A 4-month pot trial with Italian ryegrass (Lolium multiflorum Lam.) was carried out in a greenhouse. The soil was collected from the 0- to 15-cm soil layer of an agricultural field at UTAD (Universidade de Trás-os-Montes e Alto Douro), Vila Real, Portugal, air-dried, and sieved to 4 mm for the pot experiment and to 2 mm for analysis. The soil is non-calcareous (CaCO₃ < 0.5%), with a sandy-loam texture, a WHC of 32%, and very low OM, total nitrogen (TN), and extractable P (Olsen) concentrations (Table 1). Two composts (A and B) were prepared from the SPS and CGW, at proportions of 4:3 and 3:4 (by volume), respectively, by the static composting system (Rutgers) in a pilot plant (Santos et al., 2016; Table 1). The composts were applied to the soil at two doses and compared with a soluble N:P:K (10:10:10) commercial fertilizer (N as NH₄NO₃).

Six treatments were tested as follows: –

AD1 and BD1: compost A and compost B at dose 1 (D1), using 10 and 11 g pot⁻¹ of composts A and B, respectively — providing 360 mg TN pot⁻¹ (equivalent to 1.44 and 1.59 t ha⁻¹ of composts A and B, respectively or 52 kg N ha⁻¹), 350–272 mg P₂O₅ pot⁻¹ and 402–439 mg K₂O pot⁻¹, respectively;

A glasshouse experiment was run in 7-L double-layer polyethylene pots (System-Kick Brauckmann) which allow leachate collection from the bottom of the external part of the pot, through a polyethylene tube. The soil (9 kg of dry soil) was placed in each pot at a density similar to that in the field (1.3 g cm³). For the compost treatments, the composts were mixed with the soil at the beginning of the experiment, and for the F treatment, the mineral fertilizer was divided into two applications (1.8 g pot⁻¹ each time): before sowing and after the first harvest. All treatments were replicated four times (the total number of pots = 24). The moisture content of the soils was adjusted to 70% of the WHC with distilled water at the beginning of the experiment; then, 0.5 g of ryegrass seeds was added per pot (equivalent to 717 g m⁻²). The soil moisture content was corrected weekly by weighing each pot and adding water to reach 70% of the WHC. Any leachate produced after watering was collected from the bottom of the external part of the pot, quantified, and frozen for further analysis. The pots were placed in mobile racks which allowed frequent changes in position (every 4 days) inside the glasshouse (without temperature control). During the experiment, the plant material was harvested three times—30, 60, and 120 days after emergence—by cutting the aerial parts 2 cm above the soil surface. The fresh and dry weights were determined before and after drying at 60°C, respectively.

The potential nitrate leaching risk in the soil was assessed after the first and second harvests. For this, the moisture content of the soil was first adjusted to 70% of WHC; then, a heavy rain event (equivalent to 25 mm) was simulated by adding 1.75 L of distilled water to the top of the pot (equivalent to 60% of WHC). The leachates produced were quantified and frozen for further analysis.

At the end of the experiment, after the third harvest, the soil was separated from the main roots, homogenized, and sampled for analysis. The roots were cleaned—first by removing most of the soil particles with tap water, then two times with abundant distilled water, and, finally, by ultrasound (5 min at 750 W)—to eliminate any remaining soil particles adhering to the roots. Plant material (aerial part and roots) was dried at 60°C for 48 h and milled for analysis.

The moisture content of the plant material was determined by drying the sample at 60°C for 48 h in an oven with forced aeration. Then, chemical analyses were performed on the dry and ground samples: for compost and plants, total organic carbon (TOC) and TN by automatic microanalysis (EuroVector EuroEA 3000 elemental analyzer); after H₂O₂/HNO₃ microwave-assisted digestion, total P by a colorimetric method, K by atomic emission spectrophotometry (iCE 3000 series, Thermo Scientific), and Cu and Zn by atomic absorption spectroscopy (AAS) (iCE 3000 series, Thermo Scientific). The soil samples were analyzed for moisture content, by oven-drying the samples at 105°C for 12 h; WHC, gravimetrically, after allowing water-saturated soil to drain for 24 h; total CaCO₃ by the calcimeter method; pH, in a saturated soil paste; electrical conductivity (EC), in a 1:5 water extract; TOC and TN, by automatic microanalysis (Navarro et al., 1993); OM concentration, calculated by multiplying TOC by 1.72 (Nelson and Sommers, 1982); NH4+–N, by a colorimetric method based on Berthelot’s reaction (Sommer et al., 1992), after 2 M KC1 extraction (1:20, w/v); NO3−–N, with a selective electrode (EPA 9210A) in a 1:20 water extract; available P, by a colorimetric method, after 0.5 M NaHCO₃ extraction (1:10, w/v) (Olsen-P; Watanabe and Olsen, 1965); available-K, by EAS, after 1 M CH₃COONH₄ extraction (1:10, w/v) (Kundsen et al., 1982); water-soluble and exchangeable Cu and Zn, by AAS, after 0.1 M CaCl₂ extraction (1:10, w/v) (McGrath and Cegarra, 1992). The leachates were analyzed for water-soluble nitrogen (N_w), using an automatic analyzer for liquid samples (TOC-V CSN Analyzer, Shimadzu), and nitrate, by ionic chromatography (850 Professional IC, Metrohm, Herisau, Switzerland). All the analyses were performed at least in duplicate, with analytical errors of <2%.

The plant N-utilization efficiency was calculated on the basis of the ARF, as suggested by Gunnarsson et al. (2010), and was reported as cumulative according to the following equation: ARF = ∑t = 13(N uptake treatment)− (N uptake control)TNadded×100

in which N uptake treatment is the total N uptake (mg pot⁻¹) of a fertilizer treatment (AD1, AD2, BD1, BD2, and F) at time n (n = harvest 1, 2, and 3); N uptake control is the total N uptake (mg pot⁻¹) of the unfertilized control at time n (n = harvest 1, 2, and 3); TN_added is the total N added to each pot (mg pot⁻¹).

The RAE of N, P, and K in the compost treatments was calculated from the nutrient uptake in the aboveground biomass of the compost treatment in comparison with the mineral fertilizer, taking into account the input in the double-dose treatments (Brod et al., 2015): RAE=100×(Uptake treatment/Applied treatment) /(Uptake mineral fert./Applied mineral fert.).

The NUE of the compost and of the mineral fertilizer was calculated by expressing as a percentage the ratio of the N_output (N uptake by plants) and N_input (N applied in the treatment) of the soil (EU Nitrogen Expert Panel, 2015). NUE=100×Noutput/Ninput

For statistical analysis, IBM SPSS Statistics 21.Ink software was used. The standard error of the means was calculated for the chemical and physicochemical characteristics of the soils and the plant biomass and nutrients; a one-way analysis of variance (ANOVA) was performed, and differences between means were determined using Tukey’s test. A two-way ANOVA was used to determine significance in the plant parameters according to the harvest and treatment and to the type of compost and application rate. The normality and homogeneity of the variances were checked using the Shapiro–Wilk and Levene tests, respectively, before ANOVA.

The shoot biomass production of L. multiflorum (dry weight) increased in the compost treatments and with the compost application rate, in comparison with the control plants, at all harvests, without significant differences between the two composts (Figure 1). The root biomass production was highest in F and only treatment AD2 gave results statistically different from the control (Figure 1). The compost treatments at D1 gave shoot biomass values similar to that of F in the first harvest, with significantly greater values at D2, due to the greater supply of inorganic N (300 and 260 mg pot⁻¹ in composts A and B, respectively) in comparison with D1 (150 and 130 mg pot⁻¹ in composts A and B, respectively) and with the first application of F (180 mg pot⁻¹). The pig manure composts had high inorganic N concentrations (Table 1), supplying adequate N for correct plant growth after soil application. Thus, this readily available N seems to be the main factor responsible for the augmented plant growth (Eghball and Power, 1999), demonstrated by the positive correlation (r = 0.98; p < 0.001) between the aerial biomass production (DM) and the TN supplied by the amendments. The dependence of DM yields of Italian ryegrass on the N supply was suggested by Douglas et al. (2003), who used different organic wastes in a field experiment. Treatment F led to the greatest plant DM production at harvests 2 and 3, as the mineral fertilizer was divided into two applications, at the beginning of the experiment and after the first harvest, providing available N for the subsequent plant growth.

Inorganic N (NH4+ and NO3−) in the composts, readily available to plants, was responsible for the effects seen at the first harvest, but organic N needed to be mineralized for plant uptake prior to the following harvests. The decreased yields at harvests 2 and 3 for the compost treatments may indicate scarce mineralization of the organic N in both composts, as observed in previous studies (Bernal et al., 1998; Moral et al., 2009), due to the microbial stability of the OM, with a low amount of degradable TOC (Bernal et al., 1998, 2009a). Indeed, the N mineralization of mature compost is usually slow and may not have been sufficient to provide enough inorganic N to the crop for the following harvests. Cordovil et al. (2006) reported a good relationship between the N supplied to ryegrass by different organic amendments and their potentially mineralizable N. Different studies revealed differences in the mineralizable N in compost; for instance, Moral et al. (2009) reported N-mineralization values ranging from 0 to 34% of TN in composts prepared from manures, but only from 1 to 4.7% of TN in those prepared from pig slurry.

Also, fresh organic wastes and immature or unstable compost can cause N immobilization in the soil after their application, mainly for high C/N ratios, reducing plant-available N and leading to low yields (Bernal et al., 1998; Douglas et al., 2003; García-Gómez et al., 2003). Increases in the biomass production of maize (12–15%) with the use of composted pig manure, in comparison with the fresh manure, were found by Loecke et al. (2004) during a 2-year experiment. Cordovil et al. (2007) also found a greater wheat production with composted pig manure than with pig slurry, and Bedada et al. (2014) observed an increase in crop production of 11%, in comparison with the control treatment, when compost prepared with cattle manure and agricultural wastes was applied. Thus, compost maturity guarantees the N mineralization after soil application, avoiding soil microbial N immobilization.

Compost application significantly increased the concentrations of the macronutrients N and K in the plants, compared with the control, at each harvest, without significant differences relative to treatment F (Table 2). Only treatment BD1, at the first harvest, reduced the N concentration in the plants with respect to F, due to the low inorganic-N concentration in compost B (Table 1). All compost treatments led to greater P concentrations in the plants than did treatment F at harvests 2 and 3, even at D1. This was probably due to a concentration effect associated with the low yields obtained at those harvests for the compost treatments (Ylivainio et al., 2008). So, both composts were able to substitute only partially for mineral fertilizer.

Besides the significant, positive responses of the N, P, and K concentrations in ryegrass to compost application, a significant decline in N with the harvesting cycle was observed, which may be due to the consumption of the easily available N and scarce N mineralization discussed previously, or to nutrient leaching after each “rain event” (discussed below). Despite the slightly higher N concentration in plants from treatment D2, relative to D1, in the first harvest, the values for the compost-based treatments did not differ significantly (P < 0.001), and only TN in BD2 was higher than the control values. Considering the optimal shoot N, P, and K concentrations in Italian ryegrass (30–42 g N kg⁻¹ DM, 3.5–5 g P kg⁻¹ DM, and 25–35 g K kg⁻¹ DM; Bergmann, 1986), N and P might have limited the growth, particularly at harvests 2 and 3. The P concentration in the aboveground biomass was within the lower range found by Ylivainio et al. (2008), who suggested that a low P concentration can be due to a dilution effect caused by the enhanced growth of L. multiflorum—as shown at harvest 2, when the P concentration was lowest for plants of treatment F, which also gave the highest yield. The N:P ratio in the aerial parts of the plants can indicate the nutrient that is restricting plant growth (Güsewell, 2004). The N:P ratio of >20 found at the first harvest, for all treatments, suggests that plant growth was initially restricted by P availability, while, at harvest 3, the N:P ratio of <10 indicates that N limited plant growth at the end of the experiment for the compost treatments (Güsewell, 2004). At the second harvest, for all treatments, the N:P ratio was adequate. These results indicate that the plant-available N from compost and F was adequate immediately after soil addition, but the precipitation or fixation of P can occur shortly after addition to the soil (e.g., due to insoluble Ca-phosphate formation), being released later on during the growing session. Organic acids in the soil, produced from root exudates, may compete for the same binding sites as P, depressing the P sorption strength of the soil (Ylivainio et al., 2008). The exudation of organic anions by the roots has been found to occur in the rhizosphere, associated with NO3−–N nutrition in preference to NH4+–N (Marschner, 1995; Hinsinger et al., 2003); also, localized release of H⁺ by some portions of the roots has been reported to occur as a response to P deficiency, favoring P solubility in the soil (Hinsinger et al., 2003). Jones (1998) reported two mechanisms of P release in the soil rhizosphere: direct ligand exchange, whereby citrate directly replaces P on ligand exchange surfaces, and complexation with metal ions in the solid phase, which constitutes the exchange matrix holding P (such as Ca²⁺ in rock phosphate). According to Jorgensen et al. (2010), composting of the separated solids from animal slurry does not alter the P availability but P precipitation as Ca or Mg phosphate may occur during composting due to the pH increase. Ylivainio et al. (2008) indicated that the high proportion of inorganic P in manure compost makes it a convenient P fertilizer for L. multiflorum due to its high root-to-shoot ratio of P (as shown in the present experiment; Tables 2 and 3); these authors referred to it as a P-efficient species with relatively low external P requirements.

The nutrient uptake by the aerial parts of L. multiflorum during the pot trial is shown in Figure 2. Compost application, particularly at the high rate (D2), significantly increased the N, P, and K accumulation by the shoots, when compared with the control (Figure 2). The uptake of N, P, and K was greatest in treatment F (324, 13.5, and 467 mg pot⁻¹, respectively), with a considerable difference from BD2 (270, 11.4, and 362 mg pot⁻¹, respectively). Composts A and B did not show any significant differences at the same application rate, but nutrient uptake increased with the compost application rate. An increase in N uptake with the compost application rate was also found by Cordovil et al. (2006), but differences were not found when organic and inorganic fertilizations were combined, suggesting that the mineral N supplied could have stimulated not only N uptake but also that of other nutrients. Also, Ylivainio et al. (2008) reported a higher P uptake by L. multiflorum when composted manure was applied at a double rate and even greater uptake at a quadruple dose, at all harvests.

High concentrations of Cu and Zn are usually found in pig slurry compost (Sáez et al., 2017). The Spanish legislation for fertilizing products (Ministerio de la Presidencia, 2013) classifies composts according to the total Cu and Zn concentrations. The composts in the present experiment belonged to class C, with Cu between 300 and 400 mg kg⁻¹ and Zn between 500 and 1000 mg kg⁻¹, limiting the use to 5 t ha⁻¹ year⁻¹ on agricultural soil. The concentrations of Zn and Cu in the plants were analyzed in order to determine their accumulation due to the use of pig slurry composts. At the first harvest, the shoot Cu and Zn concentrations were generally higher in the compost treatments than in control plants, but without statistically significant differences from treatment F. The concentrations of Zn measured in the aerial parts of the plants can be considered normal for plant nutrition (27–150 mg kg⁻¹; Kabata-Pendias, 2001) and are within the range found for grasses (12–80 mg kg⁻¹; Kabata-Pendias, 2001). Although the shoot Cu concentration remained higher in the compost treatments than in the control, the values are within the normal range for plants (5–30 mg kg⁻¹; Kabata-Pendias, 2001) and the range for grasses (1.3–33.1 mg kg⁻¹; Kabata-Pendias, 2001). The Cu concentration was decreased at the following harvests; this may be due to the ability of Cu to form chelates with the stable and humified OM (Bernal et al., 2009b). Therefore, both composts used here, A and B, did not present any risk regarding metal toxicity to plants.

The nutrient fertilizing efficiency of the compost was estimated as the ARF, calculated according to Eq. 1. With respect to the control, the N-ARF of the composts ranged from 24 to 29%, without significant differences between the compost types or doses (Table 4). Similar values (28%) were found by Thomsen and Kjellerup (1997) for manure without litter (cattle feces), being 8–14% for manure with litter (cattle feces + urine + straw), related to the inorganic-N concentration; lower values were found in winter, associated with N leaching. The mineral fertilizer treatment (F) reached the greatest N-ARF value, because all the N was in inorganic forms (NH₄NO₃), whereas in composts A and B, 26.7 and 19.0%, respectively, of the TN was in inorganic forms (Table 1). This indicates that the inorganic N present in compost A was assimilated by the plants, but partial mineralization of the organic N of compost B may have occurred, rendering it available to plants. Cordovil et al. (2006) compared the potentially mineralizable N of several organic materials (determined in incubation experiments) with the plant N uptake in a pot experiment. Their results showed that the plant N uptake from pig manure compost exceeded the potentially mineralizable N (33.3 and 25%, respectively), similar to the results found here, while compost from municipal solid waste led to very similar values of N uptake and potentially mineralizable N (19%). Also, Douglas et al. (2003) found a very low N-ARF of green compost, in spite of its high TN concentration (4.3–5.8 g kg⁻¹).

Compared with the efficiency of the mineral fertilizer, the values of RAE (Table 4) confirm that the mineralization of organic N in the mature composts occurred in the soil during the crop development, accounting for 10–20% of the TN in compost A and an average of 22% in compost B. According to Bernal et al. (1998), the potentially mineralizable N increases with composting time, representing about 9% of the TN in mature compost derived from sewage sludge. Although the NUE is considered applicable to agriculture and food production–consumption systems, this parameter also can be applied to small agronomic experiments (EU Nitrogen Expert Panel, 2015). Considering the inputs and outputs in the present system (pots), the maximum NUE occurred with F, its value decreasing with the compost application rate (Table 4). The NUE values may also indicate that N mineralization occurred in the system, representing 8–18% of TN, with greater values for compost B.

The ARF of P was low in all treatments, the highest value being in treatment F (6.6%; Table 4). This implies that this nutrient was applied in excess of crop requirements and accumulated in the soil. In the work of Ylivainio et al. (2008), the ARF of P for ryegrass in the first growing year was 20–26% for manure compost applied at different rates, with similar values in the treatments with mineral fertilizer (21.7–26.8%). Grigatti et al. (2017) found, in a 21-week pot experiment, P-ARF values for green waste compost of 10–40%, depending on the characteristics (P fractionation) of the compost. They used a P-application rate (15 mg kg⁻¹) similar to that in the present experiment; the ARF was related to the H₂O + NaHCO₃−extractable P, which was mainly inorganic P. The solubility of the P from compost may depend on the degree of maturation. During composting of animal manures, Eneji et al. (2003) found that water-soluble P decreased during composting, shifting toward the more resistant HCl-soluble P (associated with Ca and Mg), which was the dominant phase at the end of composting. The high degree of maturation of the compost used in the present experiment (Santos et al., 2016) may have reduced the availability of P. The RAE of P in compost B (74.6 and 52.7%, respectively, for doses 1 and 2) was greater than in compost A (56.1 and 40.0%, respectively, in doses 1 and 2), due to the lower concentration of P in compost B, giving a greater efficiency of P uptake by the plants. Brod et al. (2015) determined the P-RAE of different organic waste materials for ryegrass: the greatest values occurred for dairy cattle manure, with a greater efficiency than the mineral fertilizer (Ca(H₂PO₄)₂), followed by liquid digestate and chicken manure (72–78%), with lower values for fish sludge and solid digestate (47–70%), for one or up to four harvests in a soil of pH 6.9. These values were well predicted using the NaHCO₃−extractable P (Grigatti et al., 2017).

A very different behavior was shown by K, with 100% of the K added to the soil in F assimilated by the plants while the ARF reached very high values for composts A and B, without differences according to the compost type and dose (Table 4). This is due to the highly available (mainly soluble) forms of K present in the composts; the fraction not taken up by plants can be retained in the soil as exchangeable and soluble forms, but can also be lost by leaching. The high EC of the leachates of the compost treatments (up to 630 μS cm⁻¹ in the first rain event and up to 263 μS cm⁻¹ in the second), with respect to the control (116 and 127 μS cm⁻¹, rain event 1 and 2, respectively) and treatment F (219 and 170 μS cm⁻¹, rain event 1 and 2, respectively), may indicate the leaching of soluble salts, including K.

Losses of N during the experiment could have occurred by leaching or denitrification (Hoffman and Van Cleemput, 2004; Burgos et al., 2006; Demuras et al., 2016). For leaching to occur, N must be in a water-soluble, mobile form, and the rainfall must be abundant enough to transport N through the soil. Nitrate is the N form most susceptible to leaching, since ammonium N can be retained in the soil cation exchange capacity. In this sense, water-soluble TN (N_w) present in the leachate from the first rain event occurred as nitrate (data not shown); then, only N_w was measured in the following leachates.

Considering the threshold of 50 mg l⁻¹ for NO3− concentration in ground and surface water set by the European Legislation (European Union, 1998), all the leachate samples were far away from that value (maximum NO3− concentration 5.28 mg l⁻¹). All the fertilizer treatments had higher N losses than the control during the experiment (1.7- to 2.1-times greater), without significant differences between treatments or rain events (Figure 3). The amount of N leached during the rain events did not differ statistically among the fertilizer treatments, and similar values also occurred during the irrigation in the soils treated with compost B (Figure 3). For TN, the total proportion lost (during irrigation and rain events) was lowest when each compost was applied at a high rate (Figure 3). The WHC of the composts may have had a strong influence, reducing N leaching (Fulekar, 2010), as compost can improve soil properties to achieve more efficient water utilization. In fact, compost B, prepared with a greater amount of CGW than compost A, produced a slightly lower volume of leachate during the rain events than compost A, especially at D2 (2,100 ± 395 and 1,695 ± 380 ml pot⁻¹ in AD2 and BD2, respectively; 2,180 ± 212 ml pot⁻¹ in F). In terms of the doses applied, no significant differences were found when comparing the amounts of N leached (Figure 3). A higher NO3−–N application might imply a greater N leaching risk, due to the excess of NO3− with respect to the plant requirements. However, the lack of a significant change in the loss of N by leaching (or NO3− accumulation in the soil, Table 5) indicates that the NO3−–N provided by the compost was used by the plants (as the ARF, RAE, and NUE also indicate). Burgos et al. (2006) pointed out that N losses by leaching tend to be higher when poorly stabilized organic materials rich in N are applied, but also stated that NO3− losses by leaching from compost-amended soils are usually very low. Demuras et al. (2016) concluded from a field experiment that the replacement of organic with mineral N-fertilization systems was not effective at mitigating NO3− leaching, although the residual N remained in the soil for further mineralization. In fact, they did not find any reduction in N leaching during winter, when mineral fertilizer was used instead of organic N.

The chemical properties of the soils at the end of the experiment are shown in Table 5. The TOC in the soil showed small and non-significant differences between the treatments with compost and the control. Evanylo et al. (2008) found small differences in the TOC in soils amended with compost at low doses, indicating that a relevant proportion of the OM from the compost may have been mineralized during the experiment. All the treatments were able to increase TN in the soil, at the end of the experiment, with respect to the control soil, without differences among the fertilizing treatments. Increases in TN in the soil after application of composts produced from livestock manure or from municipal solid wastes have been found in pot studies (Soumaré et al., 2003; Evanylo et al., 2008; Karami et al., 2011)—indicating that the resistant fraction of the organic N in mature composts accumulated in the soil, being incorporated into the humified OM.

The lack of significant differences in the soil NH4+–N concentration among the treatments indicates its complete nitrification, in similar proportions in all treatments. The lowest concentrations of NO3−–N were observed in the soils amended with compost A, in spite of its high NO3−–N concentration (Table 1), but the RAE and NUE results indicate lower N mineralization than in compost B. No statistically significant differences occurred for the rest of the treatments, suggesting that the inorganic N supplied by the amendments may have been used by the plants or lost from the soil. The low N losses by leaching during the two rain events (Figure 3) indicate that the N applied in excess accumulated in the soil, as losses by denitrification may have been irrelevant at a soil moisture content of <70% WHC. According to Estavillo et al. (1996), N losses by denitrification from mineral fertilizers can be 9–17% during spring/autumn, reaching 2% from mature composts (Cambardella et al., 2003). In fact, the TN accumulation in the soil with the compost and mineral fertilizer treatments occurred as organic N (Table 5).

The concentrations of available P and K were lowest for treatment AD1 and greatest in treatment F (Table 5), which may be due to the excess P supplied by the treatments, giving a low ARF, as discussed previously, and the highly soluble nature of the K supplied by all treatments. However, a greater proportion of the excess P provided by the mineral fertilizer may have remained in the soil in the available fraction, in comparison with that provided by the compost, because P in well-matured compost can be found in less available forms (Eneji et al., 2003; Grigatti et al., 2017).

The elements Cu and Zn are essential plant micronutrients, but their accumulation in the soil can cause toxicity to plants (Kabata-Pendias, 2001). The use of compost from pig manure can lead to soil accumulation, this being one of the main environmental risks associated with compost use. In fact, limits for the Cu and Zn concentrations in compost for agricultural use have been set in several countries and in the EU for soil protection (Bernal et al., 2017). In the present experiment, the composts did not significantly increase the available Cu and Zn concentrations in the soil, with respect to the control (Table 5), and similar values were obtained for Zn in treatment F. These results, together with the plant concentrations found, indicate that these elements in the compost had a beneficial effect, as plant micronutrients. Considering the total concentrations of Cu and Zn in the compost and the application rate, the increases in their total levels in the soil were low (Cu < 0.4 and Zn < 1.0 mg kg⁻¹ soil), with a scarce risk of soil accumulation at the compost rates applied. However, their pollution risk due to accumulation resulting from repeated applications over a long period needs to be tested in field experiments.