# Metal immobilization and soil amendment efficiency at a contaminated sediment landfill site: a field study focusing on plants, springtails, and bacteria

Valérie Bert, Christine Lors, Jean-François Ponge, Lucie Caron, Asmaa Biaz, Marc Dazy, Jean-François Masfaraud

► To cite this version:

Valérie Bert, Christine Lors, Jean-François Ponge, Lucie Caron, Asmaa Biaz, et al. Metal immobilization and soil amendment efficiency at a contaminated sediment landfill site: a field study focusing on plants, springtails, and bacteria. Environmental Pollution, 2012, 169, pp.1-11. 10.1016/j.envpol.2012.04.021. hal-00704382

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Submitted on 5 Jun 2012

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1 Environmental Pollution

2 Full Research Paper

3 **Title:** Metal immobilization and soil amendment efficiency at a contaminated sediment

4 landfill site: a field study focusing on plants, springtails, and bacteria

5 **Author names and affiliations:**

6 Valerie Bert<sup>a, b,*</sup>, Christine Lors<sup>b, c, d</sup>, Jean-François Ponge<sup>e</sup>, Lucie Caron<sup>a</sup>, Asmaa Biaz

7 <sup>e</sup>, Marc Dazy<sup>f</sup>, Jean-François Masfaraud<sup>f</sup>

8 <sup>a</sup> INERIS, RISK, DRC, Technologies et Procédés Propres et Durables, Parc

9 Technologique Alata, BP2, F-60550 Verneuil en Halatte, France

10 <sup>b</sup> Centre National de Recherche sur les Sites et Sols Pollués, BP 537, 59505 Douai

11 Cedex, France

12 <sup>c</sup> Université Lille Nord de France, 1 bis rue Georges Lefèvre, 59044 Lille Cedex, France

13 <sup>d</sup> Ecole des Mines de Douai, LGCgE-GCE, 941, rue Charles Bourseul, 59500 Douai

14 Cedex, France

15

16 <sup>e</sup> Muséum National d'Histoire Naturelle, CNRS UMR 7179, 4 avenue du Petit-Château,

17 91800 Brunoy, France

18 <sup>f</sup> Université Paul Verlaine-Metz, Laboratoire "Interactions Ecotoxicologie, Biodiversité,

19 Ecosystèmes", CNRS UMR 7146, Campus Bridoux, rue du général Delestraint, 57070

20 Metz, France

21

22 <sup>*</sup>Corresponding author. Tel.: +33 3 44 55 63 82; fax: +33 3 44 55 65 56.

23 **E-mail address:** <valerie.bert@ineris.fr> (V. Bert).
24 Capsule

25 In-situ incorporation of Thomas Basic Slag into a landfilled metal-contaminated
sediment decreases metal mobility and ecotoxicity and increases bacterial activity.

27

28 Abstract

29 Metal immobilization may contribute to the environmental management strategy of
30 dredged sediment landfill sites contaminated by metals. In a field experiment,
31 amendment effects and efficiency were investigated, focusing on plants, springtails and
32 bacteria colonisation, metal extractability and sediment ecotoxicity. Conversely to
33 hydroxylapatite (HA, 3 % DW), the addition of Thomas Basic Slag (TBS, 5 % DW) to a
34 5-yr deposited sediment contaminated with Zn, Cd, Cu, Pb and As resulted in a decrease
35 in the 0.01M Ca(NO3)2 extractable concentrations of Cd and Zn. Shoot Cd and Zn
36 concentration in *Calamagrostis epigejos*, the dominant plant species, also decreased in
37 the presence of TBS. The addition of TBS and HA reduced sediment ecotoxicity and
38 improved the growth of the total bacterial population. Hydroxylapatite improved plant
39 species richness and diversity and decreased antioxidant enzymes in *C. Epigejos and*
40 *Urtica dioica*. Collembolan communities did not differ in abundance and diversity
41 between the different treatments.

42

43 **Keywords**: basic slag; *Calamagrostis epigejos*; dredged sediment; hydroxylapatite; ecotoxicity

45

46
47 **1. Introduction**

Human activities during the last decades have contaminated canal sediments
with various organic and inorganic pollutants. Those of most concern are metal(loid)s,
polyaromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), and mineral
oils. Polluted sediments in canals may be disposed on land. Indeed, the maintenance of
waterways requires dredging on a regular basis to prevent flooding, facilitate
navigation, and allow for use of a given water system. European Community policy
encourages sediment recovery (Directive 2008/98/EC). Nevertheless, due to the high
concentration of pollutants and their potential toxicity, contaminated dredged sediments
cannot be used in civil engineering as raw material or the deposit cannot be used to
produce valuable biomass. Many treatments are available for contaminated sediments,
but relatively few are applicable to metal(loid) pollution. Currently, treatment and reuse
of heavily contaminated dredged materials is not a cost-effective alternative to disposal
landfill sites (Bert et al., 2009). In the Nord-Pas-de-Calais region (France), which is
affected by intensive industrial activities, local authorities are required to manage
contaminated landfill sites where large volumes of polluted dredged materials were
deposited. The regional division of Voies Navigables de France (VNF) developed a
management strategy for its disposal sites. This strategy includes the implementation of
an environmental management system which aims to meet best practices and comply
with environmental regulation in the field of human health and environment. The VNF
is involved in the reclamation of its disposal sites into ‘natural’ and ‘green’ zones
(Prevost, 2008). Metal immobilization in metal-contaminated sediments at some landfill
sites may contribute to this environmental management strategy.
Metal immobilization aims at (1) changing speciation of trace elements (TE) in the substrate to reduce their soluble and exchangeable fractions, (2) limiting TE-uptake by plants, (3) reducing the direct exposure through soil by reducing metal availability to heterotrophic organisms, resulting in enhanced biodiversity (Vassilev et al., 2004). In situ immobilization of metal(loid)s is achieved by incorporating amendments into the soil, promoting their sorption and precipitation. In addition to immobilization effects, soil amendment can improve soil fertility by increasing pH, organic matter content, microbial activity, and moisture retention, and reducing soil compaction (Vangronsveld et al., 1995, 1996).

The efficiency of amendments is site-specific, depending on various factors among which pollutant types and soil properties (Kumpiene et al., 2008). Numerous soil amendments have been tested in pots, at pilot and field scale, for remediation purposes (Bes and Mench, 2008; Knox et al., 2001; Kumpiene et al., 2008). The effects of Thomas Basic Slag (TBS) and hydroxylapatite (HA) incorporation have only been investigated in batch and pot experiments (Bes and Mench, 2008; Boisson et al., 1999; Friesl et al., 2006; Mench et al., 1994a, 1994b; Mench et al., 2000; Misra and Chaturvedi, 2007; Negim et al., 2010; Panfili et al., 2005). Hydroxylapatite, a mineral from the phosphate group, is the major component of tooth enamel and bone mineral. It is mainly used for medical purposes. Hydroxyapatite powders can be synthesised via numerous production routes, using a range of different reactants. Some processing techniques include wet chemical methods (precipitation), hydrothermal techniques, hydrolysis of other calcium phosphates and sol-gel. Due to its particular properties including the sorption of metallic ions, HA can be useful for the management of contaminated groundwater and soil. The application of HA [Ca₁₀(PO₄)₆(OH)₂] to
contaminated soil immobilizes dissolved Pb (Ma et al., 1993). Other studies on metal stabilization by phosphorous compounds focused on reduction of plant uptake of metals, as well as reduction of their solubility and mobility (Knox et al., 2001). After HA incorporation into a contaminated soil, the concentrations of exchangeable metals decreased and plant uptake of these elements was reduced (Boisson et al., 1999).

However, arsenic uptake by plants increased and there were nutrient deficiencies. Therefore it was concluded that HA application could be effective to immobilize Zn, Pb, Cu and Cd but was inappropriate in the case of mixed metal-arsenic pollution and when potential nutrient deficiencies may occur. Thomas Basic Slag (TBS) is an alkaline by-product of steel industry used as a fertilizer by farmers. It usually increases soil pH (Bes and Mench, 2008) and contains P, Mn, Mg, and Fe (Panfili et al., 2005). Soil treatment with TBS decreased Cd and Zn-soluble and $(Ca(NO_3)_2)$ exchangeable fractions and reduced Cd concentrations in tobacco shoots (Mench et al., 1994a, 1994b).

Zinc availability decreased after TBS addition and this persisted over 5 months without any phytotoxicity (Mench et al., 2000). Bes and Mench (2008) studied the incorporation of 0.25% and 3.9% (DW) TBS into Cu-contaminated soils. Based on soil phytotoxicity, 3.9% TBS was one of the most efficient amendments even though high levels of Cu remained in the soil solution. Panfili et al. (2005) studied the effect of HA and TBS on Zn in a metal-contaminated sediment and suggested that the formation of Zn phosphate contributed to Zn immobilization.

Plants, bacteria and soil fauna, notably springtails (Collembola), can develop tolerant populations on metal-contaminated soils (Bes et al., 2010; Escarré et al., 2011; Gillet and Ponge, 2003; Lors et al., 2005; Ryan et al., 2004; Tyler et al., 1989; Vangronsveld et al., 1995, 1996). Vegetation reduces contaminant mobility by reducing
118 leaching, soil erosion, and improves the aesthetic value of formerly barren areas
119 (Ruttens et al., 2006). The vegetation itself may contribute to metal immobilization in
120 the rhizosphere and through the production of litter (Bouwman and Vangronsveld,
121 2004). When metal bioavailability and exposure decrease, due to vegetation or
122 amendment or both, sensitive plants, bacteria and soil fauna can develop and participate
123 in turn to the restoration of an ecosystem on contaminated soils (Bouwman et al., 2001;
124 Lock et al., 2003; Lock and Janssen, 2003, 2005; Vangronsveld et al., 1995, 1996).

125 This field-scale work was performed at an experimental metal-contaminated
126 sediment landfill site. The aim was to evaluate the efficiency of TBS and HA
127 incorporation on metal(oid) immobilization. Efficient soil amendments may decrease
128 metal(oid) extractability and shoot metal concentrations. In parallel, the potential
129 adverse or beneficial effects of TBS and HA incorporation were investigated on an
130 array of biological parameters related to plant and springtail communities composition
131 and diversity, total and metal specific bacterial populations (anaerobic sulphate-
132 reducing bacteria and aerobic sulphur-oxidising bacteria), springtail trophic status (gut
133 contents), sediment ecotoxicity and sublethal effects in plants (maximum photochemical
134 efficiency of photosystem II and antioxidant enzymes activity levels). This battery
135 might be relevant to assess the efficiency of soil additives.

136 This is the first time that a full-scale study of metal immobilization combined
137 with ecosystem monitoring is reported on a contaminated dredged sediment disposal
138 site.

139

140 **2. Materials and methods**

141 2.1. *Description of experimental site and sediment treatment*
In May 2002, a field trial was set up in an agricultural area in Lallaing (North of France, 50°23'17''N and 3°11'59''E). Three plots of 60 m² and 40 cm depth were dug in an uncontaminated soil. These plots were filled with freshly dredged sediments from the nearby Scarpe canal (Pont de Râches). These sediments were contaminated due to past and present non-ferrous metal processing and smelting activities. Two plots were used to assess soil additives. Thomas Basic Slag (TBS) and a synthetic hydroxylapatite (HA), both in powdered forms, were singly incorporated into the sediment at a rate of 5% and 3% DW, respectively. TBS was obtained from Cedest® (Mâcon, France) and HA was from Brenntag® (Mülheim/Ruhr, Germany). The purity of HA was certified to 99% by the supplier. The third plot remained untreated (NT). After TBS and HA incorporation, the treated and untreated plots were mechanically homogenized with a crane shovel for two hours to ensure that the mixture was homogeneous. The plots were air dried for two months to reduce the sediment water content. The three plots were further subdivided into 9 sub-plots of 20 m² each, resulting in 3 untreated plots (NT), 3 HA-treated plots and 3 TBS-treated plots. Six of these plots were further sown with grasses (Deschampsia cespitosa and Festuca rubra) while the 3 remaining plots remained unplanted. This study reported the work performed on these unsown plots.

From mid-spring 2003, spontaneous vegetation started to develop on these plots. Vegetation management (i.e. removal of selected species, harvest, fertilizer addition, and irrigation) was not carried out from this time to March 2007.

## 2.2. Collection of sediment and soil samples

In March 2007, top-sediments (0-20 cm depth) of the three tested plots (NT, HA, TBS) were sampled. Three surface sediment samples per plot were randomly collected with a hand auger to determine (pseudo)-total metal(oid) and extractable concentrations,
166 pH<sub>water</sub> and sediment ecotoxicity. Six additional randomly collected samples were taken
167 on each of the plots at a depth of about 20 cm using a hand auger for bacterial analyses
168 (3 samples/plot) and a corer of 5 cm diameter for collembolan analyses (3 samples /
169 plot). One composite sample per plot was formed with 5 sub-samples randomly
170 collected to determine the other physico-chemical characteristics (particle size, organic
171 carbon, total nitrogen, carbonates and cation exchange capacity). In parallel, an
172 uncontaminated soil (0-20 cm depth), located outside the plots, was collected. This
173 control area (CA) was selected because it had similar spontaneous vegetation to the
174 plots. Physicochemical properties were determined on a composite sample made of 5
175 sub-samples collected below the vegetation of interest. Three additional soil samples
176 were collected for collembolan analyses.

177 2.3. Sediment and soil characteristics

178 All samples were air dried until constant weight and sieved through 2 mm mesh
179 size. All analyses on composites have been performed at the Laboratoire Départemental
180 d'Analyses et de Recherche (LDAR), Laon, France using standard methods (particle
181 size (NF X 31 – 107, 2003), organic carbon (NF ISO 14 235, oxidation method, 1998),
182 total N (NF ISO 11261, colorimetric method, 1995), carbonates (NF ISO 10693, 2004),
183 cation exchange capacity (CEC, NF X31-130, 1999).

184 A sample of 0.5 g of air-dried and sieved sediment and soil was weighed into 70
185 mL teflon microwave tubes. Eight mL of *aqua regia* containing 2 mL 67% concentrated
186 HNO<sub>3</sub> and 6 mL 36% concentrated HCl (NF EN 13657, 2003) was added. Samples
187 were heated in a microwave digester (MARS Xpress, CEM Corporation<sup>®</sup>, Matthews,
188 NC) to 180 °C for 20 min, with a 30 minute ramp time. After filtration through a 0.45
189 µm Whatman filter, the pseudo-total metal(oid) concentrations (As, Cd, Pb, Zn, Cu)
were determined using an Inductively Coupled Plasma Atomic Emission Spectrometer ICP-AES (Jobin Yvon®, Longjumeau, France). To assess the analytical quality, a standard reference sediment material SRM 2704 (Buffalo River Sediment, Standard Reference Material) was subjected to the same protocol. Recoveries were: 100 % for Cd, 101% for Zn, 104% for Pb, 90% for Cu and 94% for As.

Soil pH was measured in 1:5 sediment/soil:water suspension using a glass electrode pH meter (NF ISO 10390, 1994).

Table 1 lists physico-chemical properties of contaminated sediments (NT, HA, TBS) and the uncontaminated soil (CA).

## 2.4. Extractable trace element analysis and sediment ecotoxicity

Extractable sediment metal(oid) concentrations of treated and untreated plots were determined after extraction with 0.01 M Ca(NO₃)₂. Prior to analysis, 20 g of air-dried and sieved sediment were shaken for 48 h with 40 mL of 0.01 M Ca(NO₃)₂ solution. Extracts were filtered through a 0.45 µm cellulose membrane and metal concentrations were measured using ICP-AES.

Ecotoxicity was assessed by two standardized bioassays: the bacterial Microtox® assay and the algal test, using Vibrio fischeri and Pseudokirchneriella subcapitata as test organisms, respectively. Those two assays were previously shown to be highly sensitive to leachates obtained from dredged materials (Piou et al., 2009). Prior to the bioassays, soil leachates were prepared according to the French standard AFNOR X31-210 (1991) and diluted (10 to 90%) with demineralised water. Ecotoxicity assays were run according to AFNOR T90-320 (1991) and ISO 8692 (1989) standards for Microtox® and algal bioassays, respectively. The half maximum effective concentration
213 (EC50, Microtox® bioassay) and the 50% inhibitory concentration (IC50, algal bioassay)
214 were derived from dilution-effect curves. When the toxicity of leachates was not
215 sufficient to reach 50% of effect, EC50% or IC50 values were indeterminable and the %
216 of effect observed at the 90% dilution was then reported.

217 2.5. Plant survey, biodiversity indices and plant analysis

218 2.5.1. Plant survey and similarity index

219 Vascular plants were surveyed in May and June 2007 in all plots using
220 Lambinon et al. (2004) as a standard reference book for plant taxonomy. The total
221 vegetation cover and the cover rate of mosses, herbaceous plants, shrubs and trees were
222 assessed on each plot. Individuals per species and per plot were counted and numbers
223 were related to plot area. Species richness corresponded to the number of species per
224 plot.

225 A similarity index (Sørensen, 1948) was calculated to compare the composition
226 of vegetation between (i) experimental plots (influence of treatments) and (ii)
227 experimental plots and surrounding biotopes (identification of putative biotopes that
228 may contribute to plot colonization).

229 2.5.2. Biodiversity indices

230 The diversity of vascular plants was assessed using the Shannon-Weaver index
231 H' (Shannon and Weaver, 1949), calculated as

232 $$H' = - \sum_{i=1}^{s} p_i * \log_2 p_i$$
233 where *pi* is the species proportion in a community composed of *s* species. The
234 homogeneity of species distribution was assessed through the regularity index *J* of
235 Piélou (1966).

236 **2.5.3. Plant analyses**

237 *Calamagrostis epigejos*, a grass, and *Urtica dioica*, a forb, were chosen for
238 phytotoxicity and metal content analyses as both herb species were present in all plots
239 and in the control area in sufficient abundance to sample 5 individuals per species of
240 approximately similar size and sufficient biomass to perform analyses. Plant sampling
241 occurred in June 2007.

242 **2.5.3.1 Chlorophyll fluorescence**

243 The fluorescence of chlorophyll a was measured on dark-adapted leaves of *C.*
244 *epigejos* and *U. dioica* submitted to a saturating light pulse using a portable chlorophyll
245 fluorimeter (Handy-PEA, Hansatech Instruments®, Norfolk, UK). Basal fluorescence
246 (F0) and maximum fluorescence (Fm) values were used to derive Fv/Fm ratio
247 (maximum photochemical efficiency of PSII), with the variable fluorescence Fv=Fm-
248 F0.

249 **2.5.3.2 Antioxidant enzymes**

250 Leaves from both plant species were collected from CA, NT, HA and TBS plots
251 and immediately transferred to polypropylene tubes then frozen in liquid nitrogen. At
252 the laboratory, leaf samples were crushed manually in a 125 mM phosphate buffer at pH
253 7.8 by using a porcelain mortar placed on ice. After centrifugation (15 000 g for 10 min
254 at 4° C), the supernatant was used to determine antioxidant enzyme activities.
Superoxide dismutase (SOD), catalase (CAT), ascorbate peroxidase (APX), guaiacol peroxidase (GPX), and glutathione reductase (GRD) activities were measured according to methods described by Dazy et al. (2008). The protein concentration was determined in plant extracts (Bradford, 1976); enzyme activity levels were expressed in µmoles/min/mg protein excepted for SOD for which one unit of activity corresponded to the amount of enzyme causing 50% inhibition of absorbance readings by comparison to control tubes lacking enzymes.

### 2.5.3.3 *Elemental concentrations in plant shoots*

Plant shoots were carefully washed (3 times) with distilled water and oven-dried at 40°C until constant weight. Weighed aliquots (0.5 g DW) were wet digested in 10 mL HNO₃ (65%) using a microwave digester (Mars Xpress) according to the following program: 3 min to reach 100 °C, 2 min to reach 140 °C, 2 min to reach 160 °C, 2 min to reach 180 °C and 20 min at 180 °C. Solutions were filtered to <0.45 µm (Whatman filter) and concentrations of Cd, Zn, Pb, Cu, As, Ca, Fe, Mg, Mn, P, Na, K and Al were analyzed using ICP-AES (VARIAN 720 AES-ICP, Ulys®, France). Blanks and certified reference material (trace elements in hay, IAEA V10, International Atomic Energy Agency, Vienna, Austria) were included for quality control. Recoveries were: 100% for Ca, 91% for Cu, 79% for Fe, 113% for Mg, 91% for As, 89% for Mn and Pb, 105% for K, 89% for Na, 118% for P, 87% for Zn, 93% for Cd and 67% for Al.

## 2.6 *Collembola identification and gut content observation*

After sampling at a depth of 20 cm, considering that the whole collembolan community was sampled on a surface of ~20 cm², sediments (NT, HA, TBS) and soil
278 (CA) samples were embedded in polythene bags then transported to the laboratory. The
279 extraction of arthropods started on the same day using the dry funnel (Berlese) method
280 (Edwards and Fletcher, 1971) and was completed within 7-10 days. The daytime
281 maximum temperature averaged 25-30 °C. Animals were collected and preserved in
282 ethyl alcohol until sorting and identification. Sorting was done in alcohol under a
283 dissecting microscope at x 20 magnification. All springtails were mounted in chloral-
284 lactophenol then observed in a light microscope under phase contrast at x 400
285 magnification. Identification was done at the species level using keys by Bretfeld
286 (1999), Fjellberg (1998), Gisin (1960), Hopkin (2000), Jordana et al. (1997), Potapov
287 (2001) and Zimdars and Dunger (1994), Gut contents were observed by transparency
288 and classified in 8 categories using morphological features observable in phase contrast
289 light microscopy at x 400 magnification, and were counted according to the method
290 devised by (Gillet and Ponge, 2003, 2005): algae, animal remains, bacteria, melanised
291 and hyaline hyphae, fungal spores, holorganic humus, hemorganic humus, to which
292 empty guts were added. The examination of gut contents may elucidate the composition
293 of horizons in which the animals were living (Ponge, 2000). We sought to determine
294 when toxicity forced the organisms to move to horizons in which they do not normally
295 live or delete some food resources from their current diet (Gillet and Ponge, 2003).

2.7 *Bacterial analyses*

297 Sediment samples (NT, HA and TBS) were collected in polythene bags then
298 transported to the laboratory. After homogenization, samples were maintained at 4°C
299 before performing bacterial analyses. The enumeration of total sulphur-oxidising
300 bacteria (SOB) was carried out on Nunc® Delta (Nalge Nunc International®, Roskilde,
301 Denmark) microplates of 96 wells (8 lines x 12 columns), on which each well contained
302 250 µL of specific medium. Nutrient Broth® (Difco®, Detroit, USA) medium (diluted to
303 1/10), added to fungicidal cyclohexamide solution (0.2 g L⁻¹), was used for total
304 bacterial count. The medium used to enumerate neutrophilic SOB contained 5 g
305 Na₂SO₃·5H₂O, 2 g K₂HPO₄, 0.1 g (NH₄)₂SO₄, 0.1 g MgSO₄·7H₂O, 0.1 g CaCl₂, 0.002 g
306 FeCl₃·6H₂O, 0.005 g MnSO₄·4H₂O and 3 ml bromocresol purple (c = 0.02 g L⁻¹). The
307 pH of the medium was initially adjusted to pH 6.5 with NaOH. For acidophilic SOB,
308 the medium, which contained 1.5 g KH₂PO₄, 0.1 g CaCl₂, 0.1 g MgCl₂·H₂O, and 0.1 g
309 NH₄Cl, was combined with 3 ml of bromocresol green (c = 0.02 g L⁻¹). The pH was
310 initially adjusted to 4.5 with HCl. A suspension containing 10 g of dry soil and 50 ml of
311 Ringer solution was mixed in a Waring blender for 1.5 min at high speed. After
312 decantation, the solution was successively diluted with Ringer solution to 10⁻¹ – 10⁻⁸
313 (Ramsay, 1984). Each well was inoculated with 25 µl of solution. The wells of the eight
314 lines corresponded to the eight tested dilutions. Each sample was inoculated in the wells
315 of three columns and a column was not inoculated to control the sterility of the medium.
316 As each microplate had twelve columns, only one microplate was used for the three
317 samples of a given plot. The inoculated plates were incubated at 30°C for 15 days and
318 then the coloured products were measured. Bacteria were counted using the Most
319 Probable Numbers (MPN) method (De Man, 1977). Sulphate-reducing bacteria (SRB)
320 were counted in tubes containing a solid medium specific to SRB (American Petroleum
321 Institute, Washington, DC) (Lors et al., 2004).

## 2.8 Statistical analyses

323 A one-way ANOVA was used to detect treatment effects on metal
324 concentrations, Fv/Fm ratio, antioxidant enzyme activities, species distribution,
325 biodiversity indices, distribution of gut contents and bacterial numbers. When ANOVA
326 was significant, mean values were compared with post-hoc multiple-comparison procedures. When the sample weight was too low and variance homogeneity was not fulfilled, a non-parametric statistical test followed by a multiple-comparison procedure was performed. Correlation analysis, using Pearson product-moment correlation coefficient, was performed between metal concentrations and antioxidant enzymes activities in leaves. Statistica v7 (StatSoft Inc.) was used for bacterial calculations whereas XLSTAT® (Addinsoft®, Paris, France) was used for springtail calculations. All other statistics were made with PASW statistics 18.

334 A semi-quantitative method was specially devised to make an overall assessment of beneficial or adverse effects of soil amendment (HA and TBS) addition on dredged sediments. For this purpose, the treatments (NT, HA, TBS) were ranked 1 to 3, a value of 1 being attributed to the treatment that maximized variables describing optimal features of plant, faunal and microbial communities (i.e. vegetation succession state, species richness, species diversity, equitability, abundance) and plant Fv/Fm ratio, or minimized those describing metal mobility, ecotoxicity, plant metal concentration and antioxidant enzyme activity. Some variables cannot be taken into account in these calculations, either because they may respond to other factors than soil toxicity (acid vs neutral sulphur-oxidizing bacteria, collembolan gut contents, vegetation composition, pH), or because they do not differ significantly among treatments (plant Fv/Fm ratio for *C. epigejos*). Ranks were then averaged among the different variables which had been kept for the calculations. Overall differences between the three treatments NT, HA and TBS were then tested by a sign-test, using ranks as unit values.

348 **3 Results**
349 **3.1. Metal(oid) mobility and sediment ecotoxicity**

350 Extractable Pb and As concentrations in all plots were below quantification
351 limits (0.07 and 0.1 mg kg⁻¹ DW for Pb and As, respectively). Cadmium and Zn showed
352 a similar pattern with higher 0.01 M Ca(NO₃)₂-extractable concentrations in the NT plot
353 than in the TBS-treated plot, while intermediate values were found in the HA-treated
354 plot (Table 2). Expressed as percentages of the total metal concentrations, extractable
355 Cd and Zn represented 0.14 and 0.07%, respectively, in the TBS-treated plot, whereas
356 they represented 0.37 and 0.76% in the NT plot and 0.31 and 0.61% in the HA-treated
357 plot. For Cu, no significant differences occurred across the plots.

358 Sediment leachates were moderately toxic to the bacteria *V. fischeri* since no
359 EC₅₀ could be determined. Considering the % of effect measured at the 90% dilution,
360 the ecotoxicity of sediment leachates could be ranked in the following order NT > HA >
361 TBS (Table 2). Sediment leachates were more toxic to Algae, especially those prepared
362 from NT and HA (IC₅₀ values being 12.8 and 19.2 %, respectively). TBS sediment
363 leachate had only a slight inhibitory effect on Algae.

364

365 **3.2. Plant survey, plot similarity, source of colonization and element concentrations**
366

367 The composition of the plant community was influenced by the treatments of the
368 plots as shown by plant survey and the number of individuals per species and per
369 surface (Table 3). The TBS plot was mainly colonized by herbaceous plants while HA
370 and NT plots were dominated by shrubs. The maximum height of shrubs was higher in
371 HA plot than in NT plot. The TBS plot had the lowest plant species richness (n=14).
Based on Shannon-Weaver diversity and Piélou equitability indices, the HA plot had the highest plant diversity with a more homogeneous species distribution while TBS and NT plots showed a lower diversity and at least one species that dominated the overall community (Table 3).

In all plots, the dominant family was Poaceae with 5 species for HA, 4 species for TBS and 3 species for NT. Among Poaceae, only *C. epigejos* was present on all plots. *Calamagrostis epigejos* was more present on the NT plot (46.9%), than on TBS (30%) and HA plots (19.9%). On all plots, other plant species were abundant: *Carex hirta* (11.5%) and *Glechoma hederacea* (11.7%) on the NT plot, *Betula pendula* (12.7%) and *G. hederacea* (22.8%) on the HA plot, and *Phalaris arundinacea* (32.1%), *Galium aparine* (16.4%) and *Epilobium angustifolium* (11.7%) on the TBS plot. The second most abundant family was *Asteraceae* with three species. At least two *Asteraceae* were present on all plots. Perennial species dominated the three group communities. Plant species were native to the Nord-Pas-de-Calais region with no clear specificity linked to sediment traits except for some plants showing hygrophilic characteristics (e.g. *Eupatorium cannabinum*, *C. hirta*, *Carex cuprina*, *P. arundinacea*, *Salix alba*).

The Sørensen index, comparing vegetation composition between plots and between plots and surrounding biotopes, indicated that 85% of plant species were common between NT and HA plots whereas the TBS plot showed 60% and 57% species similarity with HA and NT plots, respectively (data not shown). The surrounding biotopes consisted of a slag heap, a ruderal zone, a canal bank, and an agricultural land. A similarity index indicated that the plant colonizers of the NT and HA plots mainly originated from the slag heap, with 45 and 41% of species similarity.
This result agreed with the direction of the south-easterly prevailing wind and the seed shadow colonization hypothesis (Clark et al., 2005), as the slag heap was situated south-east from the plots. The TBS plot showed less similarity with the slag heap (33%).

The concentrations of the chemical elements were measured in *C. epigejos* and *U. dioica* as both species were sampled in all plots (Table 4). Their shoot Pb and As concentrations were below quantification limits (1 and 0.5 mg kg⁻¹ for Pb and As, respectively). Shoot Cd and Zn concentrations for both species and Cu concentration for *U. dioica* were lower in CA than in the three treatments. The shoot Cd concentration was the lowest in *C. epigejos* grown on TBS whereas for *U. dioica* it did not differ across treatments. Shoot Zn concentration was the lowest in *C. epigejos* grown on TBS. The highest shoot P and K concentrations were measured in *U. dioica* grown on TBS. No other significant effect of TBS treatment on metal concentration was observed. The HA treatment did not decrease shoot Cd and Zn concentrations of *C. epigejos* and *U. dioica* compared to NT but increased shoot Cu in *C. epigejos*. Shoot Mg, Na, and Al concentrations in *C. epigejos* and shoot Fe concentration in *U. dioica* were higher in the HA plot than in the NP plot. The HA treatment did not influence other shoot element concentrations.

413

### 3.3. Antioxidant enzyme activities and chlorophyll fluorescence of leaves

Unstressed plants typically exhibit Fv/Fm ratios of 0.8-0.83 (Laposi et al., 2009), which was the case for *C. epigejos* on all plots (Fig. 1). For all antioxidant enzymes, lowest activities were determined in CA plants. In untreated sediment (NT), *C. epigejos* enzymes were significantly increased 1.2 to 3 fold depending on the enzyme considered.
For plants collected on HA plots, the activity of SOD, CAT, and GRD were similar to CA values, except APX which remained higher. For GPX, the HA treatment had no significant effect. This enzyme was less influenced than the others by the composition of the NT sediment since its activity only increased by a factor of 1.2 compared to plants harvested on CA plots. Plant enzymes responded differently to the TBS treatment. For SOD, CAT and APX, activities levels were not influenced since their values did not differ significantly from corresponding values in NT plants. GPX and GRD were both affected by TBS treatment, but in opposite ways, since GPX increased by comparison to NT values while GRD decreased.

*Urtica dioica* collected from CA and HA plots, and to a lesser extend NT plots, had all Fv/Fm ratios around 0.8 (Fig. 2). For TBS-plants, the Fv/Fm ratios were significantly lower. As noted for *C. epigejos*, for all antioxidant enzymes, the lowest activities were found for CA plants. For the untreated sediment (NT), *U. dioica* enzymes significantly increased by a factor 3 to 5. The HA and TBS treatments had opposite effects on antioxidant enzymes. All tested enzyme activities in *U. dioica* leaves decreased for HA compared to NT and tended to recover CA values. However, differences between HA and CA plants remained significant for SOD, APX and GPX activities. In contrast, TBS plants still exhibited high enzyme activities, similar to NT values (e.g. APX and CAT) or even higher (e.g. SOD, GPX, and GRD).

### 3.4. *Correlations between shoot element concentrations and antioxidant enzyme activities*

A correlation analysis was performed, between shoot element concentrations in *C. epigejos* and *U. dioica* and their antioxidant enzyme activities in leaves, to reveal significant linear relationships (Table 5). All enzyme activities negatively correlated
with shoot Cd concentration in *C. epigejos*. Manganese concentrations were positively correlated with 4 out of the 5 enzyme activities whereas Cu, Zn, Mg, and Na concentrations were correlated with GRD. For *U. dioica*, no relationship was found between shoot Cd concentration and any of the enzyme activities, but all enzyme activities were positively correlated with shoot Cu, Zn, P, and K concentrations.

### 3.5. *Collembolan diversity and gut contents*

A total of fourteen species were found in the four areas (Table 6). Neither total abundance nor biodiversity indices (species richness, Shannon index, equitability) showed significant differences between treatments. However, the TBS plot tended to harbour more specimens and species than NT and HA plots and CA. In the NT plot, there was no decrease in any of the four biodiversity indices when compared to CA, with a mean abundance of 20 specimens per core sample, shared between four species. At the species level, *Friesea truncata* was only present in the TBS samples.

The only significant effect on the gut contents of Collembola (Fig. 3) was an increase in the percent of specimens with hyaline fungal hyphae, which increased in all cores sampled in the HA plot, reaching an average of 16% compared to less than 3% for the other treatments. Hemorganic humus (amorphous organo-mineral matter) was dominant everywhere, indicating that the sediments (treated or not) were consumed by springtails, in particular by the dominant species *Mesaphorura florae*.

### 3.6. *Total and specific bacterial numbers*

The total bacterial population in all plots varied from $10^7$ to $10^8$ bacteria g$^{-1}$ dry sediment, and was similar to that usually found in unpolluted topsoils (Robert, 1996;
Taylor et al., 2002) (Table 7). However, in the TBS plot, the total bacterial population (in $10^8$ bacteria.g$^{-1}$ dry soil) was significantly higher (1.7) than in the NT plot (0.21).

In all plots, neutrophilic sulphur-oxidising bacteria (NSOB) were the most abundant group, with about $10^5$ bacteria g$^{-1}$ dry sediment, whereas acidophilic sulphur-oxidising bacteria (ASOB) were the least abundant group with about $10^3-10^4$ bacteria g$^{-1}$ dry sediment. The amendments did not modify the size of these bacterial populations.

The sulphate-reducing bacterial microflora (SRB) was about $10^1-10^2$ bacteria g$^{-1}$ dry sediment and was of the same order of magnitude in all plots.

### 3.7. *Integration of data from the ecotoxicity battery*

Averaging ranks of all variables selected for the synthetic assessment of treatment effect (see Material and methods) indicated that TBS and HA exhibited similar favourable effects on the ecosystem when compared to the untreated sediment (NT) (Table 8). HA and TBS treatments ranked significantly above NT (one-sided t-tests, P = 0.001 and 0.011, respectively) while they did not differ from each other (two-sided t-test, P = 0.797).

## 4. Discussion

Although the TBS application rate and the soil type used in this study differed from those investigated in other works, our results are broadly similar for Zn and Cd. The incorporation of TBS (5% DW) into the metal contaminated sediment decreased the 0.01 M $Ca(NO_3)_2$-extractable Cd and Zn fractions and shoot Cd and Zn concentrations of *C. Epigejos* (Tables 2 and 4), confirming previous findings (Mench et al., 1994a, 1994b, 2000). The TBS incorporation into a Cd-Zn-contaminated soil, reduced both the
488 Ca(NO₃)₂ extractability of Cd and Zn and their contents in tobacco shoots (Mench et al., 1994a, 1994b). In addition, TBS added into a Zn-contaminated soil increased soil pH, decreased extractable Zn and shoot Zn concentration in ryegrass, 3 and 5 months after soil treatment (Mench et al., 2000). Contrary to TBS, HA incorporation (3% DW) into the sediment did not decrease significantly Cd and Zn extractability nor Cd and Zn concentrations in plants (Tables 2 and 4). These results were not consistent with those of Boisson et al. (1999) who found a decrease of extractable Cd and Zn and metal concentrations in maize when compared to the untreated soil and with increasing HA application rate (0.5, 1 and 5% DW).
489
490
491
492
493
494
495
496

497 In contrast to other works (Boisson et al., 1999; Friesl et al., 2006), As-mobilization was not observed either in the presence of HA or TBS (Table 2). No amendment effect was recorded for Cu (Tables 2 and 4) whereas TBS incorporation (0.25 and 3.9% DW) into a Cu contaminated soil decreased Cu concentrations in the primary leaves of bean (Bes and Mench, 2008).
498
499
500
501

502 In our field study, TBS was the most efficient amendment to decrease Zn and Cd mobility and transfer to the plants. We hypothesise that Zn- and Cd-phosphates were precipitated, thereby immobilizing these elements in a non-available form as suggested by Panfili et al. (2005) who studied Zn speciation in the same metal-contaminated sediment treated with TBS. In our field experiment, the low efficiency of HA compared to TBS may be explained by the low solubility of HA at neutral pH (Table 1). As suggested by Miretzky and Fernandez-Cirelli (2007), the solubility of the phosphate phase is necessary for successful in situ treatment, metal immobilization by phosphorus being limited or inefficient when the matrix is neutral or alkaline.
503
504
505
506
507
508
509
510
As the essential nutrient concentrations did not decrease in plant shoots (Table 4), our results suggested that neither TBS nor HA led to deficiency problems in the tested plant species. In the presence of HA (5% DW), Boisson et al. (1999) reported a Mn-deficiency in maize. Neither treatment promoted the uptake of essential nutrients such as Ca except in the case of *U. dioica* where shoot P and K concentrations were increased by TBS addition. In primary leaves of bean, Ca concentration was enhanced in the presence of TBS whereas P concentration was not promoted (Bes and Mench, 2008; Negim et al., 2010).

In our study, sediment ecotoxicity was reduced by the incorporation of TBS and HA, TBS being the most efficient amendment.

At excessive concentrations in the soil solution, Cd, Cu, Pb and Zn exert toxic effects on plants, including alterations in photosynthetic and respiration processes or inhibition of plant growth. They may also stimulate the formation of reactive oxygen species (Dazy et al., 2009). To prevent oxidative damage, plant cells usually use enzymatic protection mechanisms, such as superoxide dismutase, catalase, and peroxidases, and metabolites such as ascorbate and glutathione (Dazy et al., 2008).

Following amendment of the soil, a decrease in phytotoxicity and a recovery in plant antioxidant defence to control values are expected as a result of metal-immobilization processes (Ruttens et al., 2006; Vangronsveld et al., 1995, 1996). In the NT plot, all antioxidant enzyme activities of both plant species were increased compared to the CA plot, indicating that the plants suffered from oxidative stress. The increase in shoot Zn and Cu concentrations were consistent with enzymatic activity increases, suggesting that these metals accounted for the stress response recorded in plants growing on untreated dredged sediment. This was not the case for Cd, as no correlation for *U*.
dioica and no positive correlation for *C. epigejos* were found. The addition of HA permitted an almost total recovery of antioxidant enzyme activities to CA values, whereas the addition of TBS did not. Thus, HA was a better amendment for keeping the stress responses of plants at a low level. Levels of chlorophyll fluorescence (Fig. 1 and Fig. 2) indicated that the plants were not stressed on NT, HA and TBS plots. There are few studies that report the effect of TBS on phytotoxicity and an absence of studies on the effect of HA on phytotoxicity. After the incorporation of TBS (0.25% and 3.9%) into a Cu-contaminated soil, the chlorophyll density of primary leaves of bean decreased and guaiacol-peroxidase activity in roots fell to control level (Bes and Mench, 2008). In another experiment, TBS added to a Zn-contaminated soil restored guaiacol-peroxidase activity of the primary leaves in beans to control values (Mench et al., 2000).

Most of the plant species present on the NT plot were also observed on the surrounding uncontaminated biotopes (slag heap and ruderal zone), which were similarly firstly devoid of vegetation. Thus, colonization of the NT plot was not limited by the total metal concentration in the sediment, which is consistent with previous findings (Escarré et al., 2011). Madejon et al. (2006) showed that the application of amendments to a metal contaminated soil improved spontaneous colonization and establishment of early-successional plants. In our study, all plots were covered by dense vegetation (close to 100%) but differed in plant community composition and biodiversity descriptors. This may indicate that the amendment application was not a key factor for colonization and establishment of spontaneous vegetation but an important factor that could modify vegetation composition and dynamics. The addition of TBS or HA to the sediment influenced the dynamic process of plant colonization, the three plots showing differences in composition of the plant community. The large
increase in pH following addition of TBS to the sediment (7.4 to 10.2 in May 2002),
compared to the modest increase (7.4 to 7.7 in May 2002) in the presence of HA, may
be responsible for the main changes in plant community in the TBS plot compared to
other plots. Only spontaneous vegetation that is able to tolerate alkaline pHs could
colonize this plot. Elevated pH favours the establishment of herbaceous rather than
woody species (Skousen et al., 1994). Five years after the single application, the pH in
the TBS-treated sediment decreased to 7.7 while pHs of NT and HA plots were 7.0 and
7.3, respectively (Table 1). This decrease may be sufficient to favour the establishment
of *B. pendula* in the next few years, as two seedlings were noticed in the TBS plot.
*Calamagrostis epigejos* is known as a competitive social grass able to arrest plant
succession by forming extended dense areas (Prach, 2003). At maturity, it can reach 1.5
m in height thereby shading smaller or slowly-growing plant species (Rebele and
Lehmann, 2001). As the higher rate of *C. epigejos* was recorded on the TBS plot, this
plant would contribute to the maintenance of early stage succession vegetation on this
plot.

The highest level of plant species richness was found in the presence of HA (26
species out of 16 families) compared to the NT plot (21 species out of 14 families). In
contrast, TBS did not favour species richness as it accounted for the least value (14
species out of 9 families). This is paradoxical since the TBS plot had the lowest
concentrations of extractable Zn and Cd and the lowest sediment ecotoxicity, but the
competitive exclusion of plants by *C. epigejos* possibly outweighed effects of decreased
toxicity.

Concerning diversity indices, our H' values were much higher than those given
by Conesa et al. (2007) and Dazy et al. (2008) who studied plant communities growing
respectively on Zn-Pb-contaminated tailings and experimental plots filled with a
multicontaminated soil. However, such comparisons of H' values may be confounded
by different soil and pollution parameters as well as the time elapsed after colonization..
For Dazy et al. (2008), the survey was performed after 14 months of colonization
whereas in this study the survey was done 5 years after colonization.

The influence of metal contamination on natural springtail communities has
been documented (Bengtsson and Rundgren, 1988; Chauvat and Ponge, 2002; Gillet
and Ponge, 2002, 2003; Lock et al. 2003; Strojan, 1978). In these works, no major
adverse effects of metal contamination were recorded. Springtails were abundant even
in highly contaminated sites, which indicates the resistance of this group to toxic effect
of metals and other environmental stresses (Hopkin, 1997), exemplified by high EC₅₀
values in *Folsomia candida* reproduction test at 20 °C for Cd (590 µg.g⁻¹), Cu (700
µg.g⁻¹), Zn (900 µg.g⁻¹) and Pb (2790 µg.g⁻¹) (Sandifer and Hopkin, 1997). In our study,
neither total abundance nor biodiversity indices showed significant differences between
NT plot and CA, showing the normal colonization by springtails of the NT plot. The
addition of TBS and HA did not change the abundance or biodiversity indices, showing
no effect of these amendments on the springtail community. Some effect may have been
expected on the TBS plot as Lock et al. (2003) reported a significant negative
relationship between the metal extractable fraction (0.01 M CaCl₂ solution) and the
number of springtail species. In our study, the gut content analyses showed that
untreated and amended dredged sediments were consumed by springtails (Fig. 3).

In soils that are highly contaminated with metal(loid)s, a lower number of
bacteria than in uncontaminated soil could be expected, as a result of the harmful effects
of the contaminants on soil bacterial activity, especially bacterial respiration and soil
bacterial biomass (Renella et al., 2005; Sobolev and Begonia, 2008). In our study, the total bacterial population in the NT plot was similar to that usually found in unpolluted topsoils (Taylor et al., 2002).

The addition of TBS to the soil improved the total bacterial population, probably as a consequence of the low Cd-Zn-extractable fraction on this plot.

Bacteria play an important role in increasing or decreasing metal availability, notably in sediments (Bert et al., 2009). Bioleaching of metals is caused by sulphur-oxidising bacteria (SOB) whereas anaerobic sulphate-reducing bacteria (SRB) can precipitate metals. Neutrophilic and acidophilic SOB were found in similar numbers in all plots, showing no amendment effect. In all sediment plots, neutrophilic SOB were more numerous than acidophilic SOB, due to soil pH close to neutrality (Table 1). The moderate development of acidophilic SOB was consistent with the results of Lors et al. (2004) who reported an acidophilic sulphur-oxidising population of about $10^4$ bacteria $g^{-1}$ dry soil in a similar sediment. These results indicated that metals might be released in the case of metal sulphide availability.

Sulphate-reducing bacteria were present in the same low range in the three sediment plots, showing no amendment effect (Table 7). The low amount of SRB may be explained by the sampling performed on the first 20 cm depth layer corresponding to aerobic conditions, not favourable to the growth of SRB that are strictly anaerobic. However, during wet periods, the saturation of the soil pore system might induce anoxic conditions that would promote the size of the SRB community, leading to increased metal sulphide content.

## 5. Conclusions
In this paper, TBS and HA were investigated in an experimental metal-contaminated sediment landfill site to evaluate their effects on metal extractability, sediment ecotoxicity and on living organisms such as plants, bacteria and springtails.

The incorporation of TBS into the metal contaminated sediment decreased the extractable fraction of Cd and Zn. In contrast, HA did not decrease Cd and Zn extractability. TBS at a rate of 5% DW was effective in Cd and Zn immobilization, did not cause any nutrient imbalance or sediment ecotoxicity, and improved the bacterial activity.

However, TBS application had either no influence or a negative effect, on maximum photochemical efficiency of PSII and antioxidant enzymes in plants. The bioavailability of the compound(s) responsible for those biological responses was not decreased by TBS treatment. Even if plants successfully colonized the TBS plot, they remained under stress.

Our study revealed (i) the favourable influence of both HA and TBS compared to NT, (ii) the better influence of HA on vegetation development and physiological welfare (as measured by PSII measurements and antioxidant enzymes), (iii) the better influence of TBS on sediment ecotoxicity, metal immobilization, microbial population and Cd-Zn accumulation in *C. epigejos*. The choice of using metal immobilization with TBS rather than HA as a tool in the management of dredged sediment landfill sites will depend on the objectives of such strategy, i.e. to limit metal transfer and sediment ecotoxicity or to favour plant diversity and welfare. Realising both objectives requires a study of a mixture of both soil amendments.
654 Acknowledgements

655 This study was financially supported through the Programme de Recherches
656 Concertées (PRC) Environnement-Ecotoxicité: Meresbio, étude des mécanismes de la
657 restauration de la biodiversité sur un site contaminé par les métaux, by the Contrat de Plan
658 Etat/Région Nord-Pas-de-Calais and the European FEDER (ANR – ex.FRT 2005 – action
659 Régionale – CPER DT n°15006 Décision n°2005 – CPER – Nord - 176). We thank the
660 anonymous reviewers for improving the manuscript and B. Robinson for providing
661 language assistance.

662

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886
887 **Figure captions**

888 Figure 1.

889 Fv/Fm ratios and antioxidant enzyme activities (µmoles/min/mg proteins excepted for
890 SOD expressed as U/mg proteins) in *C. epigejos* leaves collected from control area
891 (CA), untreated plot (NT), plots treated with hydroxylapatite (HA) or Thomas basic slag
892 (TBS). Different letters indicate significant differences between plots according to post-
893 hoc Tukey HSD tests run when ANOVA was significant.

894 Figure 2.

895 Fv/Fm ratios and antioxidant enzyme activities (µmoles/min/mg proteins excepted for
896 SOD expressed as U/mg proteins) in *U. dioica* leaves collected from control area (CA),
897 untreated plot (NT), plots treated with hydroxylapatite (HA) or Thomas basic slag
898 (TBS). Different letters indicate significant differences between plots according to post-
899 hoc Tukey HSD tests run when ANOVA was significant.

900 Figure 3.

901 Distribution of gut contents among the 403 collembolan specimens collected in three
902 treatments and in the control area (CA). Letters indicate significant differences among
903 means at 0.05 level (ANOVA followed by SNK procedure).

904
<caption>Table 1</caption>

Main characteristics and (pseudo)-total metal concentrations of sediments (NT, untreated; HA, hydroxylapatite; TBS, Thomas Basic Slag) and the control area (CA).

For trace elements (except for CA) and pH, values are means of triplicate measurements and their standard deviation.

<table><thead><tr><th></th><th>CA</th><th>NT</th><th>HA</th><th>TBS</th></tr></thead><tbody><tr><td><strong>Particle size distribution (%)</strong></td><td></td><td></td><td></td><td></td></tr><tr><td>Sand</td><td>13.2</td><td>69.6</td><td>60.1</td><td>65.4</td></tr><tr><td>Silt</td><td>70.3</td><td>14</td><td>21.1</td><td>19.1</td></tr><tr><td>Clay</td><td>16.5</td><td>16.4</td><td>18.8</td><td>15.5</td></tr><tr><td>Organic C (%)</td><td>1.64</td><td>3.2</td><td>4.1</td><td>3.7</td></tr><tr><td>N (g kg<sup>-1</sup>)</td><td>1.6</td><td>1.62</td><td>1.67</td><td>2.04</td></tr><tr><td>CaCO<sub>3</sub> (g kg<sup>-1</sup>)</td><td>63</td><td>76</td><td>76</td><td>105</td></tr><tr><td>CEC (cmol kg<sup>-1</sup>)</td><td>13.3</td><td>7.2</td><td>8.3</td><td>11.3</td></tr><tr><td>pH water</td><td>7.6 ± 0.1</td><td>7.0 ± 0.1</td><td>7.3 ± 0.1</td><td>7.7 ± 0.3</td></tr><tr><td>Zn (mg kg<sup>-1</sup> DW)</td><td>75</td><td>3250 ± 319</td><td>2521 ± 461</td><td>3029 ± 217</td></tr><tr><td>Cd (mg kg<sup>-1</sup> DW)</td><td>0.4</td><td>84 ± 2</td><td>67 ± 11</td><td>86 ± 8</td></tr><tr><td>Pb (mg kg<sup>-1</sup> DW)</td><td>27</td><td>448 ± 82</td><td>389 ± 96</td><td>413 ± 57</td></tr><tr><td>Cu (mg kg<sup>-1</sup> DW)</td><td>17</td><td>93 ± 13</td><td>94 ± 11</td><td>123 ± 29</td></tr><tr><td>As (mg kg<sup>-1</sup> DW)</td><td>8</td><td>53 ± 4</td><td>48 ± 4</td><td>45 ± 2</td></tr></tbody></table>
911 Table 2

912 Concentrations of trace elements extracted from soils by 0.01 M Ca(NO₃)₂ (means of

913 triplicate values ± standard deviation) and corresponding leachate ecotoxicity.

914 Significant differences among means at 0.05 level were indicated by different letters. nr

915 = not reached

<table><thead><tr><th></th><th>NT</th><th>HA</th><th>TBS</th></tr></thead><tbody><tr><td>Zn (mg kg<sup>-1</sup> DW)</td><td>12.05a ± 0.96</td><td>7.77ab ± 0.33</td><td>2.07b ± 2.73</td></tr><tr><td>Cd</td><td>0.64a ± 0.12</td><td>0.41ab ± 0.04</td><td>0.12b ± 0.13</td></tr><tr><td>Pb</td><td>&lt;0.07</td><td>&lt;0.07</td><td>&lt;0.07</td></tr><tr><td>Cu</td><td>0.12a ± 0.01</td><td>0.13a ± 0.02</td><td>0.21a ± 0.14</td></tr><tr><td>As</td><td>&lt;0.1</td><td>&lt;0.1</td><td>&lt;0.1</td></tr><tr><td>Microtox EC<sub>50</sub> (%)</td><td>nr</td><td>nr</td><td>nr</td></tr><tr><td>Microtox (% luminescent inhibition at 90%)</td><td>44</td><td>20.5</td><td>6.3</td></tr><tr><td>Algae test IC<sub>50</sub> (%)</td><td>12.8</td><td>19.2</td><td>nr</td></tr><tr><td>Algal test (% inhibition at 90%)</td><td>100</td><td>100</td><td>36.4</td></tr></tbody></table>

916

917
918 Table 3

919 Composition of the plant community in the three treatment plots. Species richness is the
920 number of species. H' is the Shannon-Weaver index. J is the Piélou equitability index.

<table><thead><tr><th>Family</th><th>Species</th><th>Plant type</th><th>Plant life cycle</th><th>NT</th><th>HA</th><th>TBS</th></tr></thead><tbody><tr><td>Apiaceae</td><td>Pastinaca sativa</td><td>Herbaceous</td><td>Annual</td><td>0.47</td><td>0.46</td><td>0</td></tr><tr><td rowspan="3">Asteraceae</td><td>Eupatorium cannabinum</td><td>Herbaceous</td><td>Perennial</td><td>0</td><td>3.15</td><td>0.29</td></tr><tr><td>Achillea millefolium</td><td>Herbaceous</td><td>Annual</td><td>0.12</td><td>7.59</td><td>0</td></tr><tr><td>Cirsium arvense</td><td>Herbaceous</td><td>Perennial</td><td>0.47</td><td>0</td><td>1.39</td></tr><tr><td>Brassicaceae</td><td>Cardamine hirsuta</td><td>Herbaceous</td><td>Annual</td><td>6.18</td><td>1.76</td><td>0.19</td></tr><tr><td>Boraginaceae</td><td>Myosotis arvense</td><td>Herbaceous</td><td>Annual</td><td>0.12</td><td>0.09</td><td>0</td></tr><tr><td>Convolvulaceae</td><td>Convolvulus arvensis</td><td>Herbaceous</td><td>Perennial</td><td>0</td><td>0.46</td><td>0</td></tr><tr><td rowspan="2">Cyperaceae</td><td>Carex hirta</td><td>Herbaceous</td><td>Perennial</td><td>11.55</td><td>1.85</td><td>0.33</td></tr><tr><td>Carex cuprina</td><td>Herbaceous</td><td>Perennial</td><td>0.12</td><td>0.09</td><td>0</td></tr><tr><td rowspan="2">Lamiaceae</td><td>Glechoma hederacea</td><td>Herbaceous</td><td>Perennial</td><td>11.67</td><td>22.85</td><td>7.08</td></tr><tr><td>Lycopus europaeus</td><td>Herbaceous</td><td>Perennial</td><td>0.70</td><td>1.57</td><td>0</td></tr><tr><td>Onagraceae</td><td>Epilobium angustifolium</td><td>Herbaceous</td><td>Annual</td><td>0.35</td><td>1.2</td><td>11.76</td></tr><tr><td rowspan="3">Poaceae</td><td>Deschampsia cespitosa</td><td>Herbaceous</td><td>Perennial</td><td>1.52</td><td>1.76</td><td>0</td></tr><tr><td>Festuca rubra</td><td>Herbaceous</td><td>Perennial</td><td>3.73</td><td>6.85</td><td>0</td></tr><tr><td>Phalaris arundinacea</td><td>Herbaceous</td><td>Perennial</td><td>0</td><td>0.93</td><td>33.09</td></tr></tbody></table>
<table><tbody><tr><td>Calamagrostis epigejos</td><td>Herbaceous</td><td>Perennial</td><td>46.91</td><td>19.98</td><td>30.00</td></tr><tr><td>Holcus lanatus</td><td>Herbaceous</td><td>Perennial</td><td>0</td><td>2.22</td><td>0.1</td></tr><tr><td>Poa trivialis</td><td>Herbaceous</td><td>Perennial</td><td>0</td><td>0</td><td>0.14</td></tr><tr><td>Ranunculaceae</td><td>Ranunculus repens</td><td>Herbaceous</td><td>Perennial</td><td>0.12</td><td>1.11</td><td>0</td></tr><tr><td>Rubiaceae</td><td>Galium aparine</td><td>Herbaceous</td><td>Annual</td><td>2.22</td><td>2.13</td><td>16.40</td></tr><tr><td>Scrophulariaceae</td><td>Linaria vulgaris</td><td>Herbaceous</td><td>Perennial</td><td>0</td><td>1.67</td><td>0</td></tr><tr><td>Urticaceae</td><td>Urtica dioica</td><td>Herbaceous</td><td>Perennial</td><td>2.33</td><td>4.63</td><td>0.86</td></tr><tr><td>Betulaceae</td><td>Betula pendula</td><td>Tree</td><td>Perennial</td><td>8.05</td><td>12.67</td><td>0.05</td></tr><tr><td></td><td>Betula pubescens</td><td>Tree</td><td>Perennial</td><td>1.87</td><td>2.31</td><td>0.05</td></tr><tr><td>Fagaceae</td><td>Quercus robur</td><td>Tree</td><td>Perennial</td><td>0.12</td><td>0.28</td><td>0</td></tr><tr><td>Salicaceae</td><td>Salix sp.</td><td>Tree</td><td>Perennial</td><td>0</td><td>0.46</td><td>0</td></tr><tr><td></td><td>Salix alba</td><td>Tree</td><td>Perennial</td><td>0.12</td><td>1.02</td><td>0</td></tr><tr><td></td><td>Salix capraea</td><td>Tree</td><td>Perennial</td><td>1.40</td><td>0.93</td><td>0</td></tr><tr><td>Species richness</td><td></td><td></td><td>21</td><td>26</td><td>14</td></tr><tr><td>H'</td><td></td><td></td><td>2.7</td><td>3.6</td><td>2.35</td></tr><tr><td>J</td><td></td><td></td><td>0.62</td><td>0.77</td><td>0.62</td></tr></tbody></table>

For species, values were percent numbers of individuals per species and per unit

surface.

923
924 Table 4

925 Concentrations (mg/kg DW) in shoots of *Calamagrostis epigejos* and *Urtica dioica*.

<table><thead><tr><th></th><th>CA</th><th>NT</th><th>HA</th><th>TBS</th></tr></thead><tbody><tr><td colspan="5"><em>Calamagrostis epigejos</em></td></tr><tr><td>Cd</td><td>&lt; 0.05</td><td>2.16 (a) ± 1.4</td><td>2.51 (a) ± 1.1</td><td>0.34 (b) ± 0.18</td></tr><tr><td>Zn</td><td>23 (a) ± 9</td><td>121 (c) ± 58</td><td>98 (bc) ± 32.6</td><td>44 (ab) ± 20</td></tr><tr><td>Pb</td><td>&lt;1</td><td>&lt;1</td><td>&lt;1</td><td>&lt;1</td></tr><tr><td>Cu</td><td>4.5 (ab) ± 0.9</td><td>3.5 (a) ± 0.3</td><td>5.1 (b) ± 0.7</td><td>4.3 (ab) ± 0.2</td></tr><tr><td>As</td><td>&lt; 0.5</td><td>&lt; 0.5</td><td>&lt; 0.5</td><td>&lt; 0.5</td></tr><tr><td>Ca</td><td>405 ± 331</td><td>617 ± 115</td><td>769 ± 273</td><td>791 ± 243</td></tr><tr><td>Fe</td><td>23 ± 2</td><td>52 ± 56</td><td>33 ± 5</td><td>23 ± 1</td></tr><tr><td>Mg</td><td>453 (b) ± 121</td><td>251 (a) ± 83</td><td>532 (b)± 147</td><td>446 (ab) ± 62</td></tr><tr><td>Mn</td><td>16 (a) ± 5</td><td>51 (b) ± 8</td><td>13 (a) ± 3</td><td>15 (a) ± 1</td></tr><tr><td>P</td><td>1570 ± 247</td><td>1530 ± 189</td><td>1863 ± 875</td><td>1175 ± 235</td></tr><tr><td>Na</td><td>123 (ab) ± 4</td><td>115 (a) ± 4</td><td>131 (b) ± 12</td><td>119 (ab) ± 13</td></tr><tr><td>K</td><td>12407 (a) ± 1329</td><td>14523 (ab) ± 1543</td><td>16648 (b) ± 2620</td><td>13074 (a) ± 1288</td></tr><tr><td>Al</td><td>n.d</td><td>5 (a) ± 2</td><td>8.5 (b) ± 1.6</td><td>24 (c) ± 1</td></tr><tr><td colspan="5"><em>Urtica dioica</em></td></tr><tr><td>Cd</td><td>0.06 (a) ± 0.01</td><td>1.5 (b) ± 1.3</td><td>0.9 (ab) ± 0.3</td><td>0.36 (ab) ± 0.18</td></tr><tr><td>Zn</td><td>12 (a) ± 1</td><td>210 (b) ± 104</td><td>184 (b) ± 28</td><td>145 (b) ± 21</td></tr><tr><td>Pb</td><td>&lt;1</td><td>&lt;1</td><td>&lt;1</td><td>&lt;1</td></tr><tr><td>Cu</td><td>4.0 (a) ± 0.5</td><td>6.9 (b) ± 1.9</td><td>6.8 (b) ± 1.9</td><td>7.4 (b) ± 0.9</td></tr><tr><td>As</td><td>&lt;0.5</td><td>&lt;0.5</td><td>&lt;0.5</td><td>&lt;0.5</td></tr><tr><td>Ca</td><td>19719 ± 8227</td><td>20945 ± 3292</td><td>24398 ± 2407</td><td>23901 ± 3334</td></tr><tr><td>Fe</td><td>35 (a) ± 6</td><td>51 (ab) ± 10</td><td>60 (b) ± 12</td><td>43 (ab) ± 11</td></tr><tr><td>Mg</td><td>1726 ± 1753</td><td>1278 ± 433</td><td>2176 ± 285</td><td>1722 ± 241</td></tr><tr><td>Mn</td><td>13 ± 3</td><td>12 ± 6</td><td>11 ± 2</td><td>12 ± 2</td></tr><tr><td>P</td><td>2314 (a) ± 267</td><td>4229 (b) ± 465</td><td>4310 (b) ± 841</td><td>6096 (c) ± 563</td></tr><tr><td>Na</td><td>136 ± 38</td><td>177 ± 41</td><td>162 ± 20</td><td>149 ± 12</td></tr><tr><td>K</td><td>17355 (a) ± 4593</td><td>19029 (ab) ± 1434</td><td>23699 (bc) ± 1721</td><td>27963 (c) ± 2117</td></tr><tr><td>Al</td><td>n.d</td><td>20 ± 8</td><td>32 ± 10</td><td>28 ± 12</td></tr></tbody></table>

926 Mean values and standard deviations (n=5; mg kg⁻¹ DW). Different letters indicate

927 significant differences among plots for α =0.05.

928
<table><caption>**Table 5.** Correlation analysis between shoot element concentrations and antioxidant enzyme activities in leaves (n=20).</caption><thead><tr><th colspan="6">C. epigejos</th></tr><tr><th></th><th>SOD</th><th>CAT</th><th>APX</th><th>GPX</th><th>GRD</th></tr></thead><tbody><tr><td>Cd</td><td>-0.68**</td><td>-0.67**</td><td>-0.79***</td><td>-0.61**</td><td>-0.58**</td></tr><tr><td>Zn</td><td>0.31ns</td><td>0.35ns</td><td>0.56**</td><td>0.01ns</td><td>0.52*</td></tr><tr><td>Cu</td><td>-0.27ns</td><td>-0.48*</td><td>-0.38ns</td><td>0.00ns</td><td>-0.56*</td></tr><tr><td>Mn</td><td>0.55*</td><td>0.60**</td><td>0.53*</td><td>-0.11ns</td><td>0.80***</td></tr><tr><td>P</td><td>0.09ns</td><td>-0.02ns</td><td>-0.18ns</td><td>0.13ns</td><td>-0.23ns</td></tr><tr><td>Ca</td><td>0.37ns</td><td>0.18ns</td><td>0.38ns</td><td>0.44ns</td><td>0.08ns</td></tr><tr><td>K</td><td>0.06ns</td><td>-0.01ns</td><td>0.22ns</td><td>-0.03ns</td><td>0.14ns</td></tr><tr><td>Mg</td><td>-0.30ns</td><td>-0.43ns</td><td>-0.52*</td><td>-0.32ns</td><td>-0.56*</td></tr><tr><td>Na</td><td>-0.33ns</td><td>-0.54*</td><td>-0.19ns</td><td>-0.21ns</td><td>-0.50*</td></tr></tbody><tbody><tr><th colspan="6">U. dioica</th></tr><tr><td>Cd</td><td>0.22ns</td><td>0.24ns</td><td>0.30ns</td><td>0.32ns</td><td>0.11ns</td></tr><tr><td>Zn</td><td>0.48*</td><td>0.47*</td><td>0.56**</td><td>0.61**</td><td>0.45*</td></tr><tr><td>Cu</td><td>0.51*</td><td>0.49*</td><td>0.56*</td><td>0.60**</td><td>0.70**</td></tr><tr><td>Mn</td><td>-0.07ns</td><td>-0.02ns</td><td>-0.08ns</td><td>-0.04ns</td><td>0.08ns</td></tr><tr><td>P</td><td>0.79***</td><td>0.74***</td><td>0.79***</td><td>0.85***</td><td>0.82***</td></tr><tr><td>Ca</td><td>0.10ns</td><td>0.12ns</td><td>0.18ns</td><td>0.21ns</td><td>0.26ns</td></tr><tr><td>K</td><td>0.54*</td><td>0.45*</td><td>0.48*</td><td>0.56*</td><td>0.60**</td></tr><tr><td>Mg</td><td>-0.20ns</td><td>-0.21ns</td><td>-0.10ns</td><td>-0.07ns</td><td>-0.11ns</td></tr><tr><td>Na</td><td>0.22ns</td><td>0.24ns</td><td>0.28ns</td><td>0.22ns</td><td>0.02ns</td></tr></tbody></table>

ns: not significant, ***, **, *, significant at a probability level P< 0.001, 0.01, 0.05, respectively.
<table><caption>**Table 6.** Species distribution and biodiversity indices of collembolan species.</caption><thead><tr><th></th><th>CA</th><th>NT</th><th>HA</th><th>TBS</th><th>F value</th></tr></thead><tbody><tr><td>Dicyrtomina minuta</td><td>0</td><td>0</td><td>0</td><td>0.33±0.27</td><td>1 NS</td></tr><tr><td>Entomobrya sp.</td><td>0</td><td>1.33±1.09</td><td>0</td><td>0</td><td>1 NS</td></tr><tr><td>Folsomia candida</td><td>0.33±0.27</td><td>0.33±0.27</td><td>0</td><td>1±0.82</td><td>0.58 NS</td></tr><tr><td>Friesea truncata</td><td>0<sup>b</sup></td><td>0<sup>b</sup></td><td>0<sup>b</sup></td><td>4±1.25<sup>a</sup></td><td>6.86**</td></tr><tr><td>Hererosminthurus claviger</td><td>0.33±0.27</td><td>0</td><td>0</td><td>0</td><td>1 NS</td></tr><tr><td>Isotomodes productus</td><td>0.67±0.54</td><td>0</td><td>0</td><td>2.67±1.78</td><td>1.22 NS</td></tr><tr><td>Lepidocytus lignorum</td><td>0</td><td>0</td><td>0</td><td>0.33±0.27</td><td>1 NS</td></tr><tr><td>Mesaphorura florae</td><td>1.33±0.72</td><td>6.33±2.13</td><td>21.67±7.92</td><td>25.33±8.57</td><td>2.56 NS</td></tr><tr><td>Mesaphorura macrochaeta</td><td>3.33±2.72</td><td>0</td><td>0.67±0.27</td><td>0.67±0.54</td><td>0.75 NS</td></tr><tr><td>Mesaphorura pongei</td><td>0</td><td>0</td><td>0</td><td>1.33±1.09</td><td>1 NS</td></tr><tr><td>Orchesella sp.</td><td>0</td><td>0.33±0.27</td><td>0.33±0.27</td><td>0</td><td>0.67 NS</td></tr><tr><td>Parisotoma notabilis</td><td>6.67±1.96</td><td>4.67±1.66</td><td>5.67±2.13</td><td>0.33±0.27</td><td>1.85 NS</td></tr><tr><td>Proisotoma minima</td><td>7±2.45</td><td>9±2.94</td><td>0.67±0.54</td><td>7±2.36</td><td>1.71 NS</td></tr><tr><td>Tomocerus vulgaris</td><td>0</td><td>0</td><td>0.33±0.27</td><td>0</td><td>1 NS</td></tr><tr><td><strong>Total abundance</strong></td><td>19.67±3.07</td><td>22±5.44</td><td>29.33±6.4</td><td>43±10.6</td><td>1.53 NS</td></tr><tr><td><strong>Species richness</strong></td><td>4±0.47</td><td>4</td><td>3.67±0.27</td><td>5.67±1.19</td><td>1.28 NS</td></tr><tr><td><strong>Shannon Index</strong></td><td>1.44±0.21</td><td>1.65±0.07</td><td>1.09±0.32</td><td>1.78±0.26</td><td>1.09 NS</td></tr><tr><td><strong>Equitability Index</strong></td><td>0.72±0.04</td><td>0.82±0.04</td><td>0.56±0.15</td><td>0.75±0.03</td><td>1.25 NS</td></tr></tbody></table>

ANOVA: NS = not significant; ** P < 0.01. Post-hoc tests (SNK procedure):

significant differences indicated by different letters

943

944
945

946

Table 7. Total and specific bacterial numbers

947

948

949

950

951

952

953

Counts of total bacteria, acidophilic sulphur-oxidising bacteria (ASOB), neutrophilic sulphur-oxidising bacteria (NSOB) and sulphate-reducing bacteria (SRB) (expressed as bacteria g⁻¹ dry soil) collected from untreated plot (NT), plot treated with hydroxylapatite (HA) or Thomas basic slag (TBS). Values are expressed as means ± standard deviation in triplicate samples. Values with the same letters are not significantly different according to ANOVA used Newman-Keuls (SNK) test at P ≤ 0.05.

<table><thead><tr><th></th><th>NT</th><th>HA</th><th>TBS</th></tr></thead><tbody><tr><td><strong>Total bacteria</strong></td><td>2.12 10<sup>7</sup><br>± 5.09 10<sup>5</sup> c</td><td>5.58 10<sup>7</sup><br>± 7.90 10<sup>6</sup> b</td><td>1.75 10<sup>8</sup><br>± 1.8610<sup>6</sup> a</td></tr><tr><td><strong>ASOB</strong></td><td>3.69 10<sup>3</sup><br>± 2.31 10<sup>1</sup> a</td><td>7.36 10<sup>3</sup><br>± 3.19 10<sup>3</sup> a</td><td>2.93 10<sup>3</sup><br>± 1.29 10<sup>3</sup> a</td></tr><tr><td><strong>NSOB</strong></td><td>1.00 10<sup>5</sup><br>± 6.78 10<sup>4</sup> a</td><td>3.05 10<sup>5</sup><br>± 1.15 10<sup>5</sup> a</td><td>1.45 10<sup>5</sup><br>± 4.62 10<sup>4</sup> a</td></tr><tr><td><strong>SRB</strong></td><td>6.97 10<sup>1</sup><br>± 4.04 10<sup>1</sup> a</td><td>5.97 10<sup>1</sup><br>± 2.8910<sup>1</sup> a</td><td>5.97 10<sup>1</sup><br>± 2.89 10<sup>1</sup> a</td></tr></tbody></table>

954

955
956 Table 8

957 Treatments (NT = Untreated, HA = Hydroxyapatite, TBS = Thomas Basic Slag) ranked
958 according to 24 variables describing optimal features of plant, faunal and microbial
959 communities and sediment toxicity. Average ranks are indicated by mean ± standard
deviation.
960

<table><thead><tr><th>Variable</th><th>NT</th><th>HA</th><th>TBS</th></tr></thead><tbody><tr><td>Plant, faunal and microbial communities good health</td><td></td><td></td><td></td></tr><tr><td>Vegetation succession state</td><td>2</td><td>1</td><td>3</td></tr><tr><td>Plant species richness</td><td>2</td><td>1</td><td>3</td></tr><tr><td>Plant species diversity</td><td>2</td><td>1</td><td>3</td></tr><tr><td>Plant species equitability</td><td>2.5</td><td>1</td><td>2.5</td></tr><tr><td>Fv/Fm in U. dioica</td><td>2</td><td>1</td><td>3</td></tr><tr><td>Collembola abundance</td><td>3</td><td>2</td><td>1</td></tr><tr><td>Collembola species richness</td><td>2</td><td>3</td><td>1</td></tr><tr><td>Collembola species diversity</td><td>2</td><td>3</td><td>1</td></tr><tr><td>Collembola species equitability</td><td>1</td><td>3</td><td>2</td></tr><tr><td>Total bacteria abundance</td><td>3</td><td>2</td><td>1</td></tr><tr><td>Sediment toxicity</td><td></td><td></td><td></td></tr><tr><td>Cd mobility</td><td>3</td><td>2</td><td>1</td></tr><tr><td>Zn mobility</td><td>3</td><td>2</td><td>1</td></tr><tr><td>Microtox</td><td>3</td><td>2</td><td>1</td></tr><tr><td>Algal test</td><td>2.5</td><td>2.5</td><td>1</td></tr><tr><td>Cd in C. epigejos</td><td>2</td><td>3</td><td>1</td></tr><tr><td>Zn in C. epigejos</td><td>3</td><td>2</td><td>1</td></tr><tr><td>Cd in Urtica</td><td>3</td><td>2</td><td>1</td></tr><tr><td>Zn in Urtica</td><td>3</td><td>2</td><td>1</td></tr><tr><td>SOD in C. epigejos</td><td>3</td><td>1</td><td>2</td></tr><tr><td>CAT in C. epigejos</td><td>3</td><td>1</td><td>2</td></tr><tr><td>APX in C. epigejos</td><td>3</td><td>1</td><td>2</td></tr><tr><td>SOD in U. dioica</td><td>2</td><td>1</td><td>3</td></tr><tr><td>CAT in U. dioica</td><td>2</td><td>1</td><td>3</td></tr><tr><td>APX in U. dioica</td><td>2</td><td>1</td><td>3</td></tr><tr><td>Average rank</td><td>2.46±0.57</td><td>1.73±0.76</td><td>1.81±0.9</td></tr></tbody></table>

961

962
<figure><img src="image_1.png" /></figure>
<figure><img src="image_2.png" /></figure>
<figure><img src="image_3.png" /></figure>

970

971

972

973