3.1. Cd Sorption on the Individual Materials
As shown in Figure 1
a, Cd sorption was clearly higher on the forest soil sample (with a maximum of 32.4 mmol kg−1
) than on the vineyard soil sample and pyritic material. Expressed as percentage of Cd retained (referred to the concentration added), the amount sorbed decreased gradually with the increase of added Cd: from 92 to 54% for forest soil, from 67 to 39% for vineyard soil, and from 26 to 12% for the pyritic material.
Regarding by-products, Figure 1
b shows that the highest Cd sorption corresponded to oak ash, mussel shell, and hemp waste (maximum sorption close to 60 mmol kg−1
), whereas it was clearly lower for pine bark. Expressed as percentage, sorption was always >99% for oak ash and mussel shell, between 92 and 96% for hemp waste, and between 46 and 96% for pine bark.
Overall, the values of the sorption curves were better adjusted to the Freundlich than to the Langmuir model (Table 1
), meaning that Cd saturation would be hardly reached. Mussel shell, hemp waste and oak ash cannot be fitted to either model due to too high errors of estimation, which is frequent when sorption values are as high as those found for Cd on these materials.
In the present study, the lowest Cd sorption corresponded to three of the materials having the lowest pH values in the equilibrium solutions: pyritic material (pH between 3.21 and 3.39), vineyard soil (between 3.45 and 3.72), and pine bark (between 3.61 and 4.0), although forest soil sorbed remarkable Cd concentrations even if it had clearly acidic pH values in the equilibrium solutions (between 3.42 and 3.62). However, the pH values were much higher for the remaining materials: between 6.45 and 6.61 for mussel shell, between 8.09 and 8.48 for hemp waste, and between 12.83 and 13.01 for oak ash. It is obvious that those high pH values facilitate precipitation, which is considered within the sorption processes, although it is not adsorption.
According to Appel and Ma [53
], and to Kim et al. [54
], the main factors influencing Cd retention in soils are total Cd concentration and pH, which affects hydrolysis of the elements, organic matter solubility, and surface charge of the variable charge compounds [55
]. In acid soils, Cd is easily exchanged and available to plants, whereas, as pH increases, Cd retention is favored by sorption on variable charge compounds, by inner-sphere complexes formation, and by hydroxide precipitation [56
]. In this way, Memon et al. [58
] obtained maximum Cd adsorption on sawdust at pH > 4 because at those pH carboxyl groups are deprotonated and negatively charged, being able to electrostatically bind Cd (which at those pH, and up to pH 8, is found as Cd2+
), while at pH > 9 adsorption decreases due to hydrolysis of the metal, appearing CdOH+
, with lower affinity for sorbent surfaces. At pH < 3, variable charge surfaces tend to be positively charged, decreasing adsorption of cationic metals. In fact, Cd is mainly found as Cd2+
in the soil solution, although it can form complex ions, such as CdCl+
], and in contaminated soils the predominant soluble Cd species are the free Cd2+
ion and neutral species, such as CdSO4
, present in increasing quantities at pH > 6.5 [60
In the present study, correlation between pH in the equilibrium solution and sorbed Cd was significant only in the case of pine bark (coefficient of correlation r = −0.930, p
< 0.05). The fact that the variation of pH with the addition of Cd depends on the type of sorbent may be related to the sorption of the metal taking place by different mechanisms. When the dominant mechanism is the electrostatic attraction between the surface of the negatively charged bio-sorbent and Cd2+
, an exchange with H+
can take place, decreasing solution pH [61
]. According to Memon et al. [58
], the ion exchange mechanism could be the most frequent in Cd adsorption on organic materials (such as the hemp waste and pine bark in the present study), since its cell walls are formed by cellulose and lignin, with many hydroxyl groups present in tannins and other phenols (see Supplementary Material
), which are active ion exchangers. However, Taty-Costodes et al. [61
] and Pagnanelli et al. [62
], also indicate the presence of mechanisms other than cation exchange, such as specific adsorption and complexation processes, as well as physical adsorption, and probable micro-precipitations, which would explain the differences found in the present study in relation to pH change, since these processes do not imply H+
Regarding dissolved organic carbon (DOC) levels in equilibrium solutions, in the present study there was a tendency for DOC to decrease as the added Cd concentration increased, although it was only significant for forest soil (r = −0.969, p
< 0.01), vineyard soil (r = −0.879, p
< 0.05), and pine bark (r = −0.992, p
< 0.01). These DOC decreases may be due to the high affinity of cationic metals to form organic complexes [63
] and their subsequent precipitation. According to Park et al. [64
], the availability of metals is reduced due to their adsorption on solid surfaces, and to the formation of stable complexes with humic substances. In this sense, several studies identify organic matter as one of the main components controlling Cd distribution in soils [65
], indicating that Cd adsorption decreases when soil organic matter content is reduced. Furthermore, organic Cd complexes are not very stable, and dissociate when pH is <6 [67
On the other hand, the forest and vineyard soils in the present study had a 15% clay fraction, and they both (as well as the other materials studied) had relevant concentrations of non-crystalline oxides (especially oak ash, the pyritic material, and both soils, see Supplementary Material
). The clay minerals represent an important contribution of negatively charged surfaces, which can retain cations through electrostatic adsorption, as demonstrated by Shaheen et al. [56
] for vermiculites, smectites, imogolites and allophanes, while Chen et al. [68
] indicated that montmorillonites present high Cd removal capacity from aqueous solutions. Specifically, Serrano et al. [69
] indicate that soils with high pH and clay content have the highest Cd sorption capacities, taking into account that Cd retention would occur through precipitation reactions at high pH, while exchange reactions would dominate at low pH values.
In addition to that, Fe hydroxides play an important role in the retention of metals through a high specificity adsorption mechanism, often by direct coordination with surface oxygen, and trace elements adsorbed on these oxides could be exchanged only by other cationic metals having similar affinity for the surface, or by protons [70
]. Retention may also include isomorphic substitution and cation exchange mechanisms [56
Furthermore, total Ca content is very high in some of the materials here studied, especially in fine mussel shell, oak ash and hemp waste (see Supplementary Material
). In relation to this, Shaheen et al. [56
] indicated that the presence of free CaCO3
reduces the solubility of trace elements, which is attributed to a direct effect due to surface interactions, and to an indirect effect related to its repercussion on pH. Carbonated surfaces have a high affinity for Cd, and CdCO3
precipitates have been found on such surfaces [66
]. In fact, divalent metal cations have a tendency to associate with calcite, initially through surface adsorption reactions, and subsequently as precipitates within the calcite layers by recrystallization, giving a specific sorption with little tendency to desorption [72
]. On the other hand, the presence of carbonates in soils implies lower solubility of metallic elements, as a consequence of high pH values, which favors its precipitation (which can be considered within global sorption), although adsorption decreases at very high pH values [56
In the present study, when 6 mmol L−1 of Cd were added, and all the studied sorbent materials were considered, bivariate correlations analysis showed significant correlations between Cd sorption and pH (r = 0.933, p < 0.01), exchangeable Al (r = −0.781, p < 0.05), and exchangeable Ca (r = 0.754, p > 0.05). These results reflect the influence of pH on Cd sorption, since Cd sorption increases at higher pH due to the appearance of negative charges in variable charge compounds, and under these conditions there is less Al and more Ca in the exchange complex.
3.2. Pb Sorption on the Individual Materials
As shown in Figure 2
a, Pb sorption was clearly higher on the forest soil samples (maximum 51.4 mmol kg−1
, corresponding to 86% of 6 mmol L−1
added) than on the vineyard soil samples (maximum 36.6 mmol kg−1
, corresponding to 61% sorption), and pyritic material (maximum 35.7 mmol kg−1
, 60% sorption). When low Pb concentrations were added, the percentage sorption was between 97.7% and 99.9% for forest soil, between 77% and 99% for vineyard soil, and between 66% and 87% for pyritic material.
The higher Pb sorption on the forest soil compared to the pyritic material could be attributed to the higher organic matter content of the former. It could also explain the higher Cd sorption on the forest soil with respect to the vineyard soil, in conjunction with the Alo
contents in the forest soil sample (See Supplementary Material
), as previously pointed out by Fernández-Pazos et al. [22
] and Seco-Reigosa et al. [26
Regarding the effect of pH, Irani et al. [73
] indicated that Pb adsorption on different sorbent materials increased from pH 2 to 6, attributing it to a gradual increase in the negative charges on the sorbents, whereas at pH > 6 the formation of Pb hydroxides would lead to difficult adsorption. On the other hand, Petruzelli et al. [74
] demonstrated that Pb has great affinity for the formation of organic complexes, which remain stable till low pH values (pH 4). In acid media, usual Pb concentrations in the soil solution are between 0.003 and 0.046 mg L−1
, and under these conditions chemical speciation indicates that Pb is preferably in free form or as PbSO4
, whereas soluble organometallic complexes dominate when pH is neutral [75
]. Furthermore, according to McKenzie [76
], Fe oxides preferentially absorb Pb, in comparison with Cd.
As shown in Figure 2
b, Pb sorption was high on all by-products. Percentage sorption was always >99.8% for mussel shell and oak ash, >89% for hemp waste, and >86.9% for pine bark.
In previous works, Tofan et al. [77
] obtained 96% retention for Pb in hemp waste, with added concentrations similar to those in the present study. As for pine bark, Paradelo et al. [18
] demonstrated the high efficacy of this bio-sorbent to retain metals, especially Pb, in stable forms of low mobility, finding 100% retention when 2 mmol L−1
of Pb were added. The high tannin and lignin content of pine bark, as well as its functional groups (see Supplementary Material
), would be the cause of this great affinity for metals. High Pb adsorption capacity was also highlighted for oak ash by Papandreou et al. [78
], which found retention close to 100% at 48 h after adding 1 mmol L−1
shows that Pb sorption was better adjusted to the Freundlich model in most of the materials here studied. As in the case of Cd, mussel shell, hemp waste, and oak ash cannot be fitted to either model, which is frequent when sorption is as high as that of Pb in these materials. Reddy et al. [79
] also found a better fit for the Freundlich model using different biosorbents for Pb.
As for Cd, Pb sorption is highly pH-dependent, since this parameter affects the solubility of the metal ions, and also the ionization state of the functional groups in variable charge compounds of the sorbent surfaces [80
]. In the present work, the materials with higher pH showed clearly higher Pb sorption capacities, which is indicative of the influence of the acid-base conditions on metal retention. Lead adsorption processes on different types of biosorbents probably include various types of mechanisms, such as surface complexation, electrostatic attraction and ion exchange [80
]. According to these authors, when the exchange processes prevail, a decrease in pH in the equilibrium solutions is frequently observed, caused by an H+
increase in the solution after being exchanged with Pb2+
on the sorbent surfaces. In the present study, a decrease in pH at equilibrium with increasing Pb concentration added was also observed (except in oak ash and hemp waste); however, pH/sorbed-Pb correlations were only significant in the vineyard soil (r = −0.997, p
< 0.01) and pine bark (r = −0.995, p
< 0.01). This probably indicates the existence of ion exchange on these sorbent surfaces.
In the present study, DOC values undergo small variations in most materials when increasing Pb concentrations were added, with a DOC decrease observed in both soils, pine bark and hemp waste. DOC levels were significantly correlated with sorbed Pb just for forest soil (r = −0.897, p
< 0.05). A possible precipitation of organometallic complexes could explain DOC decreases. On the other hand, a DOC increase associated to Pb adsorption was pointed out by Karami et al. [81
] using biochar as sorbent, which would provide soluble organic compounds.
In the present study, when 6 mmol L−1 of Pb were added, and considering all the studied sorbent materials, Pb sorption showed a significant (p < 0.05) and negative correlation with exchangeable Al (r = −0.69) and with Al saturation (r = −0.757). There was also a significant (p < 0.01) and positive correlation with pH (r = 0.827). Pb sorption increases as pH grows due to the rise in negative charges, and both exchangeable Al and Al saturation decrease when pH increases.
3.4. Cd and Pb Sorption on the Amended Materials
Once the sorption and desorption of each material were analyzed separately, amendments were applied to both soils and the pyritic material, namely adding the three by-products that had given better results in the previous phases: mussel shell, hemp waste and oak ash, with individual doses of 48 t ha−1.
As shown in Figure 3
, the amendments allowed an overall increase in Cd sorbed on both soils and pyritic material, although this effect was not observed in the sorption curve corresponding to hemp waste applied on the forest soil (Figure 3
a). Oak ash was the most effective amendment in all three cases, and the smallest effect was that due to hemp waste.
The amended forest soil presented the highest sorption for all Cd concentrations added, in some cases close to 100% with the oak ash and mussel shell amendments, and always being >76% with both by-products. The oak ash amendment caused very similar effects on the vineyard soil. Cd sorption rates were always >54% with the oak ash and mussel shell amendments, and >42% with the hemp waste amendment. The oak ash amendment clearly increased Cd sorption on the pyritic material, reaching values of >97% when the lowest Cd concentration was added, and being >46% with the highest Cd concentration. The effects of the other two amendments on the pyritic material were clearly lower, with maximum Cd sorption of 75% (and minimum 40%) due to mussel shell, and between 64% and 42% when adding hemp waste.
The increased sorption after the addition of the amendments was consistent with that previously commented regarding the high Cd sorption capacity of oak ash and mussel shell, and followed the same order obtained individually for each of the materials used as amendment, that is, oak ash ≈ mussel shell > hemp waste. This took place even if the ranges of pH values in the equilibrium solutions corresponding to the amended materials were shorter than those previously found for the individual materials, all of them being <7, and, specifically: 4.24–6.25 for forest soil + mussel shell, 3.81–6.73 for forest soil + oak ash, 3.8–5.8 for forest soil + hemp waste; 3.55–4.02 for vineyard soil + mussel shell, 5.21–6.49 for vineyard soil + oak ash, 3.6–3.7 for vineyard soil + hemp waste; 3.74–4.08 for pyritic material + mussel shell, 4.05–6.51 for pyritic material + oak ash, 3.69–3.75 for pyritic material + hemp waste.
shows that the amendments increased Pb sorption as compared to the non-amended forest soil, vineyard soil and pyritic material. Oak ash was the amendment causing the highest increase in Pb sorption, always reaching >99% for any of the Pb concentrations added. Mussel shell caused a somehow lower, but similar, increase: Pb sorption >97% in forest soil and pyritic material, and >80% in vineyard soil. The hemp waste amendment gave Pb sorption >91% in the forest soil, >82% in the material pyritic, and >74% in the vineyard soil.
As in the case of Cd, the degree of enhancement in Pb sorption was consistent with that previously commented for the individual materials, and followed the same order obtained for each individual amendment, that is, oak ash = mussel shell > hemp waste. Again, the ranges of pH values in the equilibrium solutions corresponding to the amended materials were also shorter than those previously found for the individual materials, once again all of them being <7.
Vega et al. [84
] found that amending a mine soil with sludge and barley straw resulted in a great increase in Pb and Cd sorption capacity, mostly through Ca2+
displacement, although it also involved the displacement of other various exchangeable cations. Ramírez-Pérez et al. [20
] found high Cd sorption capacity for a mussel shell amended soil; in particular, the non-amended soil retained 15% Cd, rising to 87% when amended with mussel shell. Mussel shell amendment increased pH, and its high content in aragonite might be another important parameter in Cd retention, mainly due to (Cd,Ca)CO3
precipitation mechanisms [85
]. Furthermore, Shaheen and Rinklebe [86
] studied the effects of different emerging and low cost amendments on Cd and Pb retention in a contaminated floodplain soil, finding that most amendments decreased soluble + exchangeable Cd and Pb, whereas Fernández-Calviño et al. [87
] found high Cd and Pb sorption on mussel shell amended soils, attributed to the mineralogy of the mussel shell, since calcite and aragonite can effectively sorb these cationic metals.