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Journal of Soils and Sediments

, Volume 18, Issue 6, pp 2259–2270 | Cite as

Trace elements bioavailability to Triticum aestivum and Dendrobaena veneta in a multielement-contaminated agricultural soil amended with drinking water treatment residues

  • Silke Neu
  • Ingo Müller
  • Carsten Brackhage
  • Rafał Gałązka
  • Grzegorz Siebielec
  • Markus Puschenreiter
  • E. Gert Dudel
Reclamation and Management of Polluted Soils: Options and Case Studies

Abstract

Purpose

The in situ stabilization of multielement-contaminated agricultural soils has limited effectiveness when using common single amendments. This study examined the use of drinking water treatment residues (WTR), based on (hydr)oxides of Fe, Al, or Mn, as a cost-effective solution to optimize the immobilization of metals (Cd, Pb, Zn) and As.

Materials and methods

Trace elements (TE) bioavailability was assessed under semi-controlled conditions in a pot study cultivating winter wheat (Triticum aestivum L. cv. Tiger) until maturity. An Fe-based WTR and a Mn-based WTR, applied at rates of 0.5 and 1% (m/m), were related to effects of lime marl (LM) application. Additionally, a bioassay with earthworms (Dendrobaena veneta) was conducted. Both bioassays were compared with measurements of NH4NO3-soluble, diffusive gradients in thin film (DGT)-available and soil solution TE concentrations, representing well-established surrogates for mimicking the bioavailable element fractions in soil.

Results and discussion

The application of the Fe-based WTR reduced As accumulation in vegetative wheat tissues (by up to 75%) and earthworms (by up to 41%), which corresponded with the findings from soil chemical analyses and improved plant growth and earthworm body weight. However, As concentrations in cereal grains were not affected, Cd or Pb accumulation by wheat was not mitigated, and Zn uptake was enhanced. By contrast, the Mn-based WTR effected the greatest reduction in Pb uptake, and lowered Cd transfer to wheat grain (by up to 25%). Neither the NH4NO3-soluble nor DGT-available concentrations matched with Cd and Zn accumulation in plants or earthworms, indicating interferences due to competition for binding sites according to the biotic ligand model.

Conclusions

The results obtained in this study suggest that a bioassay with key species prior to field application should be mandatory when designing in situ stabilization options. The application of WTR to an agricultural soil strongly affected TE bioavailability to plants and earthworms. Low application rates tended to improve biomass production of biota. Higher application rates involved risks (e.g., P fixation, TE inputs), and none of the amendments tested could immobilize all targeted elements.

Keywords

Bioassay Fe oxide In situ stabilization Lime marl Mn oxide 

1 Introduction

Centuries of ore mining and processing have led to extensive contamination of topsoils with trace elements (TE) at many sites worldwide. The long history of mining in the Saxon Ore Mountains (Erzgebirge, Germany) has led to large-scale contamination with TE including As, Cd, Pb, and Zn, endangering current agricultural food and animal feed production (SMI 2013; BfUL 2015). Thus, efforts are needed to manage this pollutant linkage. Since conventional remediation techniques are not viable for treating trace element-contaminated soils (TECS) over a large area, gentle remediation options (GRO) offer alternative solutions for their sustainable management (Onwubuya et al. 2009; Cundy et al. 2015). The decontamination of multielement-contaminated soils by phytoextraction turned out to be infeasible in a reasonable time frame, even if supported by mobilizing agents (Nowack et al. 2006; Van Nevel et al. 2007). Thus, recent approaches concentrated on reducing the labile pool of TE. Thereby, in situ stabilization by use of soil amendments may enable simultaneous soil remediation and maintenance of agricultural production (Kumpiene 2010; Bolan et al. 2014).

The remediation of TE with partially antagonistic behavior (e.g., Cd, Pb, Zn, and As) requires the use of appropriate multifunctional stabilizing agents. Oxides and hydroxides of Fe, Mn, or Al are known to immobilize a wide range of metal(loids) in soils via the provision of sorption sites for TE, coprecipitation or formation of secondary minerals, or inner-sphere complexes with TE (Kumpiene 2010; Tiberg et al. 2016). These oxides and hydroxides are enriched to high concentrations in drinking water treatment residues (WTR) during the production of potable water. Land application of WTR for remediation purposes is therefore seen as a potentially important alternative to disposal of this by-product since it is available for free or at low costs. However, in addition to macronutrients, WTR can also contain considerable amounts of TE, depending on the origin and geochemical background of the pre-treatment water (Ippolito et al. 2011). Nonetheless, this approach has shown significant potential, as demonstrated by Nielsen et al. (2011), who showed that a single application of WTR (consisting mainly of ferrihydrite) stabilized As in pore water for a period of 3 years under field conditions. By contrast, liming agents have to be applied periodically in order to maintain soil pH at a level enabling the effective immobilization of metal cations (Bolan et al. 2014).

Wheat is one of the most widely cultivated crops worldwide and has the potential to take up considerable amounts of TE (Jamali et al. 2009; Kidd et al. 2015). During uptake, the presence of macronutrients or other TE in the soil, due to either natural abundance or through additives (e.g., from WTR application), may affect the accumulation by roots as well as translocation of TE within the plant (Foy et al. 1978; Siedlecka 1995; Pigna et al. 2010). In addition, the abundance and diversity of soil invertebrates may be diminished in TECS. Uptake of TE by earthworms is mainly controlled by passive uptake kinetics from soil pore water, with some influence of soil acidity (Spurgeon 2010). However, at mixed contamination, pronounced interactions between elements have been observed (e.g., Cd and Zn; Qiu et al. 2011).

Therefore, assessment tools of long-term TE bioavailability to key species in WTR-treated soils are of crucial importance. Basta et al. (2005) showed that conventional single or sequential extraction methods are unable to mimic bioavailability in soil-residual systems. Several studies have found that DGT is a better predictor of phytoavailability, in cases where plant uptake was limited by diffusive transport to roots and not saturated (Degryse et al. 2009; Muhammad et al. 2012; Puschenreiter et al. 2013). The DGT technique accounts for the depletion of TE at the interface of soil and roots or organisms, and the depletion-induced resupply of TE from labile pools of the solid phase to the soil solution (Zhang et al. 2001). This technique has also been found to mimic TE uptake by earthworms and human bioaccessibility (Bade et al. 2012), which in turn could be reduced by the use of WTR regarding As (Sarkar et al. 2007). Yet, DGT was not applied to soils treated with WTR, where an influx of various elements is likely to affect plant uptake of TE without simultaneous effects on DGT fluxes (Degryse et al. 2009). The objectives of this study are therefore (i) to test the potential of WTR based on (hydr)oxides of Fe, Al, or Mn for immobilizing Cd, Pb, Zn, and As in an agricultural soil with bioassays (Triticum aestivum L. cv. Tiger and Dendrobaena veneta) and (ii) to test the suitability of a conventional (NH4NO3) chemical soil extraction and DGT to assess TE bioavailability in WTR-treated soils.

2 Materials and methods

2.1 Soils and biota

Topsoil material (0–20 cm) was taken from an agricultural site contaminated with TE (Cd, Pb, Zn, and As). The soil (Con) was obtained from the control treatment of a phytoremediation field trial in Freiberg, Saxony (Kidd et al. 2015; GPS coordinates: 13° 23′ 52″ E, 50° 54′ 17″ N). Uncontaminated topsoil material (Ref) was obtained from an agricultural site nearby (GPS coordinates: 13° 16′ 49″ E, 50° 57′ 47″ N), where TE concentrations were within the range of background levels for the region (Kardel et al. 2015). This soil was used as a reference regarding element concentration and biomass production of biota. Soils Ref and Con are classified as stagnosols, derived from loess and paragneiss, and were sampled at locations with an altitude of 420 m, 630 mm average annual precipitation, and 8 °C average annual temperature (Barth and Forberg 2015).

Chemically sterilized commercial seeds of T. aestivum L. cv. Tiger were inoculated with vesicular-arbuscular mycorrhizae (VAM; Rhizophagus irregularis) using 300 l ha−1 INOQSpezial (INOQ GmbH, Schnega, Germany) in order to minimize variability under semi-controlled conditions. Compost earthworms (D. veneta) were chosen as the soil invertebrates for assessment.

Two different WTR were compared; amendment WTRA is a carbonate-enriched sludge from Wittkoppenberg waterworks in Germany, where Fe-rich groundwater was treated (Müller 2000; Marschner et al. 2008), and amendment WTRB is a Mn-rich sludge from Oborniki, Poland (Siebielec et al. 2013). The LM amendment was tested in the phytoremediation field trial parallel to this pot study. General characteristics of all amendments are presented in Table 1.
Table 1

General soil and amendment characteristics

  

Soil

  

Amendment

  

Ref

Con

  

LM

WTRA

WTRB

pH (H2O)

6.1

5.6

pH (H2O)

9.1

7.9

7.9

C org

G kg −1

11

21.6

C

g kg−1

121

20

9

CEC pot

cmol c  kg −1

15.8

21.5

Corg

g kg−1

NA

11

3

Skeleton

g kg−1

2.2

2

Ca

g kg−1

205

15

16

Sand

g kg −1

90

350

Fe

g kg−1

4.7

354

18

Silt

g kg −1

770

510

Al

g kg−1

5

22

6

Clay

g kg −1

140

140

Mn

g kg−1

0.4

6

72

Field capacity

g kg−1

220

220

P

g kg−1

<DLa

3.3

1.6

As aqua regia

mg kg−1

41

868.4

As

mg kg−1

8.9

57.2

13

As NH4NO3

mg kg−1

0.02

0.8

Cd

mg kg−1

0.5

1

4

Cd aqua regia

mg kg−1

0.7

14.4

Pb

mg kg−1

17.1

28

3.6

Cd NH4NO3

mg kg−1

0.01

0.6

Cr

mg kg−1

11.6

3.5

235.8

Pb aqua regia

mg kg−1

77.7

1635.4

Ni

mg kg−1

73.5

43

435

Pb NH4NO3

mg kg−1

0.01

1.9

Zn

mg kg−1

5.4

80

677.3

Zn aqua regia

mg kg−1

80.0

437.3

Tl

mg kg−1

0.2

0.02

1.8

Zn NH4NO3

mg kg−1

0.4

2.6

Cu

mg kg−1

NA

5.5

49.4

AV Cd (mg kg−1 NH4NO3)

0.04b/0.1

     

TV Pb (mg kg−1 NH4NO3)

0.1

     

TV As (mg kg−1 aqua regia)

200

     

TV As (mg kg−1 NH4NO3)

0.4c

     

TV Zn (mg kg−1 NH4NO3)

2c

     

Information in italics are derived from Barth and Forberg (2015)

Ref uncontaminated soil, Con untreated contaminated soil, LM lime marl, WTR A Fe-based drinking water treatment residue, WTR B Mn-based drinking water treatment residue

a0.25 g kg−1

bAction value (AV) according to BBodSchV (1999) if bread wheat or accumulating vegetables are cultivated

cTrigger values (TV) according to BBodSchV (1999) for expected impairment of plant growth

2.2 Experimental setup

Prior to use in the pot experiment, the soil material was homogenized, sieved to < 4 mm, and treated with steam. The WTR amendments were freeze-dried and applied to the soil at rates of 0.5 and 1% by dry weight. The WTRA treatments will hereafter be referred to as A-0.5 (0.5% application) and A-1 (1% application) and analogous for WTRB (B-0.5 and B-1). Lime marl was added at a rate of 0.4 kg m−2, corresponding to the field trial. All amendments were sieved to < 63 μm and thoroughly homogenized with the soil material. The substrates were watered with de-ionized (DI) water to 70% of field capacity for seven days and placed in 13-l white polyethylene vessels in four replicates. An initial number of 22 seeds per pot were sown in October 2013. During wintertime, pots were arranged outside in a sand bed for vernalization. In early spring, the number of seedlings was reduced to 16 plants per pot, corresponding to plant densities in the field. Pots were randomly set up in greenhouses with filtered ambient air. The soils were fertilized with amounts corresponding to 90 kg ha−1 N (CH4N2O), 18 kg ha−1 P (P fertilizer with 40% P2O5), and 100 kg ha−1 K (K fertilizer with 60% K2O). A second N fertilization was done with (NH4)2SO4 corresponding to 42 kg ha−1 N during bolting. Water content was maintained close to field capacity by daily watering with DI water. When the flag leaf was fully developed, six leaves per pot were sampled across subjacent leaf levels in order to assess TE and nutrient status of the plants. During harvest, biomass was separated into grain, straw, and roots (captured by sieving). Fresh soil samples were sieved to < 2 mm as required by German legislation (BBodSchV 1999). Aliquots were air dried for use in the earthworm experiment and for chemical analyses of soil other than DGT.

The earthworm experiment followed the method used by Siebielec (2010), and was conducted in three replicates per treatment. After incubation at 20 ± 2 °C at daytime with moisture equal to field water capacity, 0.5-kg soil portions were transferred to 1-l glass jars. The initial weight of earthworms was recorded prior to putting five individuals into each jar. The jars were stored at 15 °C and soil moisture was maintained at field water capacity. Earthworms were removed from the soil after four weeks, and mortality and weight were recorded.

2.3 Chemical analyses

Detailed information on the chemical analyses described in the following sections is provided in Online Resource 1. All chemicals were of analytical grade, and extractions were made up with DI water (18 MΩ cm−1, E-pure system). The validity of chemical analyses was confirmed by including blanks and standard reference material BCR 281, rye grass (EU-JRC) for plant analyses, and WEPAL-ISE 979 (Rendzina soil) for aqua regia and NH4NO3 extraction. Microwave-assisted acid digestions were conducted using the system MARS5 (CEM Corp., Matthews, USA) unless otherwise stated. Measurements of TE in extraction solutions were performed by ICP-MS (PQ ExCell, Thermo Fisher Scientific Inc., UK). Nutrients and effective cation exchange capacity (CECeff) were measured by ICP-OES (IRIS Intrepid II XSP, Thermo Fischer Scientific).

2.3.1 Amendment and soil analysis

Element concentrations of the amendments were determined following digestion with aqua regia. For characterization of Con and Ref, pseudo-total and potentially plant-available element fractions were analyzed using standard procedures of aqua regia (DIN ISO 11466: 1997) and NH4NO3 (DIN ISO 19730: 2008). Effective cation exchange capacity (CECeff) was assessed according to Section 3.2.1.8 of BMEL (2014). Soil pH was analyzed initially and after harvest as described in DIN ISO 10390: 2005. Remaining fresh soil samples were analyzed with DGT according to Zhang et al. (2001) in separate steps for metals (chelex gel) and As (Fe oxide gel). To assess the concentration of TE in soil solution (Csoln), the remaining water-saturated soils were centrifuged according to Muhammad et al. (2012) and the filtered supernatant was analyzed by ICP-MS. The interfacial TE concentration (CDGT) was calculated according to Zhang et al. (2001), using a diffusion coefficient of 5.69 × 10−6 cm2 s−1 for As according to Fitz et al. (2003). The dimensionless indicator for the extent of TE resupply from labile pools of the solid phase to the soil solution R (Harper et al. 2000; Zhang et al. 2001) was calculated as the ratio between CDGT and Csoln.

2.3.2 Plant and earthworm analysis

Plant material was washed with DI water and subsequently dried at 60 °C to constant weight. Samples were finely ground and digested using HNO3 and H2O2 (DIN EN 13805: 2014–12).

Earthworms were cleaned and kept on moist filter papers in glass vessels for three days for full depuration. The filter papers were replaced once per day to prevent secondary contamination of earthworms by their excretions. The earthworms were washed with DI water, dried on paper towels, and subsequently lyophilized. They were digested using HNO3 and H2O2.

2.4 Statistical analyses

The pot experiment was conducted in a completely randomized design. Treatment effects on element concentrations and plant biomass production were evaluated using one-way analysis of variance (ANOVA), followed by pairwise comparison using the Bonferroni post hoc test for adjustment of probabilities. Statistical analyses were performed using PASW Statistics 21 (SPSS, Inc., Somers, NY, USA).

3 Results

3.1 Soil characteristics and changes in trace elements mobility

The main soil characteristics are shown in Table 1. Soil pH values did not change significantly after 7 days of incubation compared to the initial level of pH 5.6 (moderately acid), except for a decrease in treatment B-1. By contrast, CECeff increased in A-1 and B-0.5 (Fig. 1). Until harvest, pH increased in treatments LM, A-0.5, and A-1 (each p < 0.01) to weakly acid conditions (Fig. 1), whereas WTRB did not affect pH (B-0.5) or slightly enhanced it (B-1; p < 0.01).
Fig. 1

Initial pH (pHi) and CECeff (CECi cmolc kg −1 ) and pH at harvest (pHh) in untreated (Con) and treated soils (LM lime marl, A-x- and B-x Fe-based (WTRA) and Mn-based (WTRB) drinking water treatment residue, where x is the application rate in % (m/m)). Values are means ± SD (n = 4)

Soil Con was characterized by high labile concentrations of As, Cd, Pb, and Zn, which exceed German trigger or action values for agricultural soils (BBodSchV 1999; Table 1). At the end of the pot experiment, all utilized methods (NH4NO3, CDGT, and Csoln) indicated a reduction of TE mobility by both WTR beyond the effect of pH modification (Fig. 2, Online Resource 2, Table S1—Electronic Supplementary Material). The pH values increased at a similar rate in LM and A-1, but (with the exception of Zn) TE mobility decreased the most in A-1 (As by ≤85%, Cd by ≤80%, Pb by ≤62%). Arsenic mobility still decreased by ∼60% in treatment A-0.5, where Cd and Pb were also immobilized effectively. In treatment B-0.5, TE mobility showed the lowest decrease, accompanied by a negligible change of pH compared to Con. Both WTR similarly reduced labile Zn, with higher doses being more effective. However, no treatment reduced mobile TE to the level of Ref. Along with a high total concentration of Ni in WTRB, an increase of NH4NO3-soluble Ni within the tolerable range (BBodSchV 1999) was observed in both treatments (Online Resource 2, Table S5—Electronic Supplementary Material).
Fig. 2

Trace elements mobility in treated soils (LM lime marl, A-x- and B-x Fe-based (WTRA) and Mn-based (WTRB) drinking water treatment residue, where x is the application rate in percentage (m/m)) at the end of the pot experiment expressed as percentage of untreated soil (Con). Values are means ± SD (n = 4)

The order and significance of treatment effects on the reduction of TE mobility varied with respect to the analytical method used (Fig. 2, Online Resource 2, Table S1—Electronic Supplementary Material). The potential resupply of depleted TE from the solid phase to the soil solution as assessed by the indicator R (CDGT/Csoln) tended to increase in treated soils when compared to Con (Online Resource 2, Table S2—Electronic Supplementary Material).

3.2 Plant biomass production, nutritional status, and earthworm performance

Insignificant (p > 0.05) changes in biomass production of wheat plants were observed with the different treatments. The supply of LM or low levels of WTR enhanced grain yield, but higher application rates of both WTR reversed this trend (Fig. 3a). This was accompanied by changes in the nutritional status recorded by leaf analyses (Online Resource 2, Table S3—Electronic Supplementary Material). While Con and Ref had a P leaf concentration of 3.5 g kg−1 DW−1, the P status decreased by 38% in A-0.5 and by 50% in A-1. The concentration of Ca in leaves was significantly lower in all treatments but LM compared to Con.
Fig. 3

Variability in a grain and straw biomass production of wheat (Triticum aestivum L. cv. Tiger) and b weight loss per earthworm (Dendrobaena veneta) expressed as percentage of initial weight from untreated (Con) and treated soils (LM lime marl, A-x and B-x Fe-based (WTRA) and Mn-based (WTRB) drinking water treatment residue, where x is the application rate in percentage (m/m)). Values are means ± SD (n = 4). Average earthworm, grain (black), and straw biomass (gray) in uncontaminated soil (Ref) are represented by dashed lines. Significant differences (p < 0.05) are indicated by an asterisk. The number of perished worms is indicated within the respective bars

Overall, the earthworms showed little mortality (Fig. 3b). Moderate mortality occurred in Con, LM, and B-1, where earthworms were also characterized by a loss of body weight. Compared to these treatments, weight loss was significantly lower in treatment A-0.5 (p < 0.05).

3.3 Changes in chemical elements accumulation in biota

As shown in Fig. 4, the amendments altered TE concentration in plant and earthworm tissue when compared to Con. Samples treated with WTRA showed a strong decrease of As transfer into straw (40% in A-0.5 and 48% in A-1 (p < 0.05); see also Online Resource 2, Table S4—Electronic Supplementary Material) and leaves (68 and 75% (p < 0.01)) as well as into earthworm tissue (26% (p < 0.05) and 41% (p < 0.01)). Root uptake of Pb decreased in A-0.5 (p < 0.05) and A-1 (p < 0.01), whereas uptake of Cd increased in A-0.5 (p < 0.05) (see also Fig. 5a). In grains, both WTRA treatments enhanced Zn concentration (p < 0.01), whereas WTRB lowered the Cd status (B-0.5 by 23% and B-1 by 25% (p < 0.05)). Treatment B-1 reduced Pb concentration in straw, leaves (both p < 0.01), and earthworm tissue (p < 0.05) as well as leaf As (p < 0.01). Treatment LM only reduced root uptake of Pb (p < 0.05). However, the TE status of biota in treated soils did not reach levels as low as in Ref, except for Zn (grain, leaves) and Pb (leaves). Besides the TE listed above, enhanced Ni concentrations were found in biota exposed to WTRB treatments (Online Resource 2, Table S5—Electronic Supplementary Material).
Fig. 4

Concentration of As (a), Cd (b), Pb (c), and Zn (d) in plant compartments of Triticum aestivum L. cv. Tiger and in tissue of Dendrobaena veneta from untreated (Con) and treated soils (LM lime marl, A-x and B-x Fe-based (WTRA) and Mn-based (WTRB) drinking water treatment residue, where x is the application rate in percentage (m/m)). Values are means ± SD (n = 4). Means of Ref are represented by dashed lines. Asterisks represent significant differences between treatments and Con, *for p < 0.05; **for p < 0.01

3.4 Correlation of tissue concentrations in Triticum aestivum and Dendrobaena veneta with soil chemical analyses

Tables 2 and 3 show the Pearson correlation coefficients between TE concentration in soil, biomass production, and TE concentration in plants and earthworms. Thousand grain weight (TGW) of wheat was inversely correlated with As (NH4NO3), Cd (CDGT), and Zn (CDGT > Csoln) in soil. Shoot biomass or shoot TE removal did not correlate with soil chemical characteristics (data not shown). Correlations were found between labile As concentrations in the soil and As uptake into straw (r ≤ 0.79; p < 0.01) or leaves (r ≤ 0.86; p < 0.01), and between NH4NO3-soluble Pb concentration and Pb uptake into roots (r = 0.77; p < 0.01) or leaves (r = 0.59; p < 0.01). For all other combinations, the correlation was weak, inverse, or insignificant. Overall, NH4NO3 provided predictability equal to or better than CDGT, whereas Csoln was a better predictor than CDGT for Pb.
Table 2

Pearson correlation coefficientsa between NH4NO3-soluble (NH4NO3) or DGT-available (CDGT) or soil solution (Csoln) TE concentrations in contaminated soil (untreated and treated) and thousand grain weight (TGW) or TE concentrations in plant compartments of wheat (Triticum aestivum L. cv. Tiger)

 

TGW

As root

As straw

As leaf

As grain

As NH4NO3

−0.56 (<0.01)

−0.23 (0.28)

0.79 (<0.01)

0.86 (<0.01)

−0.53 (<0.01)

As CDGT

−0.39 (0.07)

−0.39 (0.85)

0.75 (<0.01)

0.78 (<0.01)

−0.54 (<0.01)

As Csoln

−0.40 (0.06)

−0.11 (0.63)

0.48 (0.02)

0.65 (<0.01)

−0.40 (0.06)

 

TGW

Cd root

Cd straw

Cd leaf

Cd grain

Cd NH4NO3

−0.40 (0.06)

−0.61 (<0.01)

−0.55 (<0.01)

−0.05 (0.80)

0.05 (0.78)

Cd CDGT

−0.64 (<0.01)

−0.74 (<0.01)

−0.65 (<0.01)

−0.07 (0.76)

−0.15 (0.50)

Cd Csoln

−0.11 (0.63)

−0.33 (0.12)

−0.11 (0.61)

0.16 (0.48)

0.01 (0.97)

 

TGW

Pb root

Pb straw

Pb leaf

Pb grain

Pb NH4NO3

−0.31 (0.15)

0.77 (<0.01)

−0.04 (0.87)

0.59 (<0.01)

−0.32 (0.14)

Pb CDGT

−0.34 (0.11)

0.19 (0.37)

−0.11 (0.62)

0.24 (0.27)

−0.25 (0.26)

Pb Csoln

−0.07 (0.76)

0.48 (0.02)

−0.42 (0.04)

0.03 (0.87)

−0.51 (0.02)

 

TGW

Zn root

Zn straw

Zn leaf

Zn grain

Zn NH4NO3

−0.65 (0.77)

−0.40 (0.06)

−0.37 (0.08)

−0.20 (0.35)

−0.23 (0.30)

Zn CDGT

−0.46 (0.03)

−0.11 (0.62)

−0.40 (0.06)

−0.25 (0.25)

−0.41 (0.05)

Zn Csoln

−0.44 (0.04)

0.53 (<0.01)

−0.65 (<0.01)

−0.50 (0.01)

−0.58 (<0.01)

If significant (p < 0.05), data are highlighted with bold font

aValues in parentheses represent the significance of the correlation (p value)

Table 3

Pearson correlation coefficientsa between NH4NO3-soluble (NH4NO3) or DGT-available (CDGT) or soil solution (Csoln) TE concentrations in contaminated soil (untreated and treated) and biomass of earthworms (Dendrobaena veneta) expressed as percentage of initial weight and TE concentrations in earthworm tissue

 

Biomass

As tissue

As NH4NO3

−0.65 (<0.01)

0.86 (<0.01)

As CDGT

−0.42 (0.08)

0.80 (<0.01)

As Csoln

−0.45 (0.06)

0.81 (<0.01)

 

Biomass

Cd tissue

Cd NH4NO3

−0.50 (0.03)

−0.29 (0.23)

Cd CDGT

−0.62 (<0.01)

−0.47 (0.05)

Cd Csoln

−0.33 (0.18)

−0.54 (0.02)

 

Biomass

Pb tissue

Pb NH4NO3

−0.48 (0.04)

0.25 (0.31)

Pb CDGT

−0.25 (0.33)

0.08 (0.75)

Pb Csoln

−0.31 (0.22)

−0.11 (0.66)

 

Biomass

Zn tissue

Zn NH4NO3

−0.21 (0.40)

0.30 (0.23)

Zn CDGT

−0.34 (0.17)

0.27 (0.28)

Zn Csoln

−0.47 (<0.05)

−0.05 (0.83)

If significant (p < 0.05), data are highlighted with bold font

aValues in parentheses represent the significance of the correlation (p value)

Earthworm biomass was inversely correlated with soil As (NH4NO3), Cd (CDGT > NH4NO3), Pb (NH4NO3), and Zn (Csoln). In contrast to the other TE, As concentration in earthworm tissue correlated well (r ≤ 0.86; p < 0.01) with labile As in soil measured by all kinds of chemical analyses.

3.5 Correlations between TE accumulation and biomass production of biota

Cadmium uptake into plant roots (r = 0.53; p < 0.05), As uptake into vegetative tissue (straw r = 0.63; p < 0.01, leaves r = 0.70; p < 0.01), and Pb uptake into all aboveground plant compartments (straw r = 0.57; p < 0.05, leaves r = 0.57; p < 0.05, grain r = 0.52; p < 0.05) correlated with the accumulation of the respective TE in earthworm tissue. Earthworm biomass correlated with TGW (r = 0.67; p < 0.01) and grain dry weight (r = 0.50; p < 0.05).

4 Discussion

4.1 Characteristics of WTR influencing their immobilization efficiency for TE and implications for the use in agricultural soils

The efficiency of Fe oxide-containing soil additives for As immobilization has been examined by numerous studies (e.g., Basta et al. 2005; Kumpiene et al. 2008). Moore et al. (2000) found an Fe/As molar ratio of 2 and higher to be the most effective. The treatments with WTRA represent an Fe/As molar ratio of 2.7 (A-0.5) and 5.5 (A-1), respectively, with increasing stabilization efficiency for As indicated by the soil chemical analyses and concentration in biota. The stimulation of Zn uptake by wheat in WTRA treatments supports the findings for Fe oxide treatments by Mench et al. (2006). The results indicate that WTRA, consisting of Fe and Al compounds, reduced oxyanion availability in soils with subsequent P limitation in crops, as suggested by Sarkar et al. (2007). Frequent P fertilization in agricultural soils may, in turn, lead to the mobilization of As due to a competition between P and As for sorption sites. However, according to these authors, As was long-term stabilized even in WTR-treated sandy soils fertilized with triple super phosphate due to the formation of inner-sphere complexes. Within weakly to moderately acid soil conditions, Mn oxides were reported to immobilize TE like Cd, Pb, and Zn more effectively than Fe hydroxides (Mench et al. 1999, 2006). Accordingly, WTRB most successfully decreased Cd (B-0.5) and Pb (B-1) in the soil solution, and exclusively reduced Cd status of wheat grains (B-0.5 and B-1). Both WTR reduced labile TE in a moderately acid soil under semi-controlled conditions. However, if WTR are used in agricultural soils under field conditions, higher fluctuations in pH (liming) and Eh (heavy rainfall events and irrigation) are likely to hamper effective surface sorption of cations and anions (Kumpiene et al. 2008; Bolan et al. 2014).

Depending on the geochemical origin of the processed natural water, WTR may contain high amounts of potentially toxic elements (Ippolito et al. 2011). Measured concentrations of As, Cd, and Pb in WTR used in this study corresponded with ranges reported by other authors, whereas concentrations of Zn, Cr, and Ni in WTRB were relatively high (Ippolito et al. 2011; Nielsen et al. 2011). No enhanced concentrations of As, Cd, Pb, or Zn in treated soils or biota were found that could be related to an introduction by WTR, indicating strong sorption onto (hydr)oxide surfaces. However, labile (NH4NO3) concentrations of Ni were significantly enhanced in WTRB treatments (p < 0.05), albeit at a moderate level.

4.2 Effects of WTR on plants and earthworms

A marked dose responsive decrease of P in leaf tissue was recorded in WTRA treatments. Hence, a strong fixation of P parallel to that of As in soil by Fe oxides might be confirmed (Sarkar et al. 2007). Nevertheless, A-0.5 showed a non-significant increase of plant dry-matter yield compared to Con. For a higher application rate (A-1), the increase, particularly of grain biomass, reversed almost to the level of Con. Hence, alleviated TE toxicity might be counteracted by a deficiency in P, as found earlier in plants grown on soils treated with Fe- or Al-based WTR (Müller and Pluquet 2000; Ippolito et al. 2011). Earthworm mortality was prevented and body weight was enhanced by WTRA, but as with plants, a higher application rate impaired the positive response. However, the results obtained in this study suggest that reduced As bioavailability improves the health of earthworms, which in turn implies improvement of soil functionality (Nahmani et al. 2007). In the case of WTRB, plant biomass and earthworm body weight and mortality responded negatively to a higher application rate as well. This may be attributed to enhanced Ni uptake into plant roots and grains in case of B-1 and by earthworms (p < 0.05) in both treatments, pointing to oxidative stress or inhibition of nutrient uptake (Maleri et al. 2007; Chen et al. 2009).

4.3 Factors limiting the predictability of TE status of biota by analytical methods

The predictability of As concentration in vegetative plant organs and earthworm tissue was satisfactory (r ∼ 0.8) by the selected methods. For oligochaete worms, uptake routes for As are typically related to the soil pore water compartment and further affected by soil characteristics, in particular acidity (Peijnenburg et al. 1999). Despite the strong reduction of available As in soil, root uptake was not lowered by WTRA treatments, but translocation to shoots (Online Resource 2, Table S4—Electronic Supplementary Material). Although leaf P status simultaneously decreased, the molar ratio of P/As was still significantly higher when compared to the other treatments. The influence of P on the mitigation of As uptake and translocation in durum wheat (Triticum durum L.) and winter wheat (T. aestivum L.) has been intensively studied by various authors, who also found most of As being retained in the roots (Pigna et al. 2009; Brackhage et al. 2014). No significant changes in leaf Fe status were identified, from which further conclusions could be drawn about its role in As translocation within the plant. In addition to phosphate transporters, the pathway of silicic acid was found to be relevant for the accumulation of arsenous acid by plants (Zhao et al. 2010). This study did not focus on interactions with Si. However, Si should be considered in future studies. If phytoavailable, it might influence As and Cd (Rizwan et al. 2012) uptake and/or translocation in treatments with WTR, which can contain considerable amounts of Si (Nielsen et al. 2011).

Root uptake of Pb was correlated with Pb concentrations measured in NH4NO3− and soil solution (Csoln). The latter was superior to CDGT, which might be due to overestimation of metal concentrations in roots since the fraction bound to the apoplast was not removed before measuring root TE concentrations. Furthermore, Pb translocation from plant roots to shoots is known to be limited by selective cellular uptake mechanisms (Nowack et al. 2006). The weak correlations obtained with chemical extractants or CDGT also comply with the findings of Senila (2014).

For soils at or near an equilibrium state, significant correlations are typically found between Cd and Zn concentrations extracted with NH4NO3 or DGT and their concentration in plants (Gryschko et al. 2005; Senila et al. 2012). By contrast, the observed correlations were either inverse or insignificant, which corresponds with the findings from Huynh et al. (2010), who attributed it to plant removal from the labile pool. This was indicated by high concentrations of Zn in plant tissue found in this study as well. Since root uptake of Zn correlated best with Csoln, Zn uptake was not limited by diffusion, which was supported by a considerable resupply to the soil solution (see Online Resource 2, Table S2). According to the biotic ligand model (Spurgeon 2010), the uptake of Cd as a non-essential element for plants and organisms like earthworms is determined by the presence of competing cations like Ca2+ and Zn2+ as well as protons (Oste et al. 2001). Strong deviations were found between the ratio of Cd/Zn in roots and earthworms and Cd/Zn ratio in the available pool in soil. More precise, a relative enrichment of Cd in plants (Fig. 5a) and earthworms (Online Resource 2, Fig. S1—Electronic Supplementary Material) was found with decreasing Cd/Zn ratio in soil. Cadmium concentration in roots could be explained by a linear regression with pH (R 2 = 0.45, p < 0.001), while a multiple linear fitting including available Zn improved the correlation (R 2 = 0.56, p < 0.001). Thereby, higher concentrations of H+ and/or available Zn, represented by Con, B-0.5, and B-1, lowered root uptake of Cd possibly due to preferential binding of Zn and protons to root membranes and transfer via specific transporters. Within the plants, an inhibition of Cd translocation additionally modified the Cd/Zn ratio towards a relative enrichment of Zn, especially in the WTRA treatments (see Fig. 5b). In a chelant-buffered hydroponic study, Green et al. (2003) found that high ionic Zn2+ activities did not mitigate root uptake of Cd, but decreased its translocation to shoots. Consequently, when competing cations impact accumulation by plants and earthworms and further alter translocation within plants, neither chemical extractants nor DGT could serve as appropriate biomimic surrogates (Degryse et al. 2009).
Fig. 5

Concentration ratios between Cd and Zn of earthworm and plant tissue related to that in soil solution (Csoln) (a) and of aboveground tissues related to root tissue (b) in untreated (Con) and treated soils (LM lime marl, A-x and B-x Fe-based (WTRA) and Mn-based (WTRB) drinking water treatment residue, where x is the application rate in percentage (m/m))

5 Conclusions

The application of WTR strongly affected TE bioavailability in soil, TE tissue concentration, and biomass of plants and earthworms in this study. However, none of the treatments was able to mitigate the transfer of all targeted elements. The Fe-based WTRA was the most effective at reducing As accumulation, whereas the Mn-based WTRB was the most successful at reducing Pb and Cd (plant) accumulation in biota. In both cases, the biota responded negatively to higher application rates. Thus, the use of WTR for in situ stabilization of TE-contaminated agricultural soils can be quite effective, but it also holds risks related to nutrient limitation or input of potentially toxic elements. In this study, the transfer of TE to cereal grains, which is most relevant in practice, could not be deduced from chemical soil analyses. Likewise, observed reductions of TE concentration in leaves or earthworms, which might be assessed by short-term tests, were not indicative of grain TE status. The predictability of Cd and Zn bioavailability can be weak in soils with high labile shares of both elements and/or soils treated with WTR, where an influx of various elements causes interactions during uptake. Therefore, such amendments should be tested in a bioassay with a set of key species prior to application at field scale.

Notes

Acknowledgements

The authors gratefully thank Prof. B. Marschner for providing the drinking water treatment residue WTRA and A. Weiske for ICP-measurements. This work was financially supported by the European Commission under the Seventh Framework Programme for Research (FP7-KBBE-266124, Greenland).

Compliance with ethical standards

Conflict of interest

The authors declare that they have no conflict of interest.

Supplementary material

11368_2017_1741_MOESM1_ESM.docx (26 kb)
ESM 1 (DOCX 26.0 kb)
11368_2017_1741_MOESM2_ESM.docx (84 kb)
ESM 2 (DOCX 83.9 kb)

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Copyright information

© Springer-Verlag Berlin Heidelberg 2017

Authors and Affiliations

  1. 1.Institute of General Ecology and Environmental ProtectionTechnische Universität DresdenTharandtGermany
  2. 2.Saxon State Office for Environment, Agriculture and GeologyDresden PillnitzGermany
  3. 3.Institute of Soil Science and Plant CultivationState Research InstitutePulawyPoland
  4. 4.Department of Forest and Soil SciencesUniversity of Natural Resources and Life Sciences Vienna – BOKUTullnAustria

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