Nickel phytomining from industrial wastes: Growing nickel hyperaccumulator plants on galvanic sludges

Journal of Environmental Management 254 (2020) 109798

Contents lists available at ScienceDirect

Journal of Environmental Management journal homepage: http://www.elsevier.com/locate/jenvman

Research article

Nickel phytomining from industrial wastes: Growing nickel hyperaccumulator plants on galvanic sludges Alice Tognacchini a, *, Theresa Rosenkranz a, Antony van der Ent b, d, Gaylord Erwan Machinet c, Guillaume Echevarria d, Markus Puschenreiter a a

University of Natural Resources and Life Sciences, Vienna, Department of Forest and Soil Sciences, Institute of Soil Research, Konrad-Lorenz Straße 24, 3430 Tulln, Austria Centre for Mined Land Rehabilitation, Sustainable Minerals Institute, The University of Queensland, St Lucia, 4072, QLD, Australia c Microhumus, Universit�e de Lorraine, ENSAIA - Laboratoire Sols et Environnement, 2 avenue de la For^et de Haye, BP 20163, 54505, Vandoeuvre-l�es-Nancy, France d Universit�e de Lorraine, Laboratoire Sols et Environnement, 54000, Nancy, France b

A R T I C L E I N F O

A B S T R A C T

Keywords: Nickel Agromining Technosols Odontarrhena chalcidica

Nickel (Ni) is used in numerous industrial processes, with large amounts of Ni-rich industrial wastes produced, which are largely sent to landfill. Nickel recovery from waste materials that would otherwise be disposed is of particular interest. Nickel phytomining represents a new technology in which hyperaccumulator plants are cultivated on Ni-rich substrates for commercial metal recovery. The aim of this study was to investigate the possibility of Ni transfer from industrial waste into plant biomass, to support recovery processes from bio-ores. Different industrial galvanic sludges (containing 85–150 g kg 1 Ni) were converted into artificial substrates (i.e. technosols) and the Ni hyperaccumulator Odontarrhena chalcidica (formerly Alyssum murale) was cultivated on these Ni-rich matrices. A greenhouse pot experiment was conducted for three months including an ultramafic soil control and testing fertilized (NPK) and unfertilized replicates. The results showed that fertilization was effective in improving plant biomass for all the substrates and that O. chalcidica was capable of viably growing on technosols, producing a comparable biomass to O. chalcidica on the control (ultramafic soil). On all technosols, O. chalcidica achieved Ni shoot concentrations of more than >1000 mg Ni kg 1 and maximum Ni uptake was obtained from one of the technosols (26.8 g kg 1 Ni, unfertilized; 20.2 g kg 1 Ni, fertilized). Nickel accumu­ lation from three of the technosols resulted to be comparable with the control ultramafic soil. This study demonstrated the feasibility of transferring Ni from toxic waste into the biomass of Odontarrhena chalcidica and that phytomining from galvanic sludge-derived technosols can provide similar Ni yields as from natural ultra­ mafic soils.

1. Introduction Being a versatile metal, nickel (Ni) is used in a wide variety of in­ dustrial processes, with large amounts of Ni-rich waste being produced annually worldwide. Galvanic sludges represent a particular type of waste derived from Ni-related industries and, apart from Ni, are usually very rich in heavy metals such as zinc (Zn), copper (Cu), chromium (Cr) or cadmium (Cd). It is estimated that each medium to large plating in­ dustry generates, monthly, between two and three tons of these residues (Rossini and Bernardes, 2006), which are typically being disposed in landfill sites without any economical or environmental benefit (Vilar­ inho et al., 2012). For conventional Ni mining to be economically viable,

ore bodies with a Ni concentration of at least 30 g kg 1 are required (Li et al., 2003b), although ores containing ~10 g kg 1 Ni are currently mined due to globally declining ore grades (Mudd, 2009), with minimal cut-off concentrations of 9.0 g kg 1. Galvanic sludges are often richer in Ni than conventional ores and therefore could represent valuable sour­ ces of Ni that are instead being lost in landfills. Considering galvanic sludge as a secondary Ni resource would address the issue of toxic waste disposal and support the creation of closed loops industrial processes. Several treatment routes have been proposed to recover valuable metals from these types of industrial waste, mainly focusing on pyrometallur­ gical routes of recovery approaches (Bernardes et al., 1996), electro­ � et al., 2003) or hydrometallurgical (Rajcevic, 1990; winning (Veglio

* Corresponding author. E-mail address: [email protected] (A. Tognacchini). https://doi.org/10.1016/j.jenvman.2019.109798 Received 27 June 2019; Received in revised form 30 September 2019; Accepted 27 October 2019 Available online 15 November 2019 0301-4797/© 2019 Elsevier Ltd. All rights reserved.

A. Tognacchini et al.

Journal of Environmental Management 254 (2020) 109798

Silva et al., 2005; Vilarinho et al., 2012). However, these Ni recovery processes are seldomly applied on galvanic sludges due to the high costs and energy consumption; also, a certain nickel content is required in the waste to allow feasible extraction. Therefore, the possibility to develop alternative approaches for Ni recovery from galvanic wastes is of particular interest. A novel plant-based approach for metal recovery from secondary resources, defined as ‘phytomining’, was first developed by Chaney (1983) targeting Ni. Phytomining is defined as the process whereby plants capable of accumulating metals in their shoots are cultivated on mineral-rich substrates, with the objective of recovering the metal from their biomass for commercial gain (Chaney et al., 1998; van der Ent et al., 2015). Later, the term ‘agromining’ has been proposed (Morel, 2013, 2015), referring specifically to the entire agronomical chain of cultivating hyperaccumulator plants on sub economic ores and recovery metals from their biomass. Nickel hyperaccumulator species, as defined by Reeves et al. (1996) and van der Ent et al. (2013b), are plants able to accumulate at least 1000 mg kg 1 Ni in their leaves while growing on their natural habitat. So far, research and development of this technology has mainly concentrated on nickel (Chaney et al., 2007; Nkrumah et al., 2016), with special regard to phytomining from natu­ rally Ni enriched (ultramafic) soils which occur across the globe. Suc­ cessful examples can be found in literature showing the feasibility of nickel recovery from ultramafic soils. Bani et al. (2015a; 2015b) implemented extensive agromining in Albania, extracting Ni from ul­ tramafic soils with the hyperaccumulator plant Odontarrhena chalcidica � (formerly Alyssum murale, Spaniel et al., 2015). Agromining field trials are currently undertaken on ultramafic outcrops in Spain (Pardo et al., 2018), Austria (Rosenkranz et al., 2019) and Greece (Echevarria et al., 2017), cropping the most studied Odontarrhena muralis sensu latu (syn. Alyssum murale) as well as local Ni hyperaccumulator species such as Odontarrhena serpyllifolia (formerly Alyssum serpyllifolium), Bornmuellera tymphaea and Bornmuellera emarginata (formerly Leptoplax emarginata). Other Ni hyperaccumulators as Odontarrhena betolonii (formerly Alyssum bertolonii) (Robinson et al., 1997b), Berkheya coddii (Robinson et al., 1997a) were tested in field experiments, while in tropical environments field trials are being conducted in South East Asia with Phyllanthus rufuschaneyi (formerly P. securinegioides) and Rinorea cf. bengalensis (Chaney et al., 2018; Nkrumah et al., 2019). Recently, phytomining studies targeting waste incineration bottom ashes (Rosenkranz et al., 2017a, 2017b) and automotive industry residues (Rue et al., 2018) investigated the capability of hyperaccumulator species to phytoextract Ni from these types of wastes. So far, the possibility of a Ni extraction from galvanic sludge into plant biomass for metal recovery has not been investigated. Contrary to natural soils, waste materials containing metals of interest seldom exhibit favourable conditions for plant estab­ lishment and growth (Rosenkranz et al., 2017a). Galvanic sludges generally have plant adverse characteristics such as high salinity, extreme acidity or alkalinity, absence of structure, lack of essential plant nutrients and microbial activity. Moreover, these waste materials often contain mixtures of metallic elements, which can inhibit root growth and impair plant development (Rue et al., 2018). The formulation of appropriate substrates prior to plant cultivation is an essential step in order to improve structure, provide a source of organic matter and a balanced nutrient supply. The aim of the present study was to test the Ni hyperaccumulator plant Odontarrhena chalcidica for cultivation on sub­ strates derived from Ni-rich galvanic sludge, to investigate the feasibility of phytomining application on this type of toxic waste. Different waste materials were obtained from industries operating in Ni plating pro­ cesses and artificial substrates (i.e. technosols) were formulated to support plant growth. This study was conducted within the scope of the LIFE-Agromine Project (Cropping hyperaccumulator plants on nickel rich soils and wastes for the green synthesis of pure nickel compounds, http: //life-agromine.com).

2. Materials and methods 2.1. Characterization of waste materials Seven types of Ni-rich galvanic sludges (C1, C2, W1, W2, W3, W4 and W5) were obtained from different electroplating industries. After collection, the sludges were manually homogenised, stored in plastic containers for two months to allow equilibration and later characterised. All the analyses were performed on a subsample of the stabilized sludges with no previous exsiccation. Water content was measured on 10 g of each material dried at 105 � C and all the results were converted to dry weight. The pH was measured in ultra-pure water (w/v ratio 1:2.5) with a pH-meter at collection time and after two months of stabilization. Pseudo-total metal concentrations were determined after digestion in HCl 37% and HNO3 65% at 1:3 ratio (aqua regia) in a digestion block at 155 � C for 3 h. DTPA (diethylene triamine pentaacetic acid) (Lindsay and Norvell, 1978) extractions were performed to assess plant available metal fractions.1 Digested samples and DTPA extracts were determined for macronutrients and metal concentrations through ICP-MS (Elan 9000 DRCe, Perkin Elmer) and ICP-OES (Optima 8300, Perkin Elmer). Besides the galvanic sludges, a discarded serpentinite sand (Q) from gravel production was collected from a quarry located in the region of Burgenland (Austria), in correspondence with one of the main ultra­ mafic outcrops in Austria (Wenzel and Jockwer, 1999). This material was characterised for pH, pseudo-total metal concentrations and DTPA extractable metals following the methods previously described. 2.2. Formulation of waste-derived substrates Following the waste characterization, three galvanic sludges which were considered potentially suitable for the formulation of artificial soils were selected. The waste-derived soils constructed in the context of this study were defined as “technosols”.2 The selection criterion applied on the galvanic sludges was based on the Zn:Ni and Cu:Ni ratios of total and DTPA extractable fractions. To avoid the occurrence of Cu toxicity in the final substrates and reduce possible Zn interference with Ni uptake (Rue et al., 2018), only waste materials characterised by a Zn:Ni and Cu:Ni ratio lower than 0.25 were considered to be suitable. Due to its ultra­ mafic characteristics, the serpentinite sand Q was used as a mineral matrix in the formulation of the technosols. This, to improve the texture and mimic the ultramafic conditions typical of Ni hyperaccumulator plant habitats. Compost provided by the Waste Management Depart­ ment of the City of Vienna and obtained from municipal organic waste was employed as a source of organic matter. For each of the three selected galvanic sludges, eight types of technosols were prepared, mixing the sand Q with different % of galvanic sludge on a wet weight basis (1, 5, 10 or 50% w/w fresh sludge) as well as with 1 or 2% w/w of compost, and 24 substrates were obtained. Detailed technosol formu­ lations are shown in Table 1. Letters A to H (Table 1) were assigned to the final code of each technosol to indicate the % of sludge and compost. All the substrates were then homogenised by hand, lixiviated with deionized water in order to remove excess salts, air dried, stored in closed plastic containers and left to equilibrate for two months before analysis. On these formulations, pH and DTPA extractable metals were determined as previously described. To identify the waste-derived technosols which would better sustain Ni phytomining, the DTPA 1

DTPA assessments were developed for application on soils and their effec­ tiveness as indicators of metal bioavailability for sludge-type substrates is debated, due to the limited structure and exchangeable sites of those materials. Nevertheless, we have applied DTPA extractions in order to allow comparisons with soil DTPA data of our study and from literature. 2 technosols with lower case “t”, not be confused with soils originated from pedogenesis of artificial materials, thus belonging to the group of Technosols (uppercase “t”). 2

A. Tognacchini et al.

Journal of Environmental Management 254 (2020) 109798

Table 1 Formulation of the initial 24 technosols from 3 galvanic sludges and corresponding codes assigned. In light grey: technosols selected according to the DTPA values; in dark grey: the final 5 technosols selected according to the DTPA values and IG% test.

extractable fractions of each material were considered as a first selection criterion. DTPA extractions are commonly employed in agromining evaluations to assess metal plant availability. According to Echevarria et al. (1998) and Massoura et al. (2004), no chemical extraction method can accurately predict Ni uptake by hyperaccumulator plants, thus this extraction only provides an indication of the labile Ni pool to which hyperaccumulator plants are exposed. Technosols characterised by the lowest Ni-DTPA extractable fraction (Ni-DTPA below 40 mg kg 1) were discarded to avoid limitation in Ni uptake by plants. Zn:Ni and Cu:Ni ratios were again verified following the same criterion as for the pure sludge. As reported by Wang et al. (2004), when assessing metal toxicity, a poor correlation is often highlighted between metal bioavailability and plant toxicity. Thus, in metal toxicity assessments it is advisable to integrate chemical data of bioavailability with a biological indicator. A phytotoxicity test based on roots elongation and germination inhibition was therefore conducted on 19 (from initial 24) technosols selected after the DTPA and metal ratios methods. The test species Lolium perenne (commercially available seeds with guaranteed germination > 90%) was used, as well as the Ni hyperaccumulator species Odontarrhena chalcidica (seeds from the location of Pojsk€ e, Albania). Ten seeds of Lolium perenne and 20 seeds of Odontarrhena chalcidica were germinated on 100 mm Ø Petri dishes filled with 10 g of each substrate. A control on filter paper was included and two replicates were prepared for each substrate. After 15 days, germination rates, roots and shoots elongation were measured. The germination index (IG) was calculated according to Martignon (2009) and expressed as a percentage of the control (IG%): IG% ¼

and four replicates were included for each type of substrate, for a total of 48 pots. Seeds of O. chalcidica were germinated on filter paper and after two weeks ten seedlings were transplanted into the substrates. Controlled growth greenhouse conditions were set to 60% humidity, day/night cycle of 16/8 h, with day/night temperature of 25/15 � C. Pore water samplers were installed for monitoring Ni and micro­ nutrients along the experiment and sampling was performed two weeks after transplanting time (t ¼ 0) and at the end of the experiment (t ¼ 1). Plant biomass was harvested from each pot after 3 months. Plant shoots were cut and thoroughly rinsed with deionized water to remove possible traces of contamination originated from the growing substrates, then dried at 60 � C for 24 h. A dry biomass subsample from each replicate was digested in HNO3 65% and H2O2 30% in a 5:1 ratio in a digestion block at 155 � C for 4 h and then analysed for total element concentrations with ICP-MS (Elan 9000 DRCe, Perkin Elmer) and ICP-OES (Optima 8300, Perkin Elmer). Pore water samples were analysed for macronutrients with ICP-OES (Optima 8300, Perkin Elmer) and for trace elements with ICP-MS (Elan 9000 DRCe, Perkin Elmer). 2.4. Statistical analyses Differences in plant biomass and Ni concentration in O. chalcidica among all unfertilized and all fertilized substrates as well as elements concentrations in pore water samples were assessed through one-way ANOVA and Tukey’s HSD post-hoc test. If homogeneity of variance was not met, Welch ANOVA and the Games-Howell post-hoc tests were considered. If normal distribution was not met, the non-parametric Kruskal-Wallis H test with Dunn’s pairwise comparison and Bonferroni correction were applied. Differences in plant biomass and Ni concen­ trations between fertilized and unfertilized replicates within each sub­ strate were assessed through Student’s T-test for independent samples or through the non-parametric Man-Whitney U test when normal distri­ bution was not met. Correlations among variables were assessed through Pearson’s coefficient r. All statistical tests were performed with the program IBM SPSS statistics 24 considering a significance level of p < 0.05.

ðN � RÞ � 100 ðN control � R controlÞ

with N being the number of germinated seeds and R the average roots elongation. For each galvanic sludge (C1, W1, W2), the technosols resulting in lower phytotoxicity were selected for the phytomining pot experiment. 2.3. Phytomining pot experiment A greenhouse pot experiment was undertaken growing the Ni hyperaccumulator species Odontarrhena chalcidica on five technosols (selected as previously described and derived from the galvanic sludges C1, W1 and W2) and a control ultramafic soil (S). The ultramafic soil (S) was collected from an agricultural land located in the region of Bur­ genland (Austria) and analysed as previously described for the sludge and serpentinite sand (Q). Fertilized replicates (-F) were included in order to compare biomass and Ni yield in O. chalcidica, adding a 0.1% w/ w of Substral® Osmocote slow release garden fertilizer (NPK (Mg) in a 17:9:11:(2) ratio; trace elements: B, Cu, Mn, Fe, Mo and Zn) to the substrates. Two kg of each technosols and of the control soil were used

3. Results 3.1. Waste and ultramafic soil characteristics Based on the results of total and bioavailable metals, the three galvanic sludges C1, W1 and W2 were selected for the formulation of artificial substrates, while four sludges were discarded due to excessive Zn and Cu concentrations. A characterization of the selected industrial wastes C1, W1 and W2, the serpentinite sand Q and the control ultra­ mafic soil S is presented in Table 2. All three selected sludges resulted in 3

A. Tognacchini et al.

Journal of Environmental Management 254 (2020) 109798

Table 2 Pseudo-total concentration, DTPA extractable metals (average values � relative standard deviation %, n ¼ 2), water content and pH of selected galvanic sludges, serpentinite sand (Q) and ultramafic soil (S). < LOQ ¼ below the limit of quantification. LOQ

C1 1

mg kg a) Pseudo-total Ca Mg K Na Fe Al Cr Ni

0.428 0.017 0.142 0.081 0.206 0.345 0.706 0.827 1

0.534 0.140 0.008 0.009 0.065 0.005

μg kg

W1 1

g kg

212 � 22 5.15 � 2.3 1.27 � 6.6 8.45 � 3.4 2.56 � 15 0.573 � 5.7
mg kg Zn Cu Cd Co Pb Mn

g kg

mg kg

1


b) DTPA Al 1.57 Cr 0.314 Mn 0.109 Fe 11.4 Ni 0.368 Co 0.082 Cu 0.050 Zn 1.02 Cd 0.357 Pb 0.239 Water % fresh sludge pH initial pH after two months

mg kg

1

< LOQ < LOQ < LOQ < LOQ 3405 ± 6.2 < LOQ 1.33 � 46 < LOQ < LOQ < LOQ 45.4 10.1 8.9

W2 1

g kg

48.1 � 0.4 18.1 � 1.6 3.71 � 1.2 9.95 � 2.4 31.3 � 2.7 21.2 � 1.8 9.23 � 3.2 84.9 ± 1.3 mg kg

1

1100 � 5.3 367 � 0.4 < LOQ < LOQ 44.3 � 4.4 687 � 0.2 mg kg

1

< LOQ < LOQ 3.90 � 15 116 � 18 228 ± 20 < LOQ 6.89 � 6.2 41.5 � 14 < LOQ 2.82 � 1.3 64.9 8.2 8.2

highly elevated total Ni concentrations, ranging from 84.9 to 150 g kg 1 on a dry weight basis, while for the quarry sand Q and the control soil S the Ni concentration was respectively 2.63 and 1.45 g kg 1. Zn con­ centration was below the quantification limit for the waste C1, but much higher in the sludge W2 (1580 mg kg 1), while the highest total Cu and lead (Pb) concentrations were detected in waste C1 (Cu 932 mg kg 1, Pb 106 mg kg 1). Calcium (Ca) was more abundant than magnesium (Mg) in all three industrial wastes, while a higher sodium (Na) concentration was measured in the sludges C1 (8.45 g kg 1) and W1 (9.95 g kg 1). High Ca and Na concentrations occurred presumably due to the addition of calcium and sodium hydroxide (Ca(OH)2, NaOH) during the sludge precipitation process (Schugerl et al., 1996). In conformity with the total content, Ni DTPA (NiDTPA) extractable fractions (Table 2b) of all sludges resulted to be remarkably high, with extremely high values for the waste C1 (3405 mg kg 1). Despite the highest total Cu concentration in C1, its Cu bioavailable fraction (1.33 mg kg 1) resulted to be lower than in wastes W1 and W2, whereas the highest Zn availability was recorded for the W1 waste (41.5 mg kg 1). The quarry sand Q resulted in a typical ultramafic geochemical composition, and thus very similar to the con­ trol soil S. In fact, ultramafic soils are usually enriched in Mg, Fe, Mn, Cr, Ni, and Co (Echevarria, 2017). However, Ni, Fe, Mn and Cu resulted to be less available in the quarry sand Q than in the soil S, reflecting the different genesis of those ultramafic substrates. At collection time, all sludges resulted to be alkaline or very alkaline (C1 pH 10.1), with the pH of waste C1 being reduced to 8.9 after two months of stabilization.

Q 1

g kg

140 � 3.4 22.7 � 3.7 1.64 � 5.6 1.65 � 0.7 105 � 3.2 0.299 � 22 2.73 � 9.0 104 ± 1.0 mg kg

1

1580 � 2.6 648 � 4.2 < LOQ < LOQ 32.3 � 1.7 1024 � 3.8 mg kg

1

< LOQ < LOQ 1.08 � 4.7 < LOQ 290 ± 7.6 < LOQ 6.81 � 60 14.2 � 15 < LOQ < LOQ 64.4 8.8 8.6

S 1

g kg

11.6 � 2.4 211 � 12 0.042 � 38 – 47.3 � 7.6 13.2 � 8.7 1.88 � 4.9 2.63 ± 1.7 mg kg

1

107 � 25 37.2 � 41 < LOQ 99.5 � 11 < LOQ 909 � 3.4 mg kg

1

1

6.7 � 3.7 103 � 8.3 0.905 � 7.7 – 64.0 � 1.9 19.9 � 2.3 1.84 � 7.5 1.45 ± 13 mg kg

1

82.6 � 14 28.2 � 27 0.310 � 20 113.0 � 3.7 19.2 � 52 1308 � 6.8 mg kg

1

< LOQ < LOQ 0.722 � 4.0 16.5 � 0.2 15.5 ± 3.0 0.592 � 2.8 < LOQ < LOQ < LOQ < LOQ


8.1 8.1

6.1

commercial plant, the germination rate of Odontarrhena chalcidica was very low even in the control (below 30%); therefore, it was not possible to derive the IG% for this species. The germination rate of Lolium perenne in the control was 100% (Table 5, in supplementary materials). For each galvanic sludge, the formulations that resulted in the best germination index with Lolium perenne (Fig. 5, in supplementary materials) were selected for phytomining pot experiments: C1D (IG% 47.0), W1B (IG% 27.0) and W2H (IG% 88.5). The technosols C1F (IG% 15.3) and W2F (IG % 84.0) were also included to test substrates containing 10% industrial waste. Bioavailable metal fractions and pH of the selected artificial soils are presented in Table 3. After two months of stabilization the pH of all technosols was ranging between 8.2 and 8.6. Considering the percentage of galvanic sludge present in each technosol (see code letters in Table 1), it can be noticed that the NiDTPA increased in all substrates, except for W2H, when compared with the pure sludge. On the contrary, in the W2 sludge-derived substrates, a considerable increase in DTPA extractable Cu and Zn was observed as a consequence of technosol formulations. In accordance with the very high Ni availability in the C1 galvanic waste, the technosol C1F resulted in the highest DTPA extractable Ni concentration. 3.3. Pot experiment 3.3.1. Biomass O. chalcidica seedlings survived in all technosols and ultramafic soil replicates, with a higher survival rate in the control soil (data not shown). Shoots dry weight per pot is presented in Fig. 1. The maximum average dry shoots biomass obtained was 2.83 g per pot in the fertilized control soil. NPK fertilization resulted in a significant plant growth in­ crease in all the substrates, with a dry biomass 2.1–8.2-fold higher than in the unfertilized pots. Although O. chalcidica suffered from a slight

3.2. Technosols characteristics and selection Due to the discarding of five of the initial 24 waste-derived tech­ nosols based on the DTPA-extractable metal concentrations, the phyto­ toxicity test was performed on the remaining 19. Being a non4

A. Tognacchini et al.

Journal of Environmental Management 254 (2020) 109798

Table 3 DTPA extractable metals and pH of the selected technosols. Average values � relative standard deviation % (n ¼ 2). < LOQ ¼ below the limit of quantification. DTPA

LOQ

μg kg Al Cr Mn Fe Ni Co Cu Zn Cd Pb pH

2.03 0.33 0.02 11.29 1.68 0.01 0.08 14.44 0.02 0.08

C1D 1

mg kg

C1F 1


mg kg

W1B 1

mg kg


W2F 1


mg kg

W2H 1


mg kg

1


treatments, the maximum Ni accumulation in O. chalcidica was observed for the C1F technosol (26.8 g kg 1 Ni in C1F, 20.2 g kg 1 Ni in C1F-F), whereas the lowest Ni were obtained for the W2H technosols, respec­ tively 10.1 g kg 1 and 4.6 g kg 1. In both technosols and control soil, fertilization resulted in Ni concentration reduction from 24 to 53%, although for the substrate C1F it was not statistically significant. Contrarily, the amount of Ni extracted from each pot (i.e. Ni yield, Fig. 2b) resulted to be significantly higher in all the fertilized replicates due to the higher biomass. From the fertilized technosols C1F and W1B it was possible to extract respectively 36.5 and 30.9 mg of Ni per pot on average, which resulted to be comparable to the average Ni obtained from the fertilized ultramafic soil (37.1 mg pot 1). Cobalt (Co) (Fig. 3) uptake in the dry biomass of O. chalcidica resulted to be significantly higher than the control soil when growing on the C1D, W1B and W2F substrates. A remarkable Co concentration was recorded in the unfer­ tilized W1B replicates (346 mg kg 1), 20 times higher than in the un­ fertilized control. The highest Zn concentration in plant shoots (Table 4) was observed for W1, W2F and W2H, although the difference with the control S was not statistically significant for the fertilized replicates. The substrate W2H also resulted in a significantly higher Cu concentrations (Table 4) in the biomass: 3.7 (unfertilized) and 5.9 (fertilized) times higher than the Cu uptake in the controls. Significant (Pearson’s r; p < 0.05) linear strong positive correlations resulted among Zn (r ¼ 0.849 noF, r ¼ 0.795 F), Cu (r ¼ 0.987 noF, r ¼ 0.979 F) and Co (r ¼ 0.933 noF, r ¼ 0.877 F) DTPA extractable fractions and shoots con­ centration in O. chalcidica, while no significant correlation between Ni plant uptake Ni-DTPA was detected. Furthermore, no significant nega­ tive correlation was observed among Zn, Cu, Mn shoot concentrations and Ni plant uptake.

Fig. 1. Shoot dry weight of Odontarrhena chalcidica per pot. Average values � standard deviation (n ¼ 4). Statistical difference (p < 0.05) among substrates is indicated with different letters, lowercase for unfertilized, upper­ case for fertilized replicates.*indicates statistical difference (p < 0.05) between fertilized and unfertilized treatments. C1D, C1F, W1B, W2F and W2H ¼ technosols, S ¼ control ultramafic soil.

3.3.3. Pore water Fig. 4 shows the Ni pore water concentrations (Nirhiz) of unfertilized (Fig. 4a) and fertilized (Fig. 4b) pots. The substrate C1F resulted in the pore water with the highest concentration of Ni, considerably higher than all other technosols and control soil. Copper, Zn, Co, Al, Cr, Cd and Pb resulted to be under the quantification limit (Table 5 in supple­ mentary material) in all substrates. Sodium (Na) pore water concen­ trations (Table 5 in supplementary material) were considerably higher in C1F and W2H technosols compared to all other substrates. Calcium, Mg and K were generally more abundant in the artificial soils than in the control S, while Fe pore water concentrations resulted to be comparable in all substrates. In general, an increase of Ni concentration in the pore water was more frequently observed in the fertilized treatments (see Fig. 4).

growth inhibition in all the fertilized constructed soils, this reduction resulted to be significant only for the W2H substrate. Contrarily, in all unfertilized technosols a significant biomass reduction was observed in comparison to the unfertilized control soil S. Correlations (Pearson’s r, p < 0.05) among plant and substrate variables showed a significant positive correlation between plant biomass and MnDTPA (r ¼ 0.911) in the unfertilized replicates and between plant biomass and Cu concen­ trations in O. chalcidica (r ¼ 0.931) in the fertilized pots. 3.3.2. Nickel and other elements in shoots Nickel concentration in the dry biomass of O. chalcidica is presented in Fig. 2a. Nickel uptake in O. chalcidica dry shoots from the ultramafic soil S resulted to be in average 20.2 g kg 1 Ni in the unfertilized repli­ cates and 14.0 g kg 1 Ni when fertilization was applied. Average Nishoot accumulation from the unfertilized C1F and W1B substrates resulted to be higher than the control S, with only W1B being statisti­ cally validated. Among fertilized replicates, the Ni plant uptake of C1D, C1F, W1B, and W2F resulted to be statistically comparable with the control S. Nevertheless, C1D and W2F average Ni shoots concentration was lower than on the ultramafic soil. In the fertilized and unfertilized

4. Discussion 4.1. Galvanic sludge and technosols characterization and selection The selection method applied on the technosols resulted to be only 5

A. Tognacchini et al.

Journal of Environmental Management 254 (2020) 109798

Fig. 2. a) Nickel shoots concentrations (mg kg 1) and b) nickel yield per pot (mg pot 1) of Odontarrhena chalcidica. Average values � standard deviation (n ¼ 4). Statistical difference (p < 0.05) among substrates is indicated with different letters, lowercase for unfertilized replicates, uppercase for fertilized replicates.* indicates statistical difference (p < 0.05) between fertilized and unfertilized treatments. C1D, C1F, W1B, W2F and W2H ¼ technosols, S ¼ control ultramafic soil.

growing on the W2H substrate. Being Cu concentrations in pore waters under the limit of quantifi­ cation (Table 5 in supplementary material), it is likely that roots of L. perenne were not affected by Cu during the germination test which was conducted for a limited time (two weeks). On the contrary, roots of O. chalcidica, which grew over three months on the substrates, could have been affected by the high Cu levels detected in the exchangeable and DTPA extractable Cu, presumably due to root activities leading to Cu mobilization. Also, L. perenne seems to have shown toxicity responses mainly to those elements highly present in rhizosphere water such as Ni (Fig. 4), whereas O. chalcidica would tolerate them. Hence, phytotoxicity assessments with commercial test species are not recommendable for phytomining applications and it is crucial that hyperaccumulator plants should be used as the target organisms. 4.2. Plant biomass, nickel uptake and other metals The present study shows that the Ni hyperaccumulator species O. chalcidica was capable of viably growing on toxic galvanic sludgederived technosols and to produce a comparable biomass than on nat­ ural ultramafic soils (Fig. 1). Growth inhibition in all the unfertilized technosols seems to be related both to limited nutrients supply and toxicity responses to metals. For instance, a strong positive correlation among biomass of the unfertilized plants and Mn availability (MnDTPA) was recorded (Pearson’s r, p < 0.05). When fertilization was applied, biomass reduction strongly correlated with increasing Cu concentrations in shoots of O. chalcidica, suggesting Cu toxicity as the main factor inhibiting plant growth in the W2H-F replicates. Borkert et al. (1998) reported toxic effects in cultivars accumulating Cu levels exceeding 20 mg kg 1 in their shoots. The Cu concentrations in shoots of O. chalcidica (18.4 mg kg 1 in W2H, 23.8 mg kg 1 in W2H-F; Table 4a and b), indicate that Cu toxicity might have occurred in the W2H sub­ strate. Biomass reduction is reported as a common Cu phytotoxicity symptom in literature (Ebbs and Kochian, 1997; Pietrzak and Uren, 2011; Lange, 2016). Kidd and Monterroso (2005) and Rosenkranz et al. (2017a) also observed a reduced biomass due to Cu phytotoxicity in the Ni hyperaccumulator Odontarrhena serpyllifolia. The technosol W2H can be thus considered as non-suitable for phytomining applications due to high Cu availability. Compared to the usual size of O. chalcidica under field conditions, a low biomass production was observed also on the

Fig. 3. Cobalt shoots concentrations (mg kg 1) of Odontarrhena chalcidica. Average values � standard deviation (n ¼ 4). Statistical difference (p < 0.05) among substrates is indicated with different letters, lowercase for unfertilized replicates, uppercase for fertilized replicates. * indicates statistical difference (p < 0.05) between fertilized and unfertilized treatments. C1D, C1F, W1B, W2F and W2H ¼ technosols, S ¼ control ultramafic soil. control ultramafic soil.

partially effective in terms of plant adaptation and toxicity response, also due to the phytotoxicity test failure with O. chalcidica. The most encouraging phytomining results were obtained from the C1F technosol, which would have been excluded according to the phytotoxicity test results with Lolium perenne. On the contrary, the very high germination index (IG%) obtained with Lolium perenne would have suggested the W2H technosol as a low toxicity matrix, which instead appeared to be unfavourable for O. chalcidica. Limitations of this selection method could be due to the different responses that test and hyperaccumulator plants showed to the tested substrates. A critical point was the absence of toxicity response to enhanced Cu concentrations by L. perenne when 6

A. Tognacchini et al.

Journal of Environmental Management 254 (2020) 109798

Table 4 Concentration of other elements in shoots of Odontarrhena chalcidica in a) unfertilized and b) fertilized replicates. Average values � relative standard deviation % (n ¼ 4). Statistical difference (p < 0.05) in element content among substrates is indicated with different letters. -F indicates the fertilized replicates. a) Unfertilized Mg g kg 1 K g kg 1 Ca g kg 1 Mn mg kg Fe mg kg Zn mg kg Al mg kg Cu mg kg b) Fertilized Mg K Ca Mn Fe Zn Al Cu

g kg 1 g kg 1 g kg 1 mg kg mg kg mg kg mg kg mg kg

1 1 1 1 1

1 1 1 1 1

C1D

C1F

W1B

W2F

W2H

S

6.14 � 11 a 17.3 � 11 b 32.8 � 5.4 ab 149 � 22 a 261 � 6.2 ab 33.9 � 21 c 27.8 � 65 a 4.44 � 61 b C1D-F

2.69 � 27 b 23.6 � 8.0 a 33.6 � 26 ab 40.3 � 7.5 c 255 � 20 ab 22.4 � 50 c 19.4 � 36 a 3.88 � 28 b C1F-F

15.8 � 27 a 11.1 � 13 c 39.9 � 6.5 a 97.9 � 6.0 b 268 � 14 a 134 � 7.3 b 23.1 � 64 a 3.47 � 14 b W1B-F

9.35 � 22 a 15.1 � 17 bc 31.0 � 14 ab 60.6 � 14 c 206 � 10 ab 192 � 12 a 26.4 � 24 a 9.05 � 39 b W2F-F

7.01 � 21 a 15.3 � 26 bc 24.9 � 14 b 52.0 � 36 bc 164 � 23 b 193 � 19 a 27.6 � 54 a 18.4 � 32 a W2H-F

2.10 � 22 c 10.9 � 14 c 33.0 � 13 ab 58.9 � 16 c 189 � 10 ab 60.3 � 28 c 16.6 � 24 a 4.97 � 102 b S-F

4.79 � 21 bc 18.3 � 8.8 b 31.6 � 9.9 ab 140 � 8.4 a 227 � 12 a 29.1 � 12 ab 11.2 � 51 b 5.35 � 11 ab

2.69 � 15 c 22.1 � 15 ab 33.2 � 6.5 ab 36.9 � 13 d 201 � 6.4 a 22.5 � 27 b 12.5 � 57 b 4.41 � 33 ab

10.7 � 10 a 17.1 � 33 ab 36.9 � 20 b 100 � 13 b 201 � 9.0 a 83.1 � 26 a 12.4 � 16 b 4.83 � 16 ab

5.06 � 14 b 17.1 � 8.8 b 28.5 � 6.8 ab 51.3 � 18 d 185 � 11 a 113 � 5.0 a 19.2 � 34 ab 8.64 � 19 a

6.87 � 12 b 25.5 � 9.0 a 26.5 � 13 a 38.5 � 14 d 177 � 8.3 a 104 � 15 a 13.5 � 30 ab 23.8 � 23 c

2.39 � 12 c 14.7 � 12 b 30.5 � 3.8 ab 74.7 � 8.4 c 202 � 22 a 48.2 � 24 ab 26.9 � 31 a 4.04 � 36 b

Fig. 4. Nickel pore water concentration at the beginning (T ¼ 0) and at the end (T ¼ 1) of the pot experiment, a) unfertilized and b) fertilized replicates. Average values � standard deviation (n ¼ 4). Significant difference (p < 0.05) among substrates is indicated with different letters, lowercase for unfertilized replicates, up­ percase for fertilized replicates.* indicates statistical difference (p < 0.05) between fertilized and unfertilized treatments. C1D, C1F, W1B, W2F and W2H ¼ tech­ nosols, S ¼ control ultramafic soil.

control ultramafic soil S, probably due to general growing settings or to the high number of seedlings per pot. Mineral NPK fertilization resulted to be necessary in order to achieve reasonable amounts of biomass, although a significant dilution effect in Ni shoot concentrations occurred in the ultramafic soil and in all the technosols except C1F. Although less pronounced than in our results, a Ni dilution trend was also observed by Robinson et al. (1997b) in a field experiment with Odontarrhena berto­ � �pez et al. (2016) also lonii when N fertilization was applied. Alvarez-L o reported a decrease in shoot Ni concentrations in compost-amended soils. However, in an ultramafic soil with very high DTPA-extractable Ni concentrations (Bani et al., 2015a; Bani and Echevarria, 2019), fertilization treatments increased both biomass production and Ni

concentrations in shoots thus resulting in much higher phytoextraction yields. The greatest Ni extraction from technosols was obtained on C1F in terms of either mg Ni kg 1 (Fig. 2a) and mg Ni pot 1 (Fig. 2b), followed by W1B. The highest average Ni shoot concentrations in C1F coincided with the highly available Ni (NiDTPA, Nirhiz) in the substrate. This could explain why no dilution effect of NPK fertilization on shoots Ni concentrations was observed for this treatment. However, a direct correlation among Ni availability and plant uptake was not found (Pearson’s r, p < 0.05). Higher but statistically comparable amounts of Ni yields were achieved with a NiDTPA of about 500 mg kg 1 (technosol C1F) and from the ultramafic soil (S) with a NiDTPA of 32 mg kg 1, suggesting that after a certain threshold, considerable increases in Ni 7

A. Tognacchini et al.

Journal of Environmental Management 254 (2020) 109798

bioavailability do not result in higher plant uptake. Li et al. (2003a) and Kukier et al. (2004) observed a higher Ni accumulation by Odontarrhena (synonymous Alyssum) species as soil pH was raised and thus DTPA extractable Ni declined. In this regard, no pH effect can be assessed as the different Ni availabilities in the substrates were not pH dependent but related to the technosols formulations and presumably to the Ni form in the galvanic sludges. Ni shoot content in O. chalcidica was considerably reduced in the W2H substrate (Fig. 2), where Zn and Cu resulted to be highly abundant in the biomass. Still, no significant in­ terferences of shoots Ni concentration with Zn and Cu uptake were measured (Pearson’s r, p < 0.05). In previous phytomining studies on waste-derived substrates (Rue et al., 2018) Ni accumulation in Odon­ tarrhena spp (syn. Alyssum murale) was inhibited due to competition with Zn. In our experiment, Zn concentration in O. chalcidica shoots resulted to be substantially lower than Ni, preventing interactions also in terms of nickel recovery processes. Cobalt extraction from the unfertilized tech­ nosol W1B resulted to be unexpectedly high (>300 mg Co kg 1, Fig. 3). Hyperaccumulation of cobalt was initially delimited to a dry shoot concentration above 1000 mg Co kg 1 (Brooks et al., 1980), and was recently revised to a new threshold of 300 mg Co kg 1 (van der Ent et al., 2013b). Thus, our results showed a Co hyperaccumulation behaviour in O. chalcidica when growing on the W1B technosol, despite the very low cobalt availability in the substrate (0.97 mg kg 1 CoDTPA, Table 3). Malik et al. (2000) observed Co shoots concentrations above 1000 mg kg 1 in Odontarrhena muralis s.l. and Odontharrena corsica (syn. Alyssum corsi­ cum) growing on a nickel poor soil, while Li et al. (2003a) reported maximum shoots Co uptake of about 100 mg kg 1 for the same plant species. Previous studies also reported Co accumulation to inhibit Ni uptake (Homer et al., 1991; Malik et al., 2000; Li et al., 2003a), which seems not to have occurred in our experiment, where the Co available fraction (CoDTPA) was about 46 times lower than Ni (46.3 mg kg 1 NiDTPA, Table 3).

results. Previous phytomining studies undertaken on waste-derived substrates show that Ni extraction by hyperaccumulator plants was limited. In a pioneering phytomining experiment conducted on incin­ eration bottom ashes (Rosenkranz et al., 2017a), 249 mg Ni kg 1 were extracted with Odontarrhena serpyllifolia, but the substrate did not sup­ port healthy plant growth and the biomass production was low. Rue et al. (2018) observed a satisfactory biomass production of Odontarrhena spp (syn. Alyssum murale) on automotive industry waste-derived sub­ strates, but Ni extraction was limited to 0.1%. Despite the encouraging Ni extraction rates (mg kg 1) obtained with O. chalcidica from the C1 and W1 sludges, it is unquestionable that an increase in biomass pro­ duction would be necessary for economical phytomining. Furthermore, the convenience of a metal bioconcentration and purification into plant biomass seems in this case to be limited due to the high Ni concentra­ tions of the pure sludge. Considering only the Ni fraction deriving from the industrial sludges in each pot, with one phytomining cycle up to a 0.22% from the technosol C1F and up to a 1.04% from W1B were extracted. In principle, a number of phytomining cycles of about 450 (C1F) and 96 (W1B) would be then necessary in order to recover all the Ni present in the pot, assuming a constant Ni extraction rate and that the Ni available fraction would be resupplied. In a pilot-scale experiment, � et al. (2003) successfully recovered up to 99% of metallic Ni from Veglio galvanic waste containing 9–13% Ni with a leaching and electrowinning process. Nevertheless, it has to be considered that in most cases Ni is not being recovered from galvanic sludge and that conventional pyro/hy­ drometallurgical recovery is not applicable on waste containing a lower Ni wt %. As pointed out by van der Ent et al. (2013a), phytomining may target low-grade ores that cannot be exploited profitably. In the case of galvanic sludge, phytomining could provide an alternative way of metal recycling when other methods are not economically feasible. Assuming that a dry shoot Ni content of 1–2% is further concentrated to 10–20% after plants incineration (Simonnot et al., 2018), the advantage of a phytomining approach could be the substantial bioaccumulation factor between lower-grade galvanic sludge and plant ashes. A further important aspect is the possibility to produce pure Ni-salts from the harvested biomass of Odontarrhena spp. (syn. Alyssum murale) (Barbar­ oux et al., 2012), which has a potentially higher economic value than metallic Ni. Furthermore, applying phytomining on Ni-rich waste could provide the following benefits: 1) reducing the waste to a lower grade of contamination, with consequent lower costs of landfilling; 2) creating value through Ni recycling.

4.3. Considerations about Ni phytomining Remarkably, on all the waste-derived substrates Odontarrhena chal­ cidica achieved Ni shoots concentrations of more than >1000 mg Ni kg 1 , which is defined as the Ni hyperaccumulation threshold (Brooks et al., 1977; Baker et al., 2000). Furthermore, a “hypernickelphore” behaviour was observed in O. chalcidica, as it was possible to obtain a Ni concentration above 1% from all the substrates except W2H-F and C1D-F (slightly below 1%) and even above 2% (C1F). A Ni shoots accumulation above 1% is considered to be suitable for commercially valuable phy­ tomining applications (Chaney et al., 2007; van der Ent et al., 2013b). Examples of Ni shoots concentrations in Odontarrhena muralis s.l. from field experiments on ultramafic soils can be found in literature: 11.5 g Ni kg 1 (Bani et al., 2015a), 12.4 g Ni kg 1 (Rosenkranz et al., 2019), 15.0 g Ni kg 1 (Li et al., 2003b). Thus, the Ni extraction obtained from the technosols C1F, W1B and W2F can be considered comparable not only with the control soil S but with commonly reported accumulation rates. Considering an optimal density for Odontarrhena muralis s.l of one plant per 0.25 m2 (Li et al., 2003b; Bani et al., 2015b), an amount of 1.46 kg Ni and 1.23 kg Ni per hectare could be recovered respectively from the technosols C1F-F and W1B-F with a hypothetical upscaling. From the so far largest European agromining field experiment in Albania, Bani et al. (2015a) obtained yields of 105 kg Ni per hectare with Odontarrhena chalcidica, while yields of about 200 kg Ni ha 1 were estimated under intensive agricultural conditions by Nkrumah et al. (2016). Recent agromining trials in tropical regions suggest that a Ni yield of 200–300 kg ha 1 could be achieved with the hyperaccumulators Phyl­ lanthus rufuschaneiy and Rinorea c. f bengalensis (Chaney et al., 2018; Nkrumah et al., 2019). Nevertheless, when comparing with field results it should be considered that in our study O. chalcidica was cultivated for 3 months on 2 kg pots in greenhouse conditions; therefore, proper esti­ mations of Ni yields per hectares obtained from technosols should be assessed with upscaled trials, in order to allow comparisons with field

5. Conclusions This study represents the first experimental phytomining trial con­ ducted on industrial galvanic sludges. We have demonstrated the feasibility of transferring Ni from toxic waste into the shoots of Odon­ tarrhena chalcidica at levels that would allow an economically feasible Ni recovery from the biomass. Our results show that phytomining from galvanic sludge-derived technosols can provide the same Ni yields as from ultramafic natural soils. Nevertheless, a limited plant growth was obtained from the pot experiment and further research should focus on upscaling and ameliorating plant biomass in order to achieve higher Ni yields. Phytomining could, in principle, represent a novel plant-based approach for recycling Ni from galvanic waste, but it should target sludges with low Ni concentrations, that do not allow other costeffective physic-chemical extraction methods to be implemented for the recovery of Ni. Acknowledgements The present study was conducted as part of the LIFE-Agromine Project (LIFE15 ENV/FR/000512) funded by the European Commis­ sion. We acknowledge the institute alchemia-nova GmbH for supporting part of the PhD of Alice Tognacchini.

8

A. Tognacchini et al.

Journal of Environmental Management 254 (2020) 109798

Appendix A. Supplementary data

nickel: economic and technical considerations. Plant Soil 249, 107–115. https://doi. org/10.1023/A:1022527330401. Lindsay, W.L., Norvell, W.A., 1978. Development of a DTPA soil test for zinc, iron, manganese and copper. Soil Sci. Soc. Am. J. 41, 421–428. https://doi:10.2136/sss aj1978.03615995004200030009x. Malik, M., Chaney, R.L., Brewer, E.P., Li, Y.M., Angle, J.S., 2000. Phytoextraction of soil cobalt using hyperaccumulator plants. Int. J. Phytoremediation 2, 319–329. Martignon, G., 2009. Linee guida per la misura della tossicit� a dei suoli. In: CESI Ricerca, S.p.A. (Ed.), Test di fitotossicit� a per il suolo. Rapporto 09000808, p. 148. Milano. Massoura, S.T., Echevarria, G., Leclerc-Cessac, E., Morel, J.L., 2004. Response of excluder, indicator and hyperaccumulator plants to nickel availability in soils. Aust. J. Soil Res. 42, 933–938. Morel, J.L., 2013. Using plants to “Micro-mine” metal [WWW Document]. http://www. inra.fr/en/en/Scientists-Students/Biomass/All-the-news/Using-plants-tomicro-mi ne-metals. Morel, J.L., 2015. Agromining: a new concept. In: Echevarria, G., Morel, J.L., Simonnot, M.O. (Eds.), Leaders) Workshop 3 “Agromining: from Soils to Refined Metal Products”. SGA Biennial Meeting. Nancy, France, 24–27 Aug 2015. Mudd, G.M., 2009. Nickel sulfide versus laterite : the hard sustainability challenge remains. In: Proc. “48th Annual Conference of Metallurgists”. Canadian Metallurgical Society, Sudbury, Ontario, Canada. August 2009. Nkrumah, P.N., Baker, A.J.M., Chaney, R.L., Erskine, P.D., Echevarria, G., Morel, J.L., van der Ent, A., 2016. Current status and challenges in developing nickel phytomining: an agronomic perspective. Plant Soil 406, 55–69. https://doi.org/ 10.1007/s11104-016-2859-4. Nkrumah, P.N., Tisserand, R., Chaney, R.L., Baker, A.J.M., Louis, J., Goudon, R., Erskine, P.D., Echevarria, G., van der Ent, A., 2019. The first tropical ‘metal farm’: some perspectives from field and pot experiments. J. Geochem. Explor. 198, 114–122. https://doi.org/10.1016/j.gexplo.2018.12.003. Pardo, T., Rodríguez-Garrido, B., Saad, R.F., Soto-V� azquez, J.L., Loureiro-Vi~ nas, M., � Echevarria, G., Benizri, E., Kidd, P.S., 2018. Assessing the Prieto-Fern� andez, A., agromining potential of Mediterranean nickel-hyperaccumulating plant species at field-scale in ultramafic soils under humid-temperate climate. Sci. Total Environ. 630, 275–286. https://doi.org/10.1016/j.scitotenv.2018.02.229. Pietrzak, U., Uren, N.C., 2011. Remedial options for copper-contaminated vineyard soils. Soil Res. 49 (1), 44–55. https://doi.org/10.1071/SR09200. Rajcevic, H.P., 1990. A hydrometallurgical process for the treatment of industrial wastes. Plant. Surf. Finish. 22–25. Reeves, R.D., Baker, A.J.M., Borhidi, A., Berazaín, R., 1996. Nickel-accumulating plants from the ancient serpentine soils of Cuba. New Phytol. 133, 217–224. Robinson, B.H., Brooks, R.R., Howes, A.W., Kirkman, J.H., Gregg, P.E.H., 1997. The potential of the high-biomass nickel hyperaccumulator Berkheya coddii for phytoremediation and phytomining. J. Geochem. Explor. 60, 115–126. https://doi. org/10.1016/S0375-6742(97)00036-8. Robinson, B.H., Chiarucci, A., Brooks, R.R., Petit, D., Kirkman, J.H., Gregg, P.E.H., De Dominicis, V., 1997. The nickel hyperaccumulator Alyssum bertolonii as a potential agent for phytoremediation and phytomining of nickel. J. Geochem. Explor. 59, 75–86. https://doi.org/10.1016/S0375-6742(97)00010-1. Rosenkranz, T., Kisser, J., Wenzel, W.W., Puschenreiter, M., 2017. Waste or substrate for metal hyperaccumulating plants – the potential of phytomining on waste incineration bottom ash. Sci. Total Environ. 575, 910–918. https://doi.org/10.1016/ j.scitotenv.2016.09.144. Rosenkranz, T., Kidd, P., Puschenreiter, M., 2017. Effect of bacterial inoculants on phytomining of metals from waste incineration bottom ash. Waste Manag. 73, 351–359. https://doi.org/10.1016/j.wasman.2017.12.006. Rosenkranz, T., Hipfinger, C., Ridard, C., Puschenreiter, M., 2019. A nickel phytomining trial using Odontarrhena chalcidica and Noccaea goesingensis on an Austrian serpentine soil. J. Environ. Manag. 242, 522–528. https://doi.org/10.1016/j. envman.2019.04.073. Rossini, G., Bernardes, A.M., 2006. Galvanic sludge metals recovery by pyrometallurgical and hydrometallurgical treatment. J. Hazard Mater. 131, 210–216. https://doi.org/ 10.1016/j.jhazmat.2005.09.035. Rue, M., Rees, F., Simonnot, M.-O., Morel, J.-L., 2018. Phytoextraction of Ni from a toxic industrial sludge amended with biochar. J. Geochem. Explor. 196, 173–181. https:// doi.org/10.1016/j.gexplo.2018.10.007. Schugerl, K., Burmaster, T., Gudorf, M., 1996. Selective extractive recovery of metals from heavy-metal-hydroxy sludges of a pickling plant and galvanic processing. In: Shallcross, D., Paimin, R., Prvcic, L. (Eds.), ISEC 96, vol. 2. University of Melbourne Publisher, Parkville, Australia, pp. 1549–1552. Silva, J.E., Soares, D., Paiva, A.P., Labrincha, J.A., Castro, F., 2005. Leaching behaviour of a galvanic sludge in sulphuric acid and ammoniacal media. J. Hazard Mater. B121, 195–202. Simonnot, M.-O., Vaughan, J., Laubie, B., 2018. Processing of bio-ore to products. In: van der Ent, A., Echevarria, G., Baker, A.J.M., Morel, J.L. (Eds.), Agromining: Farming for Metals: Extracting Unconventional Resources Using Plants. Springer International Publishing, Cham, pp. 39–52. � Spaniel, S., Kempa, M., Salmer� on-S� anchez, E., Fuertes-Aguilar, J., Francisco Mota, J., Al� Shehbaz, I.A., Dmitry, A., Osl�savsk� a, K., Singliarov� a, B., Zozomov� a-Lihov� a, J., Marhold, K., 2015. AlyBase – database of names, chromosome numbers, and ploidy levels of Alysseae (Brassicaceae), with a new generic concept of the tribe. Plant Syst. Evol. 301, 2463–2491. https://doi:10.1007/s00606-015-1257-3. van der Ent, A., Baker, A.J.M., van Balgooy, M.M.J., Tjoa, A., 2013. Ultramafic nickel laterites in Indonesia (Sulawesi, Halmahera): mining, nickel hyperaccumulators and opportunities for phytomining. J. Geochem. Explor. 128, 72–79. https://doi.org/ 10.1016/j.gexplo.2013.01.009.

Supplementary data to this article can be found online at https://doi. org/10.1016/j.jenvman.2019.109798. References � � Cabello-Conejo, M.I., Kidd, P.S., 2016. Organic Alvarez-L� opez, V., Prieto-Fern� andez, A., amendments for improving biomass production and metal yield of Nihyperaccumulating plants. Sci. Total Environ. 548–549, 370–379. https://doi.org/ 10.1016/j.scitotenv.2015.12.147. Bani, A., Echevarria, G., 2019. Can organic amendments replace chemical fertilizers in nickel agromining cropping systems in Albania? Int. J. Phytoremediation 21, 43–51. https://doi.org/10.1080/15226514.2018.1523871. Bani, A., Echevarria, G., Sulçe, S., Morel, J.L., 2015. Improving the agronomy of Alyssum murale for extensive phytomining: a five-year field study. Int. J. Phytoremediation 17, 117–127. https://doi.org/10.1080/15226514.2013.862204. Bani, A., Echevarria, G., Zhang, X., Benizri, E., Laubie, B., Morel, J.L., Simonnot, M.-O., 2015. The effect of plant density in nickel-phytomining field experiments with Alyssum murale in Albania. Aust. J. Bot. 63, 72–77. https://doi.org/10.1071/ BT14285. Brooks, R.R., Lee, J., Reeves, R.D., Jaffre, T., 1977. Detection of nickeliferous rocks by analysis of herbarium specimens of indicator plants. J. Geochem. Explor. 7, 49–57. https://doi.org/10.1016/0375-6742(77)90074-7. Baker, A.J.M., McGrath, S.P., Reeves, R.D., Smith, J.A.C., 2000. Metal hyperaccumulator plants: a review of ecology and physiology of a biological resource for phytoremediation of metal-polluted soils. In: Terry, N., et al. (Eds.), Phytoremediation of Contaminated Soil and Water. Lewis Publishers, Boca Raton, FL, pp. 129–158. Barbaroux, R., Plasari, E., Mercier, G., Simonnot, M.-O., Morel, J.L., Blais, J.F., 2012. A new process for nickel ammonium disulfate production from ash of the hyperaccumulating plant Alyssum murale. Sci. Total Environ. 423, 111–119. https:// doi.org/10.1016/j.scitotenv.2012.01.063. Bernardes, A.M., Bohlinger, I., Wuth, W., 1996. Thermal treatment of galvanic sludges for environmental compatibility. JOM 59–62. March 1996. Brooks, R.R., Reeves, R.D., Morrison, R.S., Malaisse, F., 1980. Hyperaccumulation of copper and cobalt—a review. Bull. Soc. R. Bot. Belg. 113, 166–172. Borkert, C.M., Cox, F.R., Tucker, M.R., 1998. Zinc and copper toxicity in peanut, soybean, rice and corn in soil mixtures. Commun. Soil Sci. Plant Anal. 29, 2991e3005. https://doi.org/10.1080/00103629809370171. Chaney, R.L., 1983. Plant uptake of inorganic waste constituents. In: Parr, J.F., Marsh, P. B., Kla, J.M. (Eds.), Land Treatment of Hazardous Wastes. Noyes Data Corp, Park Ridge, NJ, pp. 50–76. Chaney, R.L., Angle, J.S., Baker, A.J.M., Li Y.-M., 1998. Method for phytomining of nickel, cobalt and other metals from soil. US Patent 1998, 5, 711,784. Chaney, R.L., Angle, J.S., Broadhurst, C.L., Peters, C.A., Tappero, R.V., Sparks, D.L., 2007. Improved understanding of hyperaccumulation yields commercial phytoextraction and phytomining technologies. J. Environ. Qual. 36, 1429–1443. htt ps://doi:10.2134/jeq2006.0514. Chaney, R.L., Baker, A.J.M., Morel, J.L., 2018. The long road to developing agromining/ phytomining. In: Van der Ent, A., Echevarria, G., Baker, A.J.M., Morel, J.L. (Eds.), Agromining: Farming for Metals: Extracting Unconventional Resources Using Plants. Springer International Publishing, Cham, pp. 1–17. Ebbs, S.D., Kochian, L.V., 1997. Toxicity of zinc and copper to Brassica species: implications for phytoremediation. J. Environ. Qual. 26, 776–781. https://doi:10 .2134/jeq1997.00472425002600030026x. Echevarria, G., Morel, J.L., Fardeau, J.C., Leclerc-Cessac, E., 1998. Assessment of phytoavailability of nickel in soils. J. Environ. Qual. 27, 1064–1070. https://doi:10 .2134/jeq1998.00472425002700050011x. Echevarria, G., Bani, A., Benizri, E., Kidd, P.S., Kisser, J., Konstantinou, M., Kyrkas, D., Puschenreiter, M., Tognacchini, A., 2017. Life agromine: a European demonstration project for Ni agromining. In: Proceedings of the 14th International Phytotechnologies Conference, Session 4E - Part 1 - Bioremediation and Bioeconomy, pp. 25–29 (Montreal, QC). Echevarria, G., 2017. Genesis and behaviour of ultramafic soils and consequences for nickel biogeochemistry. In: van der Ent, A., Echevarria, G., Morel, J.L. (Eds.), Agromining: Farming for Metals. Springer. Homer, F.A., Morrison, R.S., Brooks, R.R., Clemens, J., Reeves, R.D., 1991. Comparative studies of nickel, cobalt, and copper uptake by some nickel hyperaccumulators of the genus Alyssum. Plant Soil 138, 195–205. https://doi.org/10.1007/BF00012246. Kidd, P.S., Monterroso, C., 2005. Metal extraction by Alyssum serpyllifolium ssp. Lusitanicum on mine-spoil soils from Spain. Sci. Total Environ. 336, 1–11. https:// doi.org/10.1016/j.scitotenv.2004.06.003. Kukier, U., Peters, C.A., Chaney, R.L., Angle, J.S., Roseberg, R.J., 2004. The effect of pH on metal accumulation in two Alyssum species. J. Environ. Qual. 32, 2090–2102. htt ps://doi:10.2134/jeq2004.2090. Lange, B., 2016. Tol� erance et accumulation du cuivre et du cobalt—implication pour la phytorem� ediation des sols contamin�es. Th� ese de doctorat, Universit� e Libre de Bruxelles, Universit�e Picardie Jules Verne, France, p. 160. Li, Y.-M., Chaney, R.L., Brewer, E.P., Angle, J.S., Nelkin, J., 2003. Phytoextraction of nickel and cobalt by hyperaccumulator Alyssum species grown on Ni-contaminate soils. Environ. Sci. Technol. 37, 1463–1468. https://doi:10.1021/es0208963. Li, Y.-M., Chaney, R.L., Brewer, E.P., Roseberg, R., Angel, J.S., Baker, A.J.M., Reeves, R., Nelkin, J., 2003. Development of a technology for commercial phytoextraction of

9

A. Tognacchini et al.

Journal of Environmental Management 254 (2020) 109798

van der Ent, A., Baker, A.J.M., Reeves, R.D., Pollard, A.J., Schat, H., 2013. Hyperaccumulators of metal and metalloid trace elements: facts and fiction. Plant Soil 362, 319–334. https://doi.org/10.1007/s11104-012-1287-3. van der Ent, A., Baker, A.J.M., Reeves, R.D., Chaney, R.L., Anderson, C.W.N., Meech, J. A., Erskine, P.D., Simonnot, M.-O., Vaughan, J., Morel, J.L., Echevarria, G., Fogliani, B., Rongliang, Qui, Mulligan, D.R., 2015. Agromining: farming for metals in the future? Environ. Sci. Technol. 49, 4773–4780. https://doi.org/10.1021/ es506031u. Vegli� o, F., Quaresima, R., Fornari, P., Ubaldini, S., 2003. Recovery of valuable metals from electronic and galvanic industrial wastes by leaching and electrowinning. Waste Manag. 23 (3), 245–252. https://doi.org/10.1016/S0956-053X(02)00157-5.

Vilarinho, C., Ribeiro, A., Carneiro, C., Castro, F., 2012. Recovery of copper and nickel hydroxide from galvanic sludge-Pilot scale experiments. In: 4th International Conference on Engineering for Waste and Biomass Valorisation, pp. 1561–1566. Wang, X., Shan, X., Zhang, S., Wen, B., 2004. A model for evaluation of the phytoavailability of trace elements to vegetables under field conditions. Chemosphere 55, 811–822. https://doi.org/10.1016/j.chemosphere.2003.12.003. Wenzel, W.W., Jockwer, F., 1999. Accumulation of heavy metals in plants grown on mineralised soils of the Austrian Alps. Environ. Pollut. 104, 145–155. https://doi. org/10.1016/S0269-7491(98)00139-0.

10