., 2006). In sewage sludge, Zn and Cu are bound to the organic matter and to Al and Fe oxides (Basta et al., 2005). Whether the decomposition of the organic substance after sludge application results in an increase or decrease in trace metal availability is still under debate (McBride, 1995; McGrath et al., 2000; Basta et al., 2005). The disposal of dredged sediments or spills of sulfidic mine wastes lead to soil contamination with trace metal sulfides (Hita and Torrent, 2005; Panfili et al., 2005; Schuwirth et al., 2006). While sparingly soluble under anoxic conditions, Zn and Cu sulfides slowly dissolve in aerated soils leading to an increase in trace metal availability (Tack et al., 1996; Hita and Torrent, 2006). Around metal smelters and foundries, atmospheric deposition may lead to soil contamination with primary metal bearing minerals, slag particles, metallic particles, and metal oxides (Morin et al., 1999; Manceau et al., 2000; Juillot et al., 2003; Voegelin et al., 2005).
For plant uptake, free Zn and Cu must be transferred through the soil solution (Zhang et al., 2001; Zhang et al., 2004). The rates at which Zn and Cu are released from contaminant phases depend on the type of soil contamination (Whiting et al., 2001) and on the physicochemical properties of the affected soil. Adsorption and precipitation reactions may reduce the concentrations of Zn and Cu in the soil solution and thereby decrease their bioavailability (Basta et al., 2005). Copper has a high affinity for organic functional groups and its solubility in topsoils is often controlled by Cu complexation with soil organic matter (Adriano, 2001). In the case of Zn, recent studies showed that layered Zn-bearing precipitates of the phyllosilicate or the layered double hydroxide (LDH) type are quantitatively relevant species that form in contaminated soils with slightly acidic to neutral pH (Manceau et al., 2000; Juillot et al., 2003; Voegelin et al., 2005). To date, the effect of these precipitates on the uptake of Zn by plants has not been investigated in detail.
In a field study on the influence of metal pollution on soil ecosystem functioning, a noncalcareous neutral soil was artificially contaminated with the filter dust from a brass foundry (Menon et al., 2005). The dust consisted of zinc oxide (ZnO, 80% w/w) and brass (Cu0.6Zn0.4, 15% w/w). From an EXAFS spectroscopy study, we know that the ZnO completely dissolved within 9 mo after contamination. About half of the released Zn adsorbed to soil organic and inorganic sorbents, while the other half formed a new mineral of the LDH type (Voegelin et al., 2005). Formation of this type of phase has been postulated to be a possible mechanism for the reduction of the Zn and Ni bioavailability in contaminated soils, provided that the precipitates are chemically stable (Ford et al., 1999; Ford and Sparks, 2000; Nachtegaal et al., 2005; Basta et al., 2005). In the case of Zn-LDH and mixed ZnNi-LDH, recent studies suggest that those precipitates may readily dissolve under acidic conditions (Voegelin et al., 2002; Voegelin et al., 2003; Voegelin and Kretzschmar, 2005). Since plants may acidify the soil rhizosphere, it is currently not clear whether the formation of Zn-LDH in soils effectively leads to a reduction in its bioavailability. Regarding the speciation of Cu and its changes with time in the same contaminated soil, no spectroscopic data are available. However, the oxidative dissolution of metallic brass particles is expected to proceed more slowly than the dissolution of the soluble ZnO.
In the context of these recent results, the objective of the present study was to investigate whether changes in the speciation of Zn and Cu on contamination of the noncalcareous soil with brass foundry filter dust are reflected in changes in the Zn and Cu availability for plants. Barley (Hordeum vulgare L.), pea (Pisum sativum L.), and sunflower (Helianthus annus L.) were grown in a growth chamber in pots containing (1) the uncontaminated field soil, (2) the same field soil freshly contaminated with filter dust, (3) and the same soil which had been incubated for 4 yr in field lysimeters after contamination with the same filter dust. Growth parameters were recorded during 30 d after planting. After harvest, the total metal contents of shoot and root tissues were determined. Pore water was collected after plant harvest and selective sequential extractions were conducted with the soils prior and after plant growth to characterize differences in metal extractability.
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MATERIALS AND METHODS
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Soil Material and Contamination with Filter Dust
The soil used in this study was obtained from the interdisciplinary project Cell-to-Tree, which addressed the influence of trace metal contamination (Zn, Cu, Cd, Pb) on ecosystem functioning from the level of single cells to whole organisms (Menon et al., 2005). The noncalcareous soil contained 15 g kg–1 organic C. It had a loamy texture (360 g kg–1 sand, 490 g kg–1 silt, and 150 g kg–1 clay) and a mixed clay mineralogy (kaolinite, mica, chlorite, vermiculite, hydroxy-interlayered vermiculite). The soil pH was 6.5 (10 g of soil in 25 mL of 10 mM CaCl2) and the effective cation exchange capacity (ECEC) was 102 mmolc kg–1. The concentrations of Zn, Cu, Cd, and Pb in the uncontaminated soil are listed in Table 1. They were within the expected range for geogenic trace metal levels in uncontaminated soils. In the Cell-to-Tree project, the top 30 cm of field lysimeters were filled with this soil material. Half of the lysimeters were subsequently contaminated by manually introducing filter dust from a brass foundry. The filter dust contained about 654 g kg–1 Zn, 65 g kg–1 Cu, 12 g kg–1 Pb, and 0.3 g kg–1 Cd. It consisted of 850 g kg–1 zincite (ZnO), 100 g kg–1 brass (Cu0.6Zn0.4), and other minerals (<50 g kg–1) (Voegelin et al., 2005). While the ZnO consisted of a very fine powder, the irregularly shaped brass platelets in the filter dust had lateral extensions of 100 to 1000 µm and were several tens of µm thick (determined by µ-X-ray fluorescence mapping, data not shown). In addition to the filter dust, a minor amount of CdO was mixed into the soil, raising the total Cd content to 5.9 mg kg–1. During 4 yr, plants were grown in the lysimeters (Menon et al., 2005). Upon termination of the project, the contaminated and uncontaminated topsoil material was separately excavated.
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Table 1. Total concentrations (± standard error) of Zn, Cu, Cd, and Pb in the soils used for the growth experiments and guide and action values of the Swiss Ordinance relating to Impacts on the Soil (VBBO, 1998).
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For the current study, excavated contaminated and uncontaminated topsoil material was used. To investigate the effect of time on metal availability, part of the uncontaminated soil was freshly contaminated with Zn and Cu using the same filter dust that had been used for the lysimeter study. Thus, three batches of soil were used for the growth experiment in this study: (1) the uncontaminated soil, (2) the uncontaminated soil freshly polluted with filter dust, and (3) the soil which had been incubated for 4 yr in the field after contamination with the same filter dust. Total contents of Zn, Cu, Cd, and Pb in these three soils are listed in Table 1.
The contaminated soils were enriched with Zn, Cu, Cd, and Pb. The 4 yr contaminated soil from the lysimeter study contains more Cd than the freshly contaminated soil because it was spiked with a minor amount of CdO in addition to the filter dust. Cadmium competes with Zn for plant uptake (Adriano, 2001). Considering the extremely high Zn/Cd ratios of 1800 in the freshly contaminated and 300 in the 4 yr contaminated soil, no significant effects of Cd on plant growth were expected in our experiments. Also toxic effects of Pb were negligible in the present study because the Pb contents in the contaminated soils were much lower than those of Zn and Cu, and because Pb is less phytotoxic than Zn or Cu (Adriano, 2001). These considerations were supported by comparison of the contents of Zn, Cu, Cd, and Pb in relation to Swiss legal guide and action values (Table 1), which indicate that the contamination with Zn and Cu is more relevant than the levels of Cd and Pb in the contaminated soils. We therefore assume for this study that the contents of Zn and Cu and possible changes in their bioavailability are the factors that influence plant growth in the freshly and 4 yr contaminated soil in comparison with the uncontaminated control soil. Due to an error in the experimental protocol, the final Zn and Cu concentrations in the freshly contaminated soil were about one third higher than in the 4 yr aged soil from the lysimeter study (Table 1). This difference in total metal loading was considered in the interpretation of the experimental results.
Growth Experiment
Pot experiments were conducted with three crop plant species to cover differences in the uptake of trace metals between different groups of plants. Barley and sunflower were chosen as representatives of monocotyledonic and dicotyledonic crop plants, respectively. Pea was chosen as a legume and also allowed us to observe the influence of trace metals on root nodulation.
For each combination of plant and soil type, pots containing about 1.1 kg soil were planted in four replicates. Before planting the seedlings, an initial amount of N was added (8.4 mM NH4NO3 in 280 mL distilled water per pot, i.e., 55 mg kg–1 N). At the same time, the soil water content was brought to about 50% of the field capacity of the soil (255 mL kg–1). The wet soil was subsequently allowed to equilibrate for 3 d in the growth chamber before the seedlings were planted. Seeds were germinated on filter paper soaked with water. After 4 d, the seedlings had root lengths of 1 to 3 cm. At this stage, the pre-equilibrated pots were planted with either 30 seedlings of barley, 7 seedlings of pea, or 7 seedlings of sunflower. The plants were grown for 30 d in the growth chamber with a day/night rhythm of 16/8 h. The intensity of the artificial illumination was 9000 to 13 000 cd m–2, depending on the location on the table. To minimize effects arising from these differences, the pots were randomly rearranged on the table every 2 to 3 d. The temperature was 21°C during the day and 16°C during the night. The relative humidity was set to 60%.
Every 2 to 3 d, the soil water content was readjusted to 255 mL kg–1 using deionized water. The amounts of water added were recorded for each pot to calculate the evapotranspiration. For each combination of plant and soil type, the shoot height, the number of leaves, and the chlorophyll content of leaves were determined from the same two replicates (pots) during plant growth. The shoot height in a pot was determined from four individual plants which were randomly selected from all plants, but excluding the shortest and the longest individual.
After 30 d, the plants were harvested. Shoots (stem and leaves) were cut off at 1 cm above soil and washed to remove soil particles. For one replicate from each soil type and plant species, the roots were collected and carefully washed with deionized water and sonicated for 10 min to remove attached soil particles. On pea roots, the total number of root nodules was counted. Sprouts and roots were dried at 60°C for 1 d and the dry mass determined.
Analysis of Trace Metals in Shoots and Roots
The total dried shoot and root material from each pot was ball milled to <50 µm. Subsamples of 250 mg were weighed into Teflon containers, spiked with 6 mL concentrated HNO3, 2 mL H2O, and 2 mL 35% H2O2, and digested in the microwave (12 min at 250 W, 6 min at 600 W, 13 min at 400 W). The extract was quantitatively transferred to 25 mL polypropylene Erlenmeyer flasks and made up to 25 mL using doubly deionized (Millipore Milli-Q synthesis A10) water. The extracts were analyzed for Zn, Cu, and other relevant elements by inductively coupled plasma–optical emission spectrometry (ICP–OES; Varian VISTA-MPX). If the concentration of an element exceeded the calibration range, the extract was diluted correspondingly and reanalyzed for the respective element.
Determination of Soil pH and Collection and Analysis of Soil Solution
Immediately after harvesting the plants, 4 g of moist soil from each pot were suspended in 10 mL distilled water to measure the soil pH. Three days after the harvest, the soil water content was readjusted to 255 mL kg–1 (about half the field capacity) and the soil covered with cellophane foil to minimize evaporation. After equilibration for another 4 d, 180 g of soil were transferred to a centrifuge vessel for the extraction of the soil solution and centrifuged during 30 min at 3560 x g (3500 rpm). Before soil solution extraction, the centrifuge vessels had been acid washed and rinsed with distilled water. The collected soil solution was passed through a nylon filter (0.45 µm) and immediately acidified with ultrapure concentrated HNO3 (100 µL/10 mL). The undiluted soil solution extracts were analyzed for Zn, Cu, and other relevant elements by ICP–OES. If the concentration of an element exceeded the calibration range, the extract was diluted correspondingly and reanalyzed for the respective element.
Sequential Extraction of Zinc and Copper
To investigate differences in metal speciation among the different soil batches and changes in metal speciation during the growth experiment, a sequential extraction procedure (SEP) consisting of seven extraction steps was performed (Zeien and Brümmer, 1989). The hypothetical interpretation of the operational fractions F1 to F7 is given in parentheses (Zeien and Brümmer, 1989; Schwartz et al., 1999): F1: 1 M NH4NO3, soil to solution ratio (SSR) of 25 mL per g of soil (readily soluble and exchangeably adsorbed metals); F2: 1 M NH4–acetate, pH 6.0 (specifically adsorbed, CaCO3 bound, and other weakly bound species); F3: 0.1 M NH2OH–HCl plus 1 M NH4–acetate, pH 6.0 (metals bound to Mn oxides); F4: 0.025 M NH4EDTA, pH 4.6 (metals bound to organic substances); F5: 0.2 M NH4–oxalate, pH 3.25 (metals associated with amorphous and poorly crystalline Fe oxides); F6: 0.1 M ascorbic acid in 0.2 M NH4–oxalate, pH 3.25, in boiling water (metals associated with crystalline Fe oxides); F7: Aqua regia: 9.3 M HCl, 3.85 M HNO3 (residual metal fraction).
Soil samples from the growth experiment were collected after the extraction of the soil solution. Soil from before and after the growth experiments was air-dried at 60°C, sieved to <2 mm. For each combination of plant and soil type, soil from two pots was selectively extracted in duplicates. The extracts were analyzed for Zn and Cu using ICP–OES. Calibration standards were prepared in the same background electrolyte as the extracts.
Statistical Analysis
For selected parameters, the effect of the soil treatment and the plant type were analyzed by a two-way analysis of variance (ANOVA) (Table 2). Where indicated, the parameters were log-transformed to obtain a normal distribution. Depending on whether a parameter was determined in replicates or not, we used the model for treatments with replicates (including interaction term) or without replicates (no interaction term), respectively. Differences between group means were tested for their significance by pairwise comparison with the least significant difference (LSD) at a probability level of p = 0.005. The significance level was adjusted based on the number of the most relevant plant-wise comparisons between different soil treatments (three comparisons for each plant, nine in total).
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Table 2. Probabilities for factors soil and plant and the interaction soil x plant from a two-way ANOVA of the parameters plant dry weight and evapotranspiration, contents of Zn and Cu in plant shoots and roots (log-transformed), concentration of Zn and Cu in soil solution (log-transformed), and soil pH.
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RESULTS AND DISCUSSION
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Plant Growth in the Uncontaminated and the Contaminated Soils
Selected growth parameters of barley, pea, and sunflower grown in the uncontaminated and the two contaminated soils are shown in Fig. 1. The results from the ANOVA are listed in Table 2. The results for the shoot dry weight and the evapotranspiration showed that both the soil treatment and the plant species were highly significant factors and that the plants differed in their reaction to the soil treatments. Sunflower reacted most strongly to soil contamination and all seedlings in the contaminated soils died within the first 2 wk of the growth experiment. The measured evapotranspiration of approximately 1.4 L in the contaminated soils planted with sunflower thus mainly resulted from water evaporation from the soil surface. A similar rate of water evaporation was found from a 3 d control experiment with unplanted soil. Assuming evaporation to account for 1.4 L of the total evapotranspiration, the transpiration of the pea plants in the contaminated soils was strongly reduced relative to the uncontaminated soil. This reduction is paralleled by a significant reduction in the pea shoot dry weight. Between the two contaminated soils, on the other hand, the shoot dry weight of pea did not vary significantly. The number of nodules on pea roots showed a strong decline from the uncontaminated soil (392 nodules in one pot, n = 1) to the freshly contaminated (42) and the 4 yr contaminated soil (54). This indicates that the rhizobia also suffered from Zn and/or Cu toxicity. This has also been observed in a field study on long-term sludge treated soils (Chaudri et al., 2000). The evapotranspiration of barley did not vary significantly between the three soil treatments, while its shoot dry weight and shoot height were lowest in the soil contaminated for 4 yr. During growth, barley leaves in the contaminated soils became increasingly chlorotic and exhibited an increase in reddish-brown spots indicating leaf necrosis. Those symptoms were more severe in the soil contaminated for 4 yr than in the freshly contaminated soil and likely resulted from a combination of increasing fungal infection and symptoms of Zn and/or Cu toxicity and concomitant Fe and/or Mn deficiency (Bergmann, 1993). After about 2 wk of growth in all soils, the basal parts of the barley shoots assumed a reddish color and also the leaves in the uncontaminated soil turned slightly chlorotic (yellowish), indicating a minor N deficiency.
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Fig. 1. Shoot height, shoot dry weight, and total evapotranspiration of barley, pea, and sunflower grown in uncontaminated soil, freshly contaminated soil, and 4 yr contaminated soil. Bars indicate the standard deviation from four replicates (pots). Different labels indicate means that differ significantly from each other (LSD test; p = 0.005).
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Overall, the soil contamination with filter dust resulted in plant toxicity and significantly affected the growth of all studied plants. The sunflower died in both contaminated soils. The growth of pea was similarly reduced in both contaminated soils, while barley growth was more severely affected in the soil contaminated for 4 yr. Thus, the growth parameters suggested that the adverse effects of Zn and Cu on plant growth persisted or even increased during 4 yr of soil aging.
Zinc and Copper Contents of Shoots and Roots
The total concentrations of Zn and Cu in plant shoots and roots are shown in Fig. 2 (columns representing the concentrations in shoots and roots are plotted upward and downward, respectively). The results from the ANOVA on the log-transformed contents are listed in Table 2. For the contents of Zn and Cu in the shoots, the ANOVA indicated that the soil treatment and the plant type were both highly significant in explaining the observed concentration levels and that the plants differed in their metal uptake in response to the soil treatments. The concentrations of Zn and Cu in the shoots of plants grown in the uncontaminated soil were within the typical range reported in the literature (Kabata-Pendias and Pendias, 2001). For plants grown in the contaminated soils, significantly higher Zn and Cu contents were found and the contents of Zn in the shoots were clearly higher than those of Cu. Barley contained significantly more Zn and Cu in its shoots than pea.
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Fig. 2. Zinc and Cu concentrations in the shoots and roots of barley, pea, and sunflower. Shoot contents are plotted upward, roots contents downward. Where indicated, contents in plants grown in uncontaminated soil were multiplied for better visibility. Bars for shoot contents show the standard deviation from 4 replicates (pots). Root contents were determined from one pot only. Different labels indicate means that differ significantly from each other (LSD test; p = 0.005; on log-transformed data).
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As observed for the shoots, the roots of plants grown in the two contaminated soils contained significantly more Zn and Cu than the roots of plants grown in the uncontaminated soils, and the Zn content of the roots were higher than the Cu content. Despite careful washing, some soil or filter dust associated with the roots might have contributed to the measured Zn and Cu contents. However, because such artifacts would cause random differences in the measured root contents, the systematic trends in the contents of Zn and Cu in the roots suggested that they were negligible. Because trace metals strongly bind to surface functional groups in the apoplasm (Marschner, 1995), we assumed that Zn and Cu sorbed in the apoplasm accounted for most of the total Zn and Cu measured in the root tissues. The results from the ANOVA indicated that the soil treatment significantly affected the metal content of the roots, while the plant type only had a significant influence on Zn in the roots, but not on Cu. The Cu contents of the roots of all studied plants were similar in one specific soil and about 5 to 20 times higher than the corresponding shoot contents, in agreement with the strong complexation of Cu by organic functional groups in the apoplasm (Marschner, 1995). Similarly, the contents of Zn in the roots of sunflower and especially pea grown in contaminated soils were also much higher than in the shoots. Only for barley grown in the contaminated soils were the Zn contents in the shoots similar or even higher than the contents of the roots.
The Fe contents of the shoots of barley and pea were in the range of 30 to 45 mg kg–1 and did not systematically vary with soil treatment, while the Mn contents increased from the uncontaminated to the freshly to the 4 yr contaminated soil (10, 29, and 46 mg kg–1 in barley and 19, 30, and 53 mg kg–1 in pea, respectively). The contents of both Fe and Mn were at the lower end of the concentration range in plant tissues reported in the literature (Bergmann, 1993; Kabata-Pendias and Pendias, 2001). While no reduction in the uptake of Fe and Mn by barley and pea due to Zn and Cu contamination could be observed, the Fe/Zn ratio (w/w) in the shoot tissues strongly decreased from the uncontaminated to the freshly and the 4 yr contaminated soil (0.636, 0.018, and 0.010 for barley and 0.605, 0.065, and 0.069 for pea, respectively). This indicated that the nutrition status of the plants grown in the contaminated soils was severely disrupted, causing growth reduction and toxicity symptoms similar to Fe deficiency (see previous section).
Due to the high levels of Zn and Cu in the contaminated soils, their contents in the plant shoots may not have been controlled by their availability only. Physiological factors may lead to a plateau response of metal uptake into shoots at increasing available metal contents in the soil (McBride, 1995; Hamon et al., 1999) and the metal contents in the plant tissues may be limited by the kinetics of their uptake by the plant rather than by the kinetics of their supply from the soil solid phase (Whiting et al., 2001). If the Zn and Cu contents in the plants were limited by the plant uptake, the Zn and Cu contents of plants grown in the 4 yr contaminated soil should have been similar to the contents in the freshly contaminated soil. Figure 2 shows that the Zn and Cu contents of plants grown in the 4 yr contaminated soil were similar to or even higher than the Zn and Cu contents of plants grown in the freshly contaminated soil. This trend was remarkable because the freshly contaminated soil contained about one third more Zn and Cu than the 4 yr contaminated soil (Table 1). Thus, the contents of Zn and Cu in the plant tissues showed no evidence for a reduction in the metal bioavailability during soil aging. On the contrary, they rather suggested that the availability of Zn and Cu for plant uptake slightly increased during 4 yr of soil aging.
Zinc and Copper Concentrations in the Soil Solution and Soil pH
The concentrations of Zn and Cu in the soil solution sampled by soil centrifugation after the growth experiment and the soil pH (H2O) measurements are shown in Fig. 3. The results from the ANOVA of these parameters are listed in Table 2. The soil pH was lower in the contaminated soils than in the uncontaminated soil, though the difference was not significant in the case of barley. This decrease in pH may have been caused by proton release from inorganic and organic soil components due to complexation reactions with Zn and Cu.
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Fig. 3. Zinc and Cu concentrations in soil solutions collected after plant growth experiments and corresponding soil pH (in distilled water). Where indicated, concentrations were multiplied for better visibility. Bars show the standard deviation from three replicates (pots). Different labels indicate means that differ significantly from each other (LSD test; p = 0.005; for dissolved Zn and Cu on log-transformed data).
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The concentration of Zn in the soil solutions was higher for the contaminated than the uncontaminated soil. The highest concentrations were observed in the contaminated soils planted with sunflower. Since the sunflower seedlings died within 2 wk after transplantation into the pots, the concentrations of Zn and Cu measured in the experiments with sunflower may be considered to represent the reference concentrations in the absence of significant influences from root exudates and metal uptake into roots and shoots. The concentration of Zn in the two contaminated soils planted with sunflower were the same, indicating that the solubility of Zn did not vary from the freshly contaminated to the 4 yr aged soil. Much lower Zn concentrations were observed in the soil solutions of the contaminated soils grown with barley and pea. From the difference in the Zn concentrations between the experiments with barley or pea and the experiments with sunflower, the difference in dissolved Zn at the time of soil solution extraction was roughly estimated as 4.6 mg Zn per pot for barley and 3.6 mg Zn per pot for pea (280 mL soil solution per pot). From the shoot dry weight (Fig. 1) and the Zn content in the shoots (Fig. 2) of barley and pea grown in the contaminated soils, the mass of Zn removed by uptake into shoots was estimated to be approximately 8.4 and 2.5 mg Zn per pot for barley and pea, respectively. Thus, the magnitude and trends of Zn uptake by barley and pea agreed relatively well with the Zn decrease in soil solution. This suggested that the differences in the dissolved Zn concentrations were effectively due to differences in the extent of Zn uptake by barley and pea. Between the two contaminated soils, no significant differences in the Zn concentrations in solution were observed.
In the contaminated soils, the plants appeared to have a strong influence on the concentration of Cu in solution, which was found to be much higher in the experiments with barley than in those with pea. The experiments with sunflower may again be considered as a plant-free reference due to the negligible biomass and early death of the sunflowers in the pot experiments. The mobilizing effect of the roots on Cu may be related to the exudation of metal binding substances. An increase in the uptake of Zn, Ni, and Cd from a calcareous soil by graminaceous plants was observed when grown under Fe-deficient conditions (Römheld and Awad, 2000). For wheat, it was found that Fe deficiency led to an increase in Cu acquisition, possibly due to an enhanced release of phytosiderophores (Chaignon et al., 2002). The much higher soil solution concentrations of Cu in the experiments with barley may thus have resulted from Zn- and/or Cu-induced Fe deficiency leading to the exudation of siderophores, which in turn could have promoted the mobilization of Cu either from the soil matrix or from corroding brass particles. Comparing the two contaminated soils, the Cu concentrations in the soil contaminated for 4 yr were consistently higher than in the freshly contaminated soil. This trend was in agreement with the higher Cu contents of the plant roots and shoots in the 4 yr aged soil (Fig. 2).
Chemical Extractability of Zinc and Copper and Changes with Soil Aging
The contaminated soils before and after the plant growth experiments were sequentially extracted following the procedure of Zeien and Brümmer (1989). The results for Zn and Cu are shown in Fig. 4. The left panels (A, C) show the relative amounts of Zn or Cu extracted in all fractions, while the right panels (B, D) show the fractions of Zn or Cu extracted in the steps F1 +F2 of the extraction procedure and the amount of Cu extracted in step F4.
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Fig. 4. Sequentially extracted fractions of Zn and Cu in freshly contaminated and 4 yr contaminated soil before and after the growth of plants (expressed in % of the total extracted amounts). (A) Fractions of Zn extracted in steps F1 through F7. (B) Sum of the fraction of Zn extracted in steps F1 and F2. (C) Fractions of Cu extracted in steps F1 through F7. (D) Sum of the fraction of Cu extracted in steps F1 and F2 and fraction of Cu extracted in step F4. The bars in panels (C) and (D) represent the 95% confidence limits of the means from 4 replicates (duplicate extraction of soil from two pots per treatment).
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Before the growth of plants, 81% of the Zn in the freshly contaminated soil were extracted in steps F1 +F2 of the SEP, due to the high solubility of ZnO and the short time (24 h in each extraction step) available for Zn redistribution into the soil matrix. The cumulative fraction F1 +F2 in the freshly contaminated soils after the growth experiment was already reduced to 61 to 63% due to the dissolution of ZnO and the sequestration of Zn into the soil matrix. Four years of soil aging in the lysimeters reduced the fraction of Zn extracted in the steps F1+F2 to 55%. Comparison with the freshly contaminated soil after the growth experiments suggested that most changes occurred in the first few months of soil contamination. On the other hand, the extractable fraction F1 +F2 was again reduced to 48 to 50% in the soils sampled after the pot experiments with the aged soil. This continued redistribution may have been provoked by the soil handling before the pot study. Soil sieving and drying may have led to an additional homogenization of the contamination in the soil and to the creation of finer soil aggregates allowing for some additional metal sorption during the pot experiment. In an earlier study (Voegelin et al., 2005) we found that the ZnO in the contaminated soil completely dissolved within 9 mo under field conditions and that half of the Zn formed a precipitate of the Zn-LDH type, while half of the Zn was adsorbed to soil organic matter and the soil clay fraction. In the sequential extraction, these large changes in Zn speciation during soil incubation resulted in a decrease of Zn in the fractions F1 +F2 from 81% before the pot experiments with freshly contaminated soil to 48% in the 4 yr contaminated soil after the pot experiments.
Contamination with Cu occurred in the form of brass platelets in the filter dust. These brass particles had lateral dimensions of several hundred µm and thicknesses of several tens of µm. Based on studies on brass corrosion in dilute aqueous solutions (Rao and Nair, 1998; Marshakov, 2005; Valcarce et al., 2005), corrosion can be estimated to proceed with a rate of 3.5 to 50 µm per year. Considering these rates and the geometry of the brass particles, we estimated that the oxidative dissolution of individual brass particles in soil requires a few months up to a few years under field conditions. Consequently, the release of Cu through oxidative dissolution of brass platelets was expected to proceed more slowly than the release of Zn from the dissolution of the finely powdered soluble ZnO. The large amounts of Cu extracted in steps F1 and F2 of the sequential extraction procedure were due to the accelerated oxidation of brass in the presence of ammonia (Dehri and Erbil, 1999) and the formation of Cu-NH3 complexes, which are thermodynamically much more stable than the respective complexes of other metal cations such as Zn, Cd, Pb, Fe, or Mn (Martell et al., 1997). Thus, the fractions F1 and F2 of the sequential extraction cannot simply be interpreted as exchangeable or mobilizable Cu; they also contained a considerable fraction of the brass added to the soil. Similar amounts of Cu were extracted with the fractions F1 +F2 from the freshly contaminated soil before and after the pot experiment. This suggested that the brass particles dissolved rather slowly and that no marked Cu redistribution occurred within the 30 d of the pot experiment. However, clear decrease of Cu in the fractions F1 +F2 from between 52 and 57% to between 37 and 44% was observed when comparing the freshly contaminated and the 4 yr contaminated soil. This decrease was accompanied by an increase of Cu extracted in the step F4 from between 26 and 29% to between 36 and 40%. The fraction F4 is considered to represent organically bound Cu. The decrease in F1 +F2 and concomitant increase in F4 from the freshly contaminated to the 4 yr contaminated soil therefore likely reflected the ongoing slow dissolution of brass particles and Cu readsorption onto soil organic matter. This interpretation is in agreement with the observed increase in the Cu contents of the roots of barley and pea in the 4 yr contaminated soil (Fig. 2) and the increase of the Cu concentrations in the soil solution (Fig. 3).
Effect of Changes in the Speciation of Copper and Zinc on their Availability
In the case of Cu, the results from the sequential chemical extraction (Fig. 4) and the soil solution data (Fig. 3) indicated that the oxidative dissolution caused the slow release of Cu from the corroding brass particles. This release was paralleled by a concomitant increase of the content of Cu in the roots and shoots of barley and pea from the freshly contaminated to the 4 yr aged soil (Fig. 2). Our results for Cu compared with the findings from other studies with soil contamination with slowly dissolving but unstable contaminants such as metal sulfides or smelter slag (Whiting et al., 2001; Dahmani-Muller et al., 2002; Panfili et al., 2005). In these systems, the availability of the metal is initially limited by the dissolution kinetics of the metal bearing phase. With increasing dissolution of the primary contaminant phase, however, freshly formed adsorbed and precipitated metal species will control the metal availability (Adriano, 2001; Basta et al., 2005; Hita and Torrent, 2005). Depending on the stability of the initial contaminant phase and the strength of metal binding to the soil matrix, this transformation may lead to an increase or a decrease in the metal availability. In the present study, we found that the availability of Cu for plant uptake increased during 4 yr of soil aging. This indicated that the Cu adsorbed to the soil matrix was kinetically more readily available for plant uptake than the initially added Cu concentrated in thermodynamically less stable brass particles. The increase in Cu bioavailability may continue until all brass has been dissolved. Over longer periods of time, on the other hand, ongoing slow transformation processes such as the incorporation of Cu into iron oxides (Ma and Uren, 1998) may again lead to a gradual decrease in the availability of Cu for plant uptake.
With respect to Zn, we were interested in the question of whether the dissolution of the ZnO and the sequestration of half of the released Zn into a Zn-LDH precipitate during 4 yr of soil aging would result in a marked reduction in the availability of Zn and its uptake by plants. The sequential extraction data indicated some redistribution of Zn from more easily extractable to more strongly bound forms (Fig. 4). However, the Zn contents of the shoots and roots of plants grown in the aged soil were similar or even higher than in the freshly contaminated soil, despite the lower total Zn content of the aged soil. Considering the high Zn contents of the contaminated soils (Table 1), the uptake of Zn into the plants may have been limited by the plant-specific uptake kinetics and plant Zn contents may not have reflected the chemical availability of Zn in the soil (McBride, 1995; Hamon et al., 1999; Whiting et al., 2001). On the other hand, the plant uptake of Zn may also have been controlled by the rate of Zn supply from the solid phase into solution (Zhang et al., 2004). The dissolution of the ZnO might have lead to the formation of two pools of Zn, one being more readily available and the other less readily available than the initially present ZnO grains. These pools might have been speculated to correspond with the adsorbed and precipitated Zn, respectively. In this case, the changes in Zn speciation would lead to a reduction of the pool of available Zn, but not of the rate of Zn supply to the soil solution. Considering that plant roots may acidify the rhizosphere soil, the observation that the Zn uptake in the aged soil was not reduced might also be explained by the limited stability of Zn-LDH type precipitates under acidic conditions (Voegelin et al., 2002; Voegelin et al., 2003; Voegelin and Kretzschmar, 2005). Finally, the spatial distribution of Zn at the micrometer scale may also affect its availability and uptake. Even though the filter dust was intensively mixed with the soil material, we expect that it was initially enriched around soil aggregates. The supply of Zn to the soil solution from dissolving ZnO grains may therefore be diffusion controlled. In the aged soil, on the other hand, the adsorbed and precipitated Zn was more homogeneously distributed in the clayey soil matrix (Voegelin et al., 2005), and the rate of Zn supply may be determined by the reaction kinetics of Zn desorption and precipitate dissolution. Our data did not allow determining which of the discussed factors effectively controlled the availability of Zn in the studied soil. However, the results clearly showed that the incorporation of half of the total Zn into an LDH-type precipitate and the adsorption of the other half to the soil matrix did not lead to a reduction in the plant uptake of Zn. Considering that the initially added ZnO is fairly soluble and readily available for plant uptake (Whiting et al., 2001; Voegelin et al., 2005), this suggested that the immobilizing effect of the formation of a Zn-bearing precipitate of the LDH-type was small. In the long-term, slow immobilization processes such as the incorporation of silicate into the Zn-LDH phase and its gradual transformation into a more stable Zn-phyllosilicate (Ford et al., 1999; Ford and Sparks, 2000) might cause a reduction in the bioavailability of Zn. However, we conclude from this study that the influence of the formation of weakly crystalline Zn precipitates on the bioavailability of Zn in soils may be limited and must be critically assessed in the light of other relevant factors such as the total Zn content, the plant specific uptake behavior, the distribution of Zn and the kinetics of Zn supply to the soil solution from all relevant Zn species.
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ACKNOWLEDGMENTS
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We thank Madeleine Goerg-Günthard and Jörg Luster from the Swiss Federal Institute of Forest, Snow, and Landscape Research (WSL) for providing access to soil from the Cell-to-Tree project and for information on the experimental setup and soil chemical properties. Sabina Pfister is acknowledged for help with soil sampling. We thank the anonymous reviewers for their constructive comments. This research was funded by ETH Zurich within the framework of the project no. 0-20966-02.
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ABSTRACT
INTRODUCTION
