The state of Hidalgo contributes to the national production with 0.5 % of Cu, 96.7 % of Mn, 2.7 % of Pb, and 2.8 % of Zn (SGM 2017). The municipality of Zimapán is a mining area representative of this state which produces undesired environmental impact of heavy metal in the residues and which could be of pronounced economic importance. This area presents a skarn of the metal type Zn-Pb-Ag-(Cu) in the form of sulfurous minerals including pyrite, arsenopyrite, sphalerite, galena, and others (Villaseñor-Cabral et al. 2000, Espinosa et al. 2009). Residues in these mines contains heterogeneous concentrations of arsenic (13 135 mg/kg), cadmium (610 mg/kg), copper (600 mg/kg), lead (3934 mg/kg), and zinc (11 363 mg/kg) (Armienta et al. 2016). After 70 years of accumulation of mine residues at this site, a significant volume has been generated and discharged in tailings ponds (Espinosa et al. 2009). If these residues get exposed to wind and rain the potential for dispersing and to contaminate the surroundings exist. Therefore, it is necessary to stabilize the mining residues for avoiding the chemical degradation of the environment.

In situ remediation techniques are utilized to stabilize mining residues, where the objective is not to change the total concentration of these metals, but to reduce the available fraction (Adriano et al. 2004). The most promising remediation techniques include the application of lime (Bolan et al. 2003), phosphates (Basta and McGowen 2004, Cui et al. 2016), biosolids (Wang et al. 2008, Placek et al. 2016), compost (Smith 2009), and more recently biochar amendments (Beesley et al. 2015, Mahar et al. 2015, Yuan et al. 2019). The application of lime initially increases pH, reduces the solubility of metals and can also mix with compost; in addition, is a low-cost material easily accessible and applied, but organic matter is transient reducing effectiveness afterwards (Gray et al. 2006, Kumpiene et al. 2008, Singh and Kalamdhad 2013). The application of phosphates forms stabilized precipitates of metal-phosphates and provides essential nutritional elements for the growth of the plant cover but they can cause leaching (Cao et al. 2009, Bolan et al. 2014, Osborne et al. 2015). Walker et al. (2004), and Singh and Agrawal (2008), have shown that the application of biosolids and compost decreases the bioavailability of metals, but their effect is variable depending on the metal, soil type, dose, and degree of organic matter humification. Most of these amendments require periodic applications and pre-treatments, which increase the application costs to ensure their success (Almas et al. 1999, Tandy et al. 2009, Cui et al. 2016, Gong et al. 2018). Biochar, on the other hand, is a solid product of biomass pyrolysis. It increases the recalcitrant organic carbon content of soil in the long term and requires a smaller number of applications compared to compost and biosolids. In addition, biochar is a porous material (Batista et al. 2018), presenting large specific surfaces for sorption of metals (Houben et al. 2013, Zhang et al. 2013, Wang et al. 2017, Wang et al. 2018, Yuan et al. 2019), improving soil physical properties (Tang et al. 2013, Bordoloi et al. 2019), and its pH value is normally around 5 to 12 (Yuan et al. 2019). Currently, there are studies where biochar has been modified by the addition of alkalis, oxidants (as O₃, H₂O₂, K₂MnO₄ and air), microwaves, CO₂ and steam to improve its sorption capacity (Zhang et al. 2016, Yuan et al. 2019). Nevertheless, it has positive and negative effects, depending on the method of activation, kind of bioassay and kind of soil (Koltowski et al. 2017).

The production and use of biochar from plants with high growth rates, present areas of opportunity to remediate acid residues generated by mining activity in Mexico. The water hyacinth has growth rates of 100-120 Mg/ha/year (Masto et al. 2013), besides its biomass possesses a strong adsorption capacity due to its high cellulose content and functional groups as carboxyl and hydroxyl (Patel 2012, Sindhu et al. 2017). For this reason, it has been used in wetlands, in solid dry form, like biochar, to remove toxic metals from aqueous solutions, wastewater and effluent treatments (Rezania et al. 2015, Sarkar et al. 2017, Neris et al. 2019).

Water hyacinth (Eichhornia crassipes Mart) is an invasive plant that for decades has affected the Endhó and Requena dams located in the municipalities of Tula, Alfajayucan, and Tepejí del Río, in the state of Hidalgo, México (Romero 1989, Soto 1989). It causes problems including an increase in sedimentation, canal blockages, invasion of water bodies, and competition with neighboring species, thereby decreasing biodiversity (Sindhu et al. 2017). This weed thrives in water bodies with high nutrient content and control can be manual, mechanic (by dredging or with a harvester machine), chemical (through herbicides), and biological (with herbivorous carps or insects). If this invasive plant is not removed from where it grows, it has the potential to recover with water availability. In addition, water hyacinth has large quantities of viable seeds that can germinate in the rainy season (Gutiérrez et al. 1994). The conversion of E. crassipes into biochar can represent a method for its management as weed and a use in the remediation of acid residues and soils contaminated with these, because it offers the possibility of neutralizing them and a greater permanence in the soil due to high resistance to microbial decomposition (Berek and Hue 2016, Li et al. 2016, Dai et al. 2017, Wang et al. 2018).

There are several studies that support the use of water hyacinth biochar as an amendment in soils or in mine residue’s remediation and polluted soils. In soils, its addiction increases the activity of active microbial biomass, soil respiration, the germination percentage and the shoots length of corn even with doses of 10 and 20 % (Masto et al. 2013); it also decreases cracking and increases water holding capacity when applied at 10 % (Bordoloi et al. 2019). For remediation it has an adsorption capacity for Zn and Pb of 22 and 45 % respectively, when is prepared with high pyrolysis temperature and at a 10 % dose (Wang et al. 2017); Cd is removed when is applied in a 2 % dose in a multimetal contaminated soil while As and Pb has little mobilization under acid precipitation (Yin et al. 2016). However, it has also shown the ability to remove aqueous anions such as phosphate (Cai et al. 2017) and arsenate (V) (Zhan et al. 2016) when it is magnetized by co-precipitation with Fe²⁺/Fe³⁺ on water hyacinth biomass before pyrolysis.

The objectives of this work were: a) to apply and evaluate biochar derived from water hyacinth (H) in mining acid residues; and b) to compare its performance with monopotassium phosphate (F = KH₂PO₄), lime [L = Ca(OH)₂] and phosphate mixtures with biochar (FH) and lime (FL) by a bioassay of barley root growth, soluble metal and pH.

The texture of the soil (M1) was clayey and its pH-value, neutral, while mine residues (M4) sieved at 2 mm showed an acidic pH-value (Table I), while the pH-value of control substrates decreased below to M1. Acid mining residues added to the substrates caused an increase in acidity and salt content. These effects were attributed to in situ oxidation and residual sulfides (> 11 % acid drainage generators) in the mining waste (Guzman 2012, Labastida et al. 2013, Armienta et al. 2016). The electrical conductivity and the soluble Cu and Zn content, increased from M2 to M4. The removable bases, the CEC-values and the soluble Pb concentrations in the M2-substrate increased but in M3 decreased (Table I). Lead concentration in M4 was greater than 400 mg/kg, concentration that is above the standard of the Official Mexican Standard NOM-147-SEMARNAT/SSA1-2004 (SEMARNAT 2007), even though the soluble Pb content in all substrates was lower than the permissible limit of 5 mg/L of the Official Mexican Standard NOM-052-SEMARNAT-2005 (SEMARNAT 2006)., In Mexico does not exist a reference concentration considered dangerous for Cu and Zn. In contrast, the United States Environmental Protection Agency considers phytotoxic concentrations of 1500 and 2800 mg/kg for these metals (USEPA 1995). Only Cu was not toxic when increasing the percentage of mine acid residues in the substrates (Table I). When the percentage of soil was increased, the soluble concentration of these metals decreased due to the dissolution of carbonates present in the soil and mine residues (Labastida et al. 2013).

TABLE I

Variable	Unit	M1	M2	M3	M4	H
pH		6.8±0.2	6.6±0.2	4.7±0.1	3.3±0.1	10.2±0.0
EC1:20	µS/m	334.5±25	415±21	468±70	481.0±43	604±32
CEC	cmol(-)/kg	27±3	46±1	24±6	22±5	42±2
RB	cmol(+)/kg	39±1	50±1	21±0	18±1	424±3
CuT	mg/kg	76±11	210±20	324±29	457±38	199±33
PbT	mg/kg	173±3	2384±3	4279±3	6490±3	< LD
ZnT	mg/kg	379±11	1090±18	1699±24	2409±31	115±20
Cu soluble	mg/kg	< LD	0±0	1±0	11±0	1±0
	mmol/kg	< LD	5±1	10±1	169±4	19±1
Pb soluble	mg/kg	39±5	64±2	29±2	48±6.0	< LD
	mmol/kg	187±23	310±12	142±12	230±27	< LD
Zn soluble	mg/kg	< LD	< LD	5±0	29±2	< LD
	mmol/kg	< LD	< LD	82±0	449±35	< LD

LD = limit of detection for metals: total Cu 10 mg/kg, Pb 5 mg/kg and Zn 8 mg/kg; soluble Pb 1 mg/kg; Zn 0.1 mg/kg and Cu 1 mg/kg. Where: M1 is 100 % soil; M2 is 35:65 % (M1:M4); M3 is 65:35 % (M1:M4); M4 is 100 % acidic mining residues; EC1:20 is the electrical conductivity measured in a 1:20 ratio; CEC is the cation exchange capacity; RB are the removable bases; and CuT, PbT, and ZnT are the total metals

The application of H decreased the bulk density, the particle density (except for M3), the field capacity and the humidity at saturation (in M2, M3, and M4) and increased the pore space (Table II). Batista et al. (2018) determined that H (pyrolyzed at 350 ºC) had a high field capacity due to the porosity, the CEC, and the specific surface area. However, our results were opposite due to the pyrolysis temperature.

TABLE II

Substrate	Bulk density	Particle density	Field capacity	Permanent wilting point	Available humidity	Humidity saturation	Porous space*
	g/cm3			cm3/cm3			%
M1	1.18 b	2.43 c	0.43 a	0.23 a	0.20 ab	0.72 a	51.5 ef
M1H	0.85 e	2.25 d	0.39 ab	0.20 b	0.19 ab	0.67 b	62.0 a
M2	1.13 bc	2.43 c	0.32 de	0.14 c	0.18 abc	0.57 d	53.2 de
M2H	0.97 d	2.32 d	0.38 bc	0.17 c	0.22 a	0.60 c	58.2 bc
M3	1.29 a	2.52 b	0.29 ef	0.15 c	0.14 cd	0.49 f	48.8 f
M3H	1.08 c	2.45 bc	0.35 cd	0.17 c	0.18 abc	0.61 c	56.1 cd
M4	1.34 a	2.63 a	0.21 g	0.09 d	0.12 d	0.47 f	49.2 f
M4H	0.99 d	2.5 bc	0.26 f	0.10 d	0.16 bcd	0.53 e	60.6 ab

Different letters in the same column indicate significant differences (p < 0.05)

The CEC of biochar measured by AgTU-method is reported (Table I) without removal of carbonates and soluble salts. Singh et al. (2010) recommend measuring the CEC with the same method but with previous removal of salts, because Ag can precipitate with sulfides in pH > 8 and overestimate this measure. However, Doumer et al. (2016) and Batista et al. (2018) reported 37 cmol^(-)/kg for H (pyrolyzed at 350 ºC) measured with ammonium acetate and barium acetate, a value close to the obtained in this work. The electrical conductivity and high alkaline pH value of the H was like those found by Singh et al. (2010), and Berek and Hue (2016).

The liming potential of the H was equivalent to 16.4 g/kg CaCO₃. The equivalent dose necessary to correct the acidity of the mine residues and bring it to pH 6.5 was 5 % for H, as shown by the curves of neutralization of the biochars (Fig. 1A). This coincides with the dose found by Wang et al. (2017) who used water hyacinth (pyrolyzed at 500 ºC) to fix Zn and Pb while Houben et al. (2013) used 10 % (w/w) of Miscanthus giganteus biochar, pyrolyzed at 600 ºC, to reduce the activity of the metals from mining waste. The acid neutralization capacity of H can be attributed to lower aromaticity and a higher abundance of carboxylic groups (Doumer et al. 2016). Figure 1B shows a pH stabilization time for H of 6 days and the proportion of particles small than 0.148 mm of 30 %. Figure 2A shows the interpolated lime dose to correct the acidity of M4 up to pH 6.5, which was 9.2 mmol of OH^- for 10 g of M4, equivalent to 34 g of Ca(OH)₂/kg of the residues. The pH stabilization time was 34 days, a longer time in comparison to that required when biochar was used as a neutralizing agent (Fig. 2B). Biochar stabilized the pH of the residues in less time than lime. The differences in time of stabilization of the M4 pH when using the biochar can be attributed to multiple factors, such as the type of biomass, the particle size and the dose, and the pyrolysis temperature, among others (Tang et al. 2013, Zaccheo et al. 2014, Wang et al. 2017). Jones et al. (2012) indicated that the neutralization potential of a wood mixture biochar (pyrolyzed at 450 ºC), was suppressed after three years of being added to the soil, mainly due to the leaching and loss of alkaline elements. Similarly, Ruttens et al. (2010) reported that over time calcareous amendments can be dissolved and leached in the soil. On the other hand, Wong et al. (1998) showed that mineral residues with pH < 3 had no potential acidity formation due to the high content of oxidized sulfurous minerals. In our case, the potential acidification value of the acidic mineral residues was zero (Cruz R. E. pers. comm. 2016), therefore the loss of liming material rarely occurs.

Fig. 1

Fig. 2

In order to define a critical reference level (p < 0.01) for root growth and use it as an indicator of phytotoxicity and treatment efficacy, the mean root length of the pristine soil (M1) was employed, this was considered to be the minimum mitigation value to be achieved. All treatments raised the pH to the M1 with the exception of M2F, M3, M3F, M4 and M4F. The root length was increased in all treatments except for M4 and M4F due to the high phytotoxic chemical activities of the H⁺ ions and the high levels of soluble Cu and Zn. This high concentration of H⁺ ions is caused by the hydrolytic oxidation of minerals such as pyrite (FeS₂), which causes the release of free sulfuric acid (Tordoff et al. 2000). The soluble Pb was increased in M2 and M2F above M1 (Fig. 3).

Fig. 3

Addition of H, FH and FL treatments to M4-substrate increased the root length and reduced the soluble Cu and Zn compared to the control (M4H, M4FL, M4FH > M1, M4L > M4F, M4) (see Fig. 3). Response of these treatments was equal to or greater than that of M1, due on the one hand to the increase in pH, which favored the decrease in the chemical activity of H⁺ ions and the phytoxicity caused mainly by the sorption of metals in the solid phase (Dai et al. 2017).

The presence of soil in 2/3 and 1/3 parts for M3 and M2 substrates was enough to mitigate the phytotoxic effect because results of root growth were equal to the treatments. For M3-substrate all treatments showed performances above the pristine soil reference level (M3, M3L, M3F, M3H, M3FL > M3FH > M1) while in M2-substrate only the control (M2, M2L, M2F, M2FL > M2H, M2FH, M1). According to figure 3, in M1 no soluble Cu was detected, but was present in the treatments of substrate (M3), supporting the positive contribution of this metal to plant response. Lime reduced soluble Pb (M3L and M2L) and soluble Zn were reduced with L, FL, H and FH. Table III shows the correlation coefficient between the pH and the concentration of the soluble metal, and between the soluble metal and the root length. The negative correlation coefficients are related to the increase in extractability of the soluble metal at low pH-values (less than 4), while reducing root growth. These relationships are indicative of phytotoxicity/phytoavailability levels of the Cu and Zn in the treatments studied. The values of the correlation coefficients do not allow to conclude that the increase of root length for substrate M3 was mainly due to the decrease in the phytotoxicity of soluble Cu, but to be phytoavailability.

TABLE III

Metal	pH-soluble metal			Soluble metal-root length
	Correlation coefficient	p-value	Correlation coefficient	p-value
Cu	-0.682	<0.001	-0.641	<0.001
Pb	-0.272	0.146	-0.070	0.071
Zn	-0.790	<0.001	-0.724	<0.001

For the M2-substrate, the H and FH treatments generated less growth of the root compared to M2. The pH of treatments with neutralizing power (L, H, FL and FH) was higher than 7 and the concentration of Pb was also lower than when lime was added (pH = 5.9); Zn was not detected in all treatments of this substrate. The increase in root length promoted by the different treatments of substrates M2 and M3 relative to the reference materials M1 and M4 was expected because on one hand, soils, and treatments with biochars (Lu et al. 2015, Jiang et al. 2016, Wang et al. 2017), lime (Bolan et al. 2003) and phosphate (Cao et al. 2009, Osborne et al. 2015) naturally tend to reduce mobility and mitigate at some degree the phytotoxicity of heavy metals such as Cu, Pb, and Zn. Adriano et al. (2004)) reported that when the amendment was applied new free organic and inorganic functional groups are available for tfor the complexation and sorption of metals, which would explain the higher root length.

On M1-substrate, the application of phosphates promoted the root growth when compared with the rest of the treatments (M1F>M1,M1H, M1FH), and the treatments with H and FH were not significantly different to M1 (M1F > M1, M1H, M1FH). Karami et al. (2011) proved that wood biochar applied at a dose of 20 % (v/v) reduced the availability of phosphorus and Mosa et al. (2018) reported phosphate sorption on water hyacinth 3biochar. When phosphate was applied together with H to the substrates, produced the same effect on growth root and soluble Cu and Zn, that when they were applied for separately. The value of pH was greater with H than with FH in M2, M3 y M4, whereas the soluble Pb content was 60 % and 88 % lower with H and with L, respectively, for the substrate M2.

Water hyacinth biochar showed to be an effective amendment to reduce the solubility of Cu and Zn and to neutralize the pH in less time than limeand it also increased the porous space of the mining residues. The reduction of soluble Pb with lime was greater than biochar in the mixtures soil:residues. However, the biochar, as the other amendments, caused root length increase when the residues were above of 65 % (M3 to M4). Its effects on pH and soluble Cu and Zn were comparable with lime, phosphate-lime and phosphate-biochar. When water hyacinth biochar was applied to pristine soil its effect on the evaluated variables was not evident.