Application of bioassays to evaluate a copper contaminated soil before and after a pilot-scale electrokinetic remediation

Environmental Pollution 157 (2009) 410–416

Contents lists available at ScienceDirect

Environmental Pollution journal homepage: www.elsevier.com/locate/envpol

Application of bioassays to evaluate a copper contaminated soil before and after a pilot-scale electrokinetic remediation Quan-Ying Wang a, b, Dong-Mei Zhou a, *, Long Cang a, b, Tian-Ran Sun a a b

State Key Laboratory of Soil and Sustainable Agriculture, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008, China Graduate School of the Chinese Academy of Sciences, Beijing 100049, China

There has been a motivation towards using biological indicators for risk assessment of contaminated soil after electrokinetic remediation.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 20 May 2008 Received in revised form 10 September 2008 Accepted 19 September 2008

Remediation programmes are considered to be complete when human risk-based criteria are met. However, these targets are often unsatisfied with the ecological parameters that may be important with regard to future soil use. Five soil subsamples, collecting along a pilot-scale soil column after electrokinetic treatment, were studied, from which about 42.0%–93.3% soil Cu had been successfully removed. A series of biological assays including soil microbial biomass carbon, basal soil respiration, soil urease activity, earthworm assays, and seed assays were used to evaluate their ecological risks. The results showed that the bioassay data from the treatment variants did not supposedly reflecting the decreased soil Cu concentrations after the electrokinetic treatment, but were highly correlated with some soil physicochemical characteristics. It suggests that bioassays are necessary to assess the ecotoxicity of soil after electrokinetic treatment. Ó 2008 Elsevier Ltd. All rights reserved.

Keywords: Contaminated soil Copper Electrokinetic remediation Bioassay

1. Introduction The environmental persistence of heavy metals associated with their intensive use by modern society has, over the years, caused metal accumulation in the biosphere (Kakkar and Jaffery, 2005), and may induce adverse effects on food quality, soil health and the environment. In response to these negative effects, there has been ongoing development of a variety of technologies to treat the soils contaminated by heavy metals (Gray et al., 2006). Electrokinetics provides a physical method for the extraction of chemicals from contaminated sites, and can be used for in situ treatment of heavy metals and organic contaminated soil (Zhou et al., 2004). Its operation principle is to apply a direct electric field in soil to drive pollutant within the soil pores towards the electrode through electromigration, electroosmosis and electrophoresis. However, the application of a direct electric current to soil results in several changes, such as pH, redox potential and electrolyte concentration, in the soil medium. These changes may impact the nature of the clay surface chemistry and the success of the electrochemical remediation (Lear et al., 2004; Al-Hamdan and Reddy, 2008).

* Corresponding author. State Key Laboratory of Soil and Sustainable Agriculture, Institute of Soil Science, Chinese Academy of Sciences, 71 East Beijing Road, Nanjing 210008, China. Tel.: þ86 25 8688 1180; fax: þ86 25 8688 1000. E-mail address: [email protected] (D.-M. Zhou). 0269-7491/$ – see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2008.09.036

An important requirement before any methods for the remediation of contaminated soil can be accepted is to demonstrate that they are effective, and that they have few adverse impacts on soil health (Lear et al., 2007). In order to meet these criteria, the contaminated soils are traditionally subjected to chemical analysis before and after treatment (Plaza et al., 2005). By lowering soil metal concentrations, treatments are assumed to reduce the potential toxicity and bioaccumulation of these metals in exposed organisms such as plants and terrestrial invertebrates. However, chemical transformation occurring during the treatment could greatly modify the bioavailability of these metals, and possibly make them more hazardous to the living organisms. Moreover, soil quality is a complex characteristic determined by physical, chemical and biological components of the soil (Pe´rez de Mora et al., 2005). Hence, chemical data alone are often insufficient to evaluate the toxic effects of the contaminants, and there is increasing interest for incorporation of toxicity tests to support management decisions for remediation (Saterbak et al., 1999; Plaza et al., 2005). The use of biological endpoints can help to appropriately define acceptable cleanup standards (Braud-Grasset et al., 1993; Debus and Hund, 1997; Dorn et al., 1998; Dorn and Salanitro, 2000), and several studies have been done to evaluate the effect of electrokinetic remediation on the soil quality. Lear et al. (2007) found that the application of electrokinetics to soil altered both the physicochemical characteristics and the exposed microbial community, and the interactions between electrokinetic processes and

Q.-Y. Wang et al. / Environmental Pollution 157 (2009) 410–416

microbial communities of soil are key to improving the efficacy and sustainability of the remediation strategy. The phytotoxicity of a Cd contaminated soil after electrokinetic remediation was also studied, and an increase in phytotoxicity was found (Giannis et al., 2008). However, until now, a comprehensive and in-depth study of individual tests with various organisms and endpoints is scarce to evaluate the soil quality after electrokinetic remediation. This study aimed to apply a series of bioassays to evaluate the ecotoxicity of a Cu contaminated soil after electrokinetic remediation, to compare the suitability and sensitivity of individual tests with various organisms and endpoints for the ecotoxicity of soil after electrokinetics, and to develop soil assessment methods based on bioassays and chemical data. 2. Materials and methods 2.1. Experimental setup The tested soil (Udic Ferrosols, 20–60 cm) was sampled from Yingtan County, Jiangxi Province. Its pH (1:2.5 soil to water), electrical conductivity (EC, 1:2.5 soil to water), organic matter content (OM, by K2CrO4), and cation exchange capacity (CEC, by ammonium acetate exchange) are 4.8, 0.28  102 mS cm1, 4.3 g kg1, and 12.2 cmol (þ) kg1, respectively. The soil Cu, Cr, Pb, Zn and Cd concentrations are 19.3, 95.3, 57.0, 31.3 and 0.04 mg kg1, respectively. About 700 kg of soil was weighted and spiked with 18 g CuSO4  5H2O per 5 kg soil in solution form, and incubated at room temperature for 2 weeks. The pilot-scale experiment was conducted in a rectangle soil box. A 5 cm thick layer of cobblestones was spread on the bottom of the box, and 10 cm opening was placed in each side of the box to hold electrodes, which were surrounded by small cobblestones. The anode and cathode consisted of four cylindrical stainless steel meshes. The Cu-spiked soil was loaded in the central soil box with a height of 70 cm. A 60-mesh nylon cloth was placed between electrode chamber and soil column to prevent migration and/or dispersion of soil particles from the cell to the electrode chambers. Three rows of four stainless steel passive electrodes were inserted in the soil box to monitor voltage drop distribution across the soil. After the water-soluble Cu was removed, 0.05 mol L1 lactic acid solution (pH 3.5, adjusted with 0.1 mol L1 NaOH) was cycled in the cathode chamber. A constant voltage drop of 80 V was applied across the soil column by a DC power source. The more detail description of the electrokinetic remediation process can be found in our earlier paper (Zhou et al., 2006). After the experiment was run for 140 days, five soil subsamples were sampled from the pilot-scale soil column in different positions, and each subsample was divided into three depth layers of 0–30, 30–50, and 50–70 cm. Because most soil animals, plants, and microorganism are living in the 0–30 cm layer, five parts (divided by the distance from anode to cathode, 0–20 cm (S1), 20–40 cm (S2), 40–60 cm (S3), 60–80 cm (S4), 80–100 cm (S5)) of this layer were used in this study.

411

2.5. Seed germination/root elongation test Dicotyledonae (mustard) and Monocotyledonae (ryegrass and wheat) were used for seed germination assays. Ten mustard, ten ryegrass, and five wheat seeds were sown separately in glass jars containing 20 g of soil in triplicate. Lids were loosely screwed on to reduce evaporation but allow aeration, and the seeds were left to  germinate at 22 C, 80% humidity, 16 h full illuminance and 8 h darkness. When 70% seeds in the clean soil germinated, the number of seeds germinated in all soil samples was recorded. At the end of incubation, the root length was also measured. 2.6. Earthworm assays Avoidance test: 400 g moist soil was used for each. The clean soil (200 g wet weight) was placed in one side of the test vessel and the same amount of test soil was placed in the opposite side. After this, the wall was gently removed and ten adult worms were left on the contact line of the soils. The box was covered with a lid (containing small holes) and the enchytraeids were kept at 20  C with a photoperiod of 16:8 h (light:dark) for 48 h. Five replicates per treatment were used. At the end of each exposure period, each test container in both phases was divided again by the introduction of the card divider. Soils from each side were emptied onto separate trays, and the numbers of worms in both control and contaminated soils were determined. Earthworm acute test: 500 g moist soil was used for each. Ten adult worms were placed in 1 l glass containers filled with test soils. To prevent the worms from escaping, test containers were covered with a polythene sheet with integrated gauze (F 1 mm) to ensure optimal ventilation. After 14 days of incubation, surviving worms were collected by hand. The test endpoint was mortality. 2.7. Basal soil respiration Basal soil respiration (BSR) at field capacity (CO2 production at 22  C without addition of glucose) was measured by alkali (Ba(OH)2  8H2O þ BaCl2) absorption of the CO2 released during the 24 h incubation period, followed by titration of the residual OH ions with standardized hydrochloric acid after adding three drops of phenolphthalein as an indicator, as reported by Isermayer (1952). Three replicates of each sample were tested. Data are expressed as mg CO2 g1 dry soil. 2.8. Data analyses All statistical analyses were carried out with the program SPSS 11.5 for Windows. Analysis of variance (ANOVA) and Student’s t-tests were performed to determine significant differences between the toxicity of clean soil, untreated soil and treated soil subsamples (S1–S5). Statistical significance is accepted when the probability of the result assuming the null hypothesis (p) is less than 0.05. The avoidance endpoint was expressed as the percentage of worms that avoided the treated soil in the test container from the total number of worms in that container. The results are the mean percentages of net responses (NR) calculated as follows: NR ¼ ½ðC  TÞ=N  100

2.2. Determination of soil physicochemical properties The soil subsamples were ground to pass through a 100-mesh screen (F 0.149 mm), and then digested with HF–HNO3–HClO4 for determination of soil Cu concentration by a Hitachi 180-80 Atomic Absorbance Spectrometer (AAS). Soil pH and EC were measured by a pH meter and an EC meter, respectively, with a ratio of 1:2.5 soil to water. Bioavailable soil Cu extracted with 0.01 mol L1 CaCl2 in a 1:2.5 (soil:extractant) ratio was measured by AAS.

where C ¼ number of worms observed in the control soil; T ¼ number of worms observed in test soil; N ¼ total number of worms per replicate. A positive (þ) NR indicates avoidance and a negative () NR indicates a non-response (or attraction) to the chemical or soil tested. A germination index (GI) was calculated by accounting for the number of grown seeds and the average sum of seeds’ root elongation in a sample as related to the control (Zucconi et al., 1981), following the formula below: GI ¼ ½ð number of grown seeds in sample= number of grown seeds in controlÞ

2.3. Soil microbial biomass carbon Soil microbial biomass carbon (SMBC) was determined by the fumigation extraction method (Vance et al., 1987). The soil subsamples (equivalent to 5.0 g dry soil) were extracted with 20 ml of 0.5 mol L1 K2SO4. The samples were shaken for  30 min, filtered, and frozen at 20 C. Simultaneously, the same soil subsamples (also  equivalent to 5.0 g dry soil) were fumigated in a dark room at 25 C with ethanolfree chloroform for 24 h and then extracted. The SMBC is calculated according to the equation: SMBC ¼ EC/KEC, where EC is the difference between extractable C from fumigated and non-fumigated samples and KEC ¼ 0.38. 2.4. Soil urease activity Soil urease activity was measured following a method developed by Guan (1986). This method involves the incubation of the soils with urea and buffer for 24 h  at 37 C. The resulting suspensions were filtered. Then, phenol and sodium hypochlorite were added. For calibration, seven standards were prepared (0.5, 1, 3, 5, 7, 10, and 13 mg NH4–N mL1). The blue-colored complexes were measured within 1 h using spectrophotometer as for urease activity at l ¼ 578 nm after a 30-min color development period.

 ðaverage sum of root lenghts in sample= average sum of root lengths in controlÞ  100

3. Results and discussion 3.1. Changes in soil physicochemical properties Electrokinetic remediation had a significant effect on the soil physicochemical properties. After the experiment was run for 140 days, about 93.3% (S1), 89.2% (S2), 82.1% (S3), 47.0% (S4), and 42.0% (S5) of soil Cu was successfully removed from the five sections, respectively, which demonstrated the success of Cu removal from contaminated soil under pilot-scale conditions. During the electrokinetic treatment, electrode reactions also changed soil pH and EC in the soil column. By the end of the study (140 days), a pH profile developed across the soil column from pH

412

Q.-Y. Wang et al. / Environmental Pollution 157 (2009) 410–416

2.97 (S1) close to the anode to pH 4.47 (S5) close to the cathode, which indicated an acidic soil environment, and the soil EC values varied from 0.51  102 (S3) to 6.97  102 mS cm1 (S1) (Table 1). The negative correlation (Table 2) between soil pH and EC shows that the EC gradually increased along with the decreased pH. Comparing the results, it appears that most of the soil Cu was accumulated in the sections close to the cathode after the electrokinetic treatment. Although no significant correlation was found between soil Cu concentration and pH (Table 2), some authors (Giannis et al., 2008) concluded that the main reason for this was the different pH in the soil sections. Meanwhile, this phenomenon also suggests the electromigration of cupric ion from the anode to the cathode (Hansen et al., 2007). In this study, bioavailable soil Cu concentrations (expressed by CaCl2-extractable soil Cu) remained higher in the S1–S5 than that in the clean soil (Table 1). Although the adverse effects of heavy metals on soil biological properties are well documented (McGrath et al., 1995; Giller et al., 1998), such studies have focused mainly on the total soil heavy metal concentrations. Soil functional quality, however, is more dependent on bioavailable soil heavy metal concentration since this fraction plays the most important role in the related parameters of soil quality such as enzyme activities or microbial related processes (Hattori, 1992; Speir et al., 1995). In this study, we focused on the CaCl2-extractable soil Cu fraction, because this is the most available fraction to soil microorganisms and plant roots, and leads to a better understanding of metal toxicity (Adriano, 2001). Soil quality can be categorized where: metal concentration above a level is deemed toxic and requires more careful assessment; below a certain level is considered non-toxic and may be reduced in priority unless there are special considerations. The soil Cu concentrations in S1–S3 after the treatment had decreased to less than 200 mg kg1, which were below the Level III (400 mg kg1) of the China Soil Environmental Quality Standard GB 156181995, and it means that these soils could be considered non-toxic.

3.2. Soil microbial biomass carbon The observed SMBC for different soil samples in this study are shown in Fig. 1. The lowest SMBC was measured in the untreated sample. For S1–S5, the SMBC ranged from 11.7 to 31.1 mg Cmic kg1 soil. A significant decline in SMBC was observed in all the soil subsamples compared to the clean soil (except S4) (Fig. 1). S4 gave the highest SMBC, suggesting that this sample was free from ‘harm’ and functioned normally although it still contained low concentration of Cu (Table 1). A significant difference was found between S1 and the untreated sample, and the same phenomena were also found for S2–S4. No significant difference was found between S5 and the untreated sample, and this was probably due to the high soil Cu concentration in S5 (481 mg kg1). SMBC has already been found to be sensitive to the changes of soil heavy metal concentrations (Giller et al., 1998). Of the soil physicochemical properties (total Cu concentration, CaCl2–Cu, EC) analyzed, none of them showed significant correlation with the

Table 1 Some physicochemical characteristics of the soil samples Soil samples

Total-Cu (mg kg1) CaCl2–Cu (mg kg1) pH

EC (102 mS cm1)

Clean soil S1 S2 S3 S4 S5 Untreated soil

29.3 55.4 89.4 148 439 481 829

2.80 6.97 2.30 0.51 1.76 2.07 2.65

0.00 9.70 13.0 10.5 12.6 13.5 295

4.81 2.97 4.04 4.47 4.15 4.08 4.60

SMBC data. However, SMBC was negatively related to soil pH (Table 2) in this study. It is known that the electrokinetics alter soil pH, which is a crucial parameter for soil microbial growth, influencing a complex range of interacting factors, such as membrane integrity and function, and also the bioavailability of nutrients and contaminants (Lear et al., 2004). Hence, the change of the SMBC may have been enhanced by the combination of low pH and soil Cu toxicity that interacted synergistically to make conditions even more harmful, and the direct effect of the applied electric field on SMBC could not be ascertained. 3.3. Soil urease activity Usually, enzyme activities decreased significantly with increasing level of soil pollution: non-contaminated soil showed the highest enzyme activity, contaminated soil the lowest, and restored soil intermediate values. Moreover, it is presumed that heavy metals, in most cases, inhibit enzymatic reactions because they bind with substrates, creating substrate complexes, or block reactive functional groups of enzymes, or react with enzymesubstrate complex. According to the urease assay (Fig. 2), however, the clean soil had the highest amount of urease activity (0.24 mg NH4–N g1 soil d1) and S1 had the least (0.14 mg NH4–N g1 soil d1), while the Cu concentration of S1 was much less than that of the other soil subsamples. The statistical analyses showed significant difference in soil enzyme activity between the clean soil and S1, and the same phenomena were found for S2, S3, S4, and S5. Significant difference was also found between the untreated soil and S2, or S3, or S4. Although soil urease activities in S2, S3 and S4 did not reach the level of clean soil, they were much higher than the untreated soil. However, no significant difference was found between the untreated soil and S1 or S5. The sources of enzyme in soil are still obscure but it is presumed to be affected by soil basic physicochemical properties, the pollutant content and so on. Soil enzyme activities in this study did not correlate significantly but positively with total soil Cu concentration or bioavailable soil Cu concentration. Furthermore, the activities did not seem to depend on the variations of the EC. From Table 2, we know that the field soil pH would not directly affect the enzyme assay as it is run using buffers at their optimal pH. However, some reports have found that the long-term effect of low pH would probably cause shifts in microbial community composition and size (Killam, 1994). Taking into account the trend of correlations between the variables, it is not explicit if the reduction of enzyme tested is due to pH or to bioavailable soil Cu or a combination of both. Moreover, to establish the relative effects of pH and bioavailable soil Cu concentrations on enzyme activity in these samples is difficult since both factors have been shown to promote a reduction in enzyme activities (Collin and Stotzky, 1992; Speir et al., 1999). Knowledge of the relative effects of electrokinetic treatment on soil urease activity is important because this enzyme is a key component in N cycle in soils and its substrate, urea, is incorporated into soil as fertilizer, animal excreta, or a breakdown product of nucleic acids. Although it is clear that N cycle functioning was modified in the soils with electrokinetic treatment, restored soils still showed significantly lower enzyme activities than the clean soil, indicating that these soils have not been fully restored and that the urease is still affected by the electrokinetic treatment such as S1. 3.4. Seed germination/root elongation test During the tests, the viability of the seed species was assessed by running controls each time when they were tested. Table 3 showed the results obtained by exposing the three seed species to soil of the

Table 2 Correlation coefficients between total soil Cu, CaCl2-extractable soil Cu, soil pH, soil EC, seed germination rate, root length, germination index of the three types of seeds (mustard, ryegrass, and wheat), earthworm avoidance response and mortality, and microbial properties of the clean soil, untreated and treated soils T–Cu

CaCl2–Cu

pH

EC

Urease

MSG

MRL

MGI

RSG

RRL

RGI

WSG

WRL

WGI

AR

EM

BSR

1.000 0.804* 0.290 0.256 0.269 0.878** 0.454 0.512 0.786* 0.739 0.737 0.815* 0.776* 0.811* 0.452 0.493 0.316 0.075

1.000 0.307 0.023 0.279 0.943** 0.801* 0.784* 0.999** 0.641 0.640 0.993** 0.990** 0.990** 0.322 0.600 0.537 0.453

1.000 0.773* 0.681 0.319 0.159 0.172 0.333 0.331 0.325 0.349 0.244 0.256 0.504 0.416 0.365 0.775*

1.000 0.416 0.105 0.472 0.436 0.032 0.132 0.120 0.078 0.013 0.051 0.099 0.546 0.395 0.294

1.000 0.167 0.395 0.371 0.250 0.800* 0.802* 0.278 0.343 0.372 0.873** 0.781* 0.469 0.599

1.000 0.754 0.796* 0.935* 0.665 0.662 0.920** 0.938** 0.929** 0.328 0.557 0.436 0.386

1.000 0.984** 0.800* 0.599 0.593 0.752 0.835* 0.792* 0.381 0.809* 0.799* 0.304

1.000 0.777* 0.645 0.638 0.724 0.821* 0.775* 0.415 0.808* 0.762* 0.242

1.000 0.610 0.610 0.994** 0.988** 0.987** 0.294 0.584 0.530 0.491

1.000 1.000** 0.628 0.692 0.715 0.894** 0.847* 0.592 0.351

1.000 0.629 0.693 0.716 0.899** 0.844* 0.593 0.348

1.000 0.979** 0.988** 0.321 0.578 0.488 0.457

1.000 0.995** 0.414 0.650 0.631 0.418

1.000 0.440 0.648 0.590 0.382

1.000 0.805* 0.606 0.534

1.000 0.708 0.245

1.000 0.084

T–Cu, total soil Cu concentration; CaCl2–Cu, CaCl2-extractable soil Cu concentration; pH, pH in soil water suspensions (1:2.5 (w:v) soil:solution ratio); EC, electrical conductivity; MSG, mustard seed germination; MRL, mustard root length; MGI, mustard germination index; RSG, ryegrass seed germination; RRL, ryegrass root length; RGI, ryegrass seed germination; WSG, wheat seed germination; WRL, wheat root length; WGI, wheat germination index; AR, earthworm avoidance response; EM, earthworm mortality; BSR, basal soil respiration; MBC, microbial biomass carbon. *Correlation is significant at the 0.05; **correlation is significant at the 0.01.

Microbial Carbon (mg kg-1soil) 35

30

25

20

15

10

5

0

e

untreated soil

a

clean soil

a

cd

0-20 c

20-40

ce

b

40-60

cd

40-60

Distance from Anode (cm)

b

20-40

a

60-80

d

60-80

de 413

80-100

be

80-100

Fig. 1. SMBCs of the clean soil, S1–S5 soil subsamples, and the untreated soil. Values are mean (n ¼ 3)  SE. Significance is shown at p < 0.05.

be

0-20

Distance from Anode (cm)

treatment sections 0–20, 20–40, 40–60, 60–80, and 80–100 cm distance from the anode for the test, and the germination indexes (GIs) were also calculated. No ryegrass seeds germinated in the untreated samples, while only a small number of mustard and wheat seeds germinated. This is probably due to the presence of high concentrations of Cu in the untreated soil (>800 mg kg1). After the electrokinetic treatment, all seeds in general had very high germination rates. While the effects of soil samples after treatment on germination of the three seed species were statistically significant compared with the untreated soil, there was no significant difference between the treatments and the clean soil (except S4 for mustard) (Table 3). No significant germination or root growth inhibition was observed on the wheat seed with any restored soil, as compared to the clean soil (Table 3), which meant that the wheat seed was unsuitable to evaluate the phytotoxicity of the soils in this study, because they were unable to discriminate between the soils. In the S3 and S5, the germination indexes of mustard were higher than the clean soil, which maybe due to stimulation occurred in the mustard seeds. Although the root length and GIs for ryegrass were significantly different from the clean soil, the values were much higher than those in the untreated soil, which meant that the soil qualities have been improved.

0.25

0.20

0.15

0.10

0.05

0.00 untreated clean soil soil

Fig. 2. Urease activity of the clean soil, S1–S5 soil subsamples, and the untreated soil. Values are mean (n ¼ 3)  SE. Significance is shown at p < 0.05.

Urease (mg NH4-N g-1 soil d-1)

Q.-Y. Wang et al. / Environmental Pollution 157 (2009) 410–416

T–Cu CaCl2–Cu pH EC Urease MSG MRL MGI RSG RRL RGI WSG WRL WGI AR EM BSR MBC

414

Q.-Y. Wang et al. / Environmental Pollution 157 (2009) 410–416

Table 3 Seed germination, average root length in cm, and germination index (GI) of the three types of seeds Root length (cm)

GI

90.0  10.0a 90.0  10.0a 86.7  5.80a 90.0  0.00a 66.7  5.80b 86.7  11.5a 25.0  7.07c

0.69  0.02a 0.29  0.05bc 0.64  0.19ab 0.71  0.47a 0.60  0.18ab 0.88  0.14a 0.00  0.00c

100  8.87a 42.4  10.9bc 88.2  19.9ab 104  68.3a 66.0  23.7ab 122  17.7a 0.00  0.00c

Ryegrass Clean soil S1 S2 S3 S4 S5 Untreated soil

100  0.00a 100  0.00a 100  0.00a 96.7  5.80a 100  0.00a 100  0.00a 0.00  0.00b

1.72  0.15a 0.52  0.16b 0.80  0.19cd 0.86  0.21d 0.59  0.13bc 0.56  0.02bc 0.00  0.00e

99.8  8.90a 30.4  9.13b 46.3  10.9bc 48.3  13.1c 34.5  7.27bc 32.0  1.21bc 0.00  0.00d

Wheat Clean soil S1 S2 S3 S4 S5 Untreated soil

100  0.00a 100  0.00a 100  0.00a 93.3  11.5a 100  0.00a 93.3  11.5a 30.0  10.0b

1.39  0.07a 1.20  0.10a 1.18  0.06a 1.18  0.63a 1.12  0.49a 1.31  0.43a 0.00  0.00b

99.8  5.05a 86.3  0.19a 84.9  4.32a 80.6  48.7a 86.8  35.3a 85.4  15.9a 0.00  0.00b

100

Avoidance Response (%)

Germination (%) Mustard Clean soil S1 S2 S3 S4 S5 Untreated soil

120

a a

a

a

a

60-80

80-100

a

80 60 40

Avoidance

20 0 -20

Non-Response

-40 -60 untreated sample

0-20

20-40

40-60

Distance from Anode (cm)

Significance is shown at p < 0.05.

The phytotoxicity of heavy metals depends on many factors such as soil characteristics and the soil metal bioavailability. All three tests with seed germination showed a significant negative correlation with the total- and CaCl2-extractable soil Cu concentrations. The root length and GI for mustard were negatively correlated with CaCl2-extractable soil Cu, while these values for ryegrass had no significant correlation with neither total- nor CaCl2-extractable soil Cu. The root length and GI for wheat were significantly and negatively correlated with the total- and CaCl2-extractable soil Cu. It is interesting to note from the phytotoxicity tests that the S1 after electrokinetic treatment is inappropriate for cultivation. This result was contrary to the data in Table 1, which suggests that the highest removal of soil Cu was achieved for S1. This is probably due to the inhibition of low pH at the soil sections after the treatment (An, 2004; Giannis et al., 2008). Seed germination was insensitive to Cu toxicity at most levels and the germination assay of the three seeds was unsuitable to assess soil toxicity after electrokinetic treatment, and germination was also reported to be less sensitive toxicity endpoints for contaminants such as methyl tetra-butyl ether (An et al., 2002), and crude oils (Dorn et al., 1998; Dorn and Salanitro, 2000), and cadmium (An, 2004). We know from Table 3 that mustard and ryegrass seeds were the sensitive ones to be used as indicating seed. Also, the root length and GI index were sensitive and robust enough to be used for the soil quality index. As a result, the phytotoxicity of the soils after the treatment was significantly reduced and the toxicity potential was less than the untreated soil. 3.5. Earthworm assays of soils In this study, both earthworm lethal and avoidance response were used to evaluate the risks in clean soil, S1–S5, and the untreated soils. All the earthworms were found alive in the soil at the end of the avoidance test. Avoidance behavior was observed in all the tests with S1–S5 and the untreated soil (Fig. 3). According to the 80% avoidance criterion for the habitat function accepted as an impact on behavior according to the ISO/17215-1 Guideline for the Earthworm Avoidance Test (ISO, 2006), using clean soil as the

Fig. 3. The avoidance tests of earthworms in the clean soil, S1–S5 soil subsamples, and the untreated soil. Data are mean (n ¼ 5)  SE. Dashed line indicates the trigger value of 80% of effect on the individuals. Significance is shown at p < 0.05.

reference one, an avoidance reaction 80% was reached for the soil subsamples S1–S5. No significant substrate avoidance occurred in all the treated soil samples when compared to the untreated one (Fig. 3), which indicated that a toxic effect, as a substrate, was assessed to be toxic at substrate preference <20%. As a rule, the increasing avoidance response from reference soil toward contaminated soils was regarded to be correlated with the increasing soil Cu concentrations. Loureiro et al. (2005) found that when exposed to copper sulphate the earthworm showed avoidance behavior at 320 mg kg1 of soil. Although only S4 and S5 exceeded 320 mg kg1, all the soil subsamples showed significant avoidance response. It is known that a low soil pH influences soil toxicity towards oligochaete species, in some cases overcoming the effect of the toxicant itself (Amorim et al., 1999). In this case, although no significant correlation was found between avoidance response and total- and CaCl2-extractable soil Cu, pH, or EC, the avoidance response towards the electrokinetic soil remediation also could be attributed to this factor. The similar results were also obtained with enchytraeids by Ro¨mbke et al. (2002), who showed that species distribution in the field was not affected by increasing Zn exposure, but could be masked by the pH values associated with the increasing amounts of soil Zn. The results in our study meant that the soils after treatment presented a limited habitat quality. Results of the 14 day earthworm acute toxicity test showed that only in the clean soil did all the worms survive at the end of the test. The untreated soil, containing the highest Cu concentration, was significantly toxic. The soil after remediation was toxic to earthworms, and mortality rates of S2–S5 ranged from 27% to 53% (Fig. 4). Although the mortalities of S2–S5 showed statistically significant difference from the clean soil, they were much lower than that of the untreated soil. It means that the remediation has reduced Cu toxicity to earthworms. Moreover, although S1 had the least soil Cu concentration, earthworms in S1 died completely. This means that the S1 was unsuitable for further reuse. No significant correlation was found between the mortality and total-Cu, CaCl2– Cu, soil pH, or soil EC. Therefore, soil Cu concentration after remediation is unlikely to account for their toxicity to earthworms. Usually, the acute toxicity test is useful for screening highly contaminated soils, and the avoidance test can provide essential data when sublethal effects of low level exposures and long-term ecological risks of soil contamination. The phenomenon that avoidance response was significantly correlated to mortality

Q.-Y. Wang et al. / Environmental Pollution 157 (2009) 410–416

b

b

100

Mortality (%)

80 d 60

cd

cd c

40 20 a

0

untreanted clean soil soil

0-20

20-40

40-60

60-80

80-100

Distance from Anode (cm) Fig. 4. Earthworm mortality (% of dead animals related to the numbers of individuals at test start) in the clean soil, S1–S5 soil subsamples, and the untreated soil after 14 days of incubation. Significance is shown at p < 0.05.

(Table 2), demonstrates that both earthworm assays were sensitive and representative biological indicators that can be used as screening tools for assessments after electrokinetic treatment in this study. 3.6. Basal soil respiration Obvious changes of values calculated from BSR data of samples were found between the clean soil and the untreated soils, as shown in Fig. 5. The clean soil had significantly higher (p < 0.05) BSR than that of S1, S2, and S3, and this was probably due to the low pH values of the treated soils. BSR had no significant correlation with total-Cu, CaCl2-extractable soil Cu, soil pH, or soil EC (Table 2), suggesting that there were other inhibitory constraints affecting the microbial processes. 3.7. Indicators of soil quality Indicators of soil quality and health, to be useful and practical, must meet certain criteria. These include: sensitivity to perturbation or contamination; a relationship to soil function; reproducibility and low temporal and spatial variability; and have simple sampling and analytical methods (Doran and Parkin, 1996; Doran and Zeiss, 2000; Nortcliff, 2002). Thus, measurements of the status

a

ab

Basal Soil Respiration (mg CO2 g-1 soil d-1)

0.05

a

bc

0.04

4. Conclusions To our knowledge, this is the first in-depth study aimed at using a series of bioassays to evaluate the environmental risks of contaminated soil after electrokinetic remediation using lactic acid as enhancing reagent for adjusting the catholyte pH. Biological assays using soil microbial biomass carbon, soil basal respiration, urease activity, earthworm (E. fetida) avoidance response and acute toxicity, and root elongation of mustard and ryegrass were shown to be good indicators of soil quality after the electrokinetic remediation. Soil remediation cannot be simply evaluated by the removal efficiency of soil pollutants. The change of soil physicochemical and biological properties should also be included. Complete detoxification of contaminated soil may need gentle remediation technology, and should be evaluated based on environmental risks.

This work is supported by the National Natural Science Foundation of China (No. 40671095) and the Knowledge Innovative Program of Chinese Academy of Sciences (No. KXCX3-SW-435).

c 0.02

and activity of specific organism contributing to soil processes have the potential to provide rapid and sensitive means characterizing soil quality. The selected test organisms for the tests in this study represented different trophic levels (producers, consumers, decomposers). Microbial assays including SMBC, BSR, enzyme activity, are commonly used as indicators of soil heavy metal contamination. The SMBC represents the size of the entire microbial community (Lorenz and Kandeler, 2005; Megharaj et al., 2000), BSR provides a measure of the decomposition of organic C within soils by those capable and active microorganisms (Torstensson, 1997), while soil enzyme activities are commonly used as indicators of soil health because they play an important role in nutrient cycling (Gianfreda et al., 2005). When considering higher organisms, earthworms are widely used because they are representative soil fauna. Plant bioassays in assessing soil pollution also merit considerable attention (Adam and Duncan, 2002). Result of soil urease activity was significantly correlated with other biological indicators, especially the results of earthworm tests, and ryegrass root elongation (Table 2), showing that soil urease activity was a valid indicator for ecological assessments of soil toxicity after electrokinetic remediation. The similar phenomena were also found for earthworm assays and BSR. Few studies have linked measured electrokinetic treatment results to ecotoxicity data. The importance of ecotoxicity tests in the complement of the chemical assessment of electrokinetic remediation is obvious from the results of this study. Using the data derived from the biological indicator assays, the soils could be ranked in order of soil health as the clean soil > S1–S5 > the untreated soil. The effective bioassays that the mustard and ryegrass root elongation test, the earthworm avoidance and acute tests, the urease assay, the MBC assay, and the BSR assay have potential to meet all the criteria mentioned above, and are appropriate for assessing the environmental risk of contaminated soil after electrokinetic treatment.

Acknowledgements

c

0.03

415

c

0.01

References

0.00 untreated clean soil soil

0-20

20-40

40-60

60-80

80-100

Distance from Anode (com) Fig. 5. BSR of the clean soil, S1–S5 soil subsamples, and the untreated soil. Values are mean (n ¼ 3)  SE. Significance is shown at p < 0.05.

Adam, G., Duncan, H., 2002. Influence of diesel fuel on seed germination. Environ. Pollut. 120, 363–370. Adriano, D.C., 2001. Trace Elements in Terrestrial Environments-Biogeochemistry, Bioavailability, and Risks of Metals, second ed. Springer-Verlag, New York. Al-Hamdan, A.Z., Reddy, K.R., 2008. Transient behavior of heavy metals in soils during electrokinetic remediation. Chemosphere 71, 860–871.

416

Q.-Y. Wang et al. / Environmental Pollution 157 (2009) 410–416

Amorim, M.J., Sousa, J.P., Nogueira, A.J.A., Soares, A.M.V.M., 1999. Comparison of chronic toxicity of Lindane (g-HCH) to Enchytraeus albidus in two soil types: the influence of soil pH. Pedobiologia 43, 635–640. An, Y.J., 2004. Soil ecotoxicity assessment using cadmium sensitive plants. Environ. Pollut. 127, 21–26. An, Y.J., Kampbell, D.H., McGill, M.E., 2002. Toxicity of methyl tert-butyl ether (MTBE) on plants (Avena sativa, Zea mays, Triticum aestivum, and Lactuca sativa). Environ. Toxicol. Chem. 21, 1679–1682. Braud-Grasset, F., Baud-Grasset, S., Safferman, S.I., 1993. Evaluation of the bioremediation of a contaminated soil with phytotoxicity tests. Chemosphere 26, 1365–1374. Collin, Y., Stotzky, G., 1992. Heavy metals alter the electrokinetic properties of bacteria, yeasts, and clay minerals. Appl. Environ. Microb. 58, 1592–1600. Debus, R., Hund, K., 1997. Development of analytical methods for the assessment of ecotoxicological relevant soil contamination. Part B-Ecotoxicological analysis in soil and soil extracts. Chemosphere 35, 238–261. Doran, J.W., Parkin, T.B., 1996. Quantitative indicators of soil quality: a minimum data set. In: Doran, J.W., Jones, A.J. (Eds.), Methods for Assessing Soil Quality Soil Science Society of America. Special Publication 49, Madison, WI, pp. 25–37. Doran, J.W., Zeiss, M.R., 2000. Soil health and sustainability: managing the biotic component of soil quality. Appl. Soil Ecol. 15, 3–11. Dorn, P.B., Salanitro, J.P., 2000. Temporal ecological assessment of oil contaminated soils before and after bioremediation. Chemosphere 40, 419–426. Dorn, P.B., Vipond, T.E., Salanitro, J.P., Wisniewski, H.L., 1998. Assessment of the acute toxicity of crude oils in soils using earthworms, Microtox, and plants. Chemosphere 37, 845–860. Gianfreda, L., Rao, M.A., Piotrowska, A., Palumbo, G., Colombo, C., 2005. Soil enzyme activities as affected by anthropogenic alterations: intensive agricultural practices and organic pollution. Sci. Total Environ. 341, 265–279. Giannis, A., Gidarakos, E., Skouta, A., 2008. Transport of cadmium and assessment of phytotoxicity after electrokinetic remediation. J. Environ. Manage. 86, 535–544. Giller, K.E., Witter, E., McGrath, S.P., 1998. Toxicity of heavy metals to microorganisms and microbial processes in agricultural soils: a review. Soil Biol. Biochem. 30, 1389–1414. Gray, C.W., Dunham, S.J., Dennis, P.G., Zhao, F.J., McGrath, S.P., 2006. Field evaluation of in situ remediation of a heavy metal contaminated soil using lime and redmud. Environ. Pollut. 142, 530–539. Guan, S.Y., 1986. Soil Enzymes and Research Methods. China Agricultural Science Press, Beijing (in Chinese). Hansen, H.K., Rojo, A., Ottosen, L.M., 2007. Electrokinetic remediation of copper mine tailings – implementing bipolar electrodes. Electrochim. Acta 52, 3355– 3359. Hattori, H., 1992. Influence of heavy metals on soil microbial activities. Soil Sci. Plant Nutr. 38, 93–100. Isermayer, H., 1952. Eine einpache Methode zur Bestimmung der Pflanzenatmung und der Karbonate in Boden, Z. Pflanzenerna¨hr. Du¨ng. Bodenk 56, 26–28. ISO (International Organization for Standardization), 2006. Soil Quality d Avoidance Test for Testing the Quality of Soils and Effects of Chemicals on Behaviour d Part 1: Test with Earthworms (Eisenia fetida and Eisenia andrei). ISO. 17512–1. Kakkar, P., Jaffery, F.N., 2005. Biological markers for metal toxicity. Environ. Toxicol. Pharm. 19, 335–349.

Killam, K., 1994. Soil Ecology. Cambridge University Press, New York, pp. 24-28. Lear, G., Harbottle, M.J., van der Gast, C.J., Jackman, S.A., Knowles, C.J., Sills, G., Thompsona, I.P., 2004. The effect of electrokinetics on soil microbial communities. Soil Biol. Biochem. 36, 1751–1760. Lear, G., Harbottle, M.J., Sills, G., Knowles, C.J., Semple, K.T., Thompson, I.P., 2007. Impact of electrokinetic remediation on microbial communities within PCP contaminated soil. Environ. Pollut. 146, 139–146. Lorenz, K., Kandeler, E., 2005. Biochemical characterisation of urban soil profiles from Stuttgart. Germany. Soil Biol. Biochem. 37, 1373–1385. Loureiro, S., Soares, A.M.V.M., Nogueira, A.J.A., 2005. Terrestrial avoidance behaviour tests as screening tool to assess soil contamination. Environ. Pollut. 138, 121–131. McGrath, S.P., Chaudri, A.M., Giller, K.E., 1995. Long-term effects of metal in sewage sludge on soils, microorganisms and plants. J. Ind. Microbiol. 14, 94–104. Megharaj, M., Singleton, I., McClure, N.C., Naidu1, R., 2000. Influence of petroleum hydrocarbon contamination on microalgae and microbial activities in a longterm contaminated soil. Arch. Environ. Contam. Toxicol. 38, 439–445. Nortcliff, S., 2002. Standardization of soil quality attributes. Agric. Ecosyst. Environ. 88, 161–168. Pe´rez de Mora, A., Ortega-Calvo, J.J., Cabrera, F., Madejo´n, E., 2005. Changes in enzyme activities and microbial biomass after ‘‘in situ’’ remediation of a heavy metal-contaminated soil. Appl. Soil Ecol. 28, 125–137. Plaza, G., Nat˛ecz-Jawaki, G., Ulgih, K., Bridmon, R.L., 2005. The application of bioassays as indicators of petroleum-contaminated soil remediation. Chemosphere 59, 289–296. Ro¨mbke, J., Notenboom, J., Posthuma, L., 2002. The effects of zinc on enchytraeids: the Budel case study. Natura Jutlandica Occasional Papers 2, 54–67. Saterbak, A., Toy, R.J., Wong, D.C.L., McMain, B.J., Williams, M.P., Dorn, P.B., Brzuzy, L.P., Chai, E.Y., Salanitro, J.P., 1999. Ecotoxicological and analytical assessment of hydrocarbon-contaminated soils and application to ecological risk assessment. Environ. Toxicol. Chem. 18, 1591–1607. Speir, T.W., Kettels, H.A., Parshotam, A., Searle, P.L., Vlaar, L.N.C., 1995. A simple kinetic approach to derive the ecological dose value, ED50, for the assessment of Cr(VI) toxicity to soil biological properties. Soil Biol. Biochem. 27, 801–810. Speir, T.W., Kettles, H.A., Percival, H.J., Parshotam, A., 1999. Is soil acidification the cause of biochemical response when soils are amended with heavy metal salts? Soil Biol. Biochem. 31, 1953–1961. Torstensson, L., 1997. Microbial assays in soils. In: Tarradellas, J., Bitton, G., Rossel, D. (Eds.), Soil Ecotoxicity. CRC Lewis Publishers, London, pp. 207–233. Vance, E.D., Brookes, P.C., Jenkinson, D.S., 1987. Microbial biomass measurements in forests soils: determination of Kc values and test of hypothesis to explain the failure of the chloroform fumigation-incubation method in acid soils. Soil Biol. Biochem. 19, 381–387. Zhou, D.M., Deng, C.F., Cang, L., 2004. Electrokinetic remediation of a Cu contaminated red soil by conditioning catholyte pH with different enhancing chemical reagents. Chemosphere 56, 265–273. Zhou, D.M., Cang, L., Alshawabkeh, A.N., Wang, Y.J., Hao, X.Z., 2006. Pilot-scale electrokinetic treatment of a Cu contaminated red soil. Chemosphere 63, 964– 971. Zucconi, F., Pera, A., Forte, M., de Bertoldi, M., 1981. Evaluating toxicity of immature compost. Biocycle 22, 54–57.