Soil contamination with metals ¹ constitutes a significant global environmental challenge (González Henao and Ghneim-Herrera, 2021; Karnwal et al., 2024). The ecotoxicity of metals in soils can be assessed by a two-fold approach (ISO 17402, 2008): (1) quantifying the total metal concentration and/or an operationally defined fraction of the total metal concentration in the soil, and (2) exposing organisms to the soil and quantifying biological responses. If the obtained metal concentrations and biological response(s) are negatively correlated, threshold values for soil metal ecotoxicity may be derived.

Many ecotoxicological studies have been performed on uncontaminated soils that have been spiked artificially with soluble metal salts (e.g., Halim et al., 2021; McKee et al., 2017; Tibihenda et al., 2022; You et al., 2024). Spiked soils consistently result in increased toxicity for plants and soil organisms relative to the toxicities observed in real-world soils that have been contaminated for decades (Neaman et al., 2020; Oorts et al., 2006; Smolders et al., 2003). The prolonged residence time and associated partitioning of metals in soils, known as “aging” (Martínez and McBride, 2001), significantly impacts metal bioavailability and toxicity (Islam, 2025) in such a way that cannot be replicated in artificially contaminated soils. Rather, metal spiking tends to selectively eradicate specific microbial species over periods of months to years, leading to an artificial loss of metabolic functioning and biogeochemical processing associated with these species (Giller et al., 1999).

High metal content in soils can disrupt soil ecosystems, prompting researchers to identify the most sensitive species for detecting early signs of toxicity (Broos et al., 2005). However, there is a lack of studies on real-world contaminated soils that compare the sensitivity of different types of organisms as bioindicators of metal toxicity. The study of Naveed et al. (2014) on soil contaminated by wood treatment with copper-based fungicides demonstrated that earthworms exhibit greater sensitivity to copper than do bacteria, nematodes, and fungi. Likewise, the study of Dovletyarova et al. (2024) demonstrated that plants exhibit greater sensitivity to metals than microorganisms do.

The results of the two latter studies can be explained by the fact that the available ecotoxicological data on metal toxicity for plants and invertebrates typically correspond to the individual level of biological organization, whereas microorganism responses are usually assessed at the community level (Santa-Cruz et al., 2021). Resistance to metal-induced stress indeed varies with the level of biological organization (Spurgeon et al., 2005). Typically, populations and communities are less sensitive to stress than are individuals. Accordingly, individual-level responses of plants and invertebrates are expected to be more sensitive to metal toxicity than are the community-level responses of microorganisms.

Moreover, multiple studies have demonstrated the existence of microbial tolerance and resilience to metals, i.e., development of resistant microbial populations and adaptive mechanisms that enable survival in metal-contaminated soils (reviews of Das et al., 2022; Mallick et al., 2015; Naik et al., 2018; Narayani and Shetty, 2013). This well-established adaptive response of soil microorganisms to metal stress may compromise their suitability for evaluating metal toxicity. However, does this inference imply that microbial indicators lack sensitivity across all contamination scenarios, or are they selectively applicable to specific metals or environmental conditions?

Further investigation is warranted to establish the robustness of microbiological responses as metal toxicity indicators across a range of real-world contamination scenarios. Thus, the purpose of this review is to evaluate the validity of microbiological responses as indicators of metal ecotoxicity in real-world contaminated soils.

Our first criterion for the selection of studies was to include only real-world contaminated soils, excluding artificially contaminated soils. However, considering that sewage sludge application to agricultural fields is a widespread practice worldwide, we also included studies on soils on which sewage sludge was applied more than 7 years prior to the study. For simplicity, these sewage sludge studies will also be referred to as “real-world contaminated soils” since the period of 7 years appears to be sufficient for metal aging in soils, based on the data reported in several long-term aging studies (Ma et al., 2006; McBride and Cai, 2016; Zeng et al., 2017).

Total metal concentration in contaminated soils may not be a consistent factor in predicting potential toxicity for plants and soil biota (ISO 17402, 2008). As a result, a wide-ranging set of approaches has been developed to estimate the “bioavailable” metal fraction in soil by linking operationally defined metal pools to specific biological responses (Kim et al., 2015). The operationally defined bioavailable metal pools are targeted by extractions with water, salt solutions, chelating agents (e.g., DTPA or EDTA), or acids (e.g., HCl or CH₃COOH) (Supplementary Table 1). In the following discussion, we will refer to these bioavailable fractions as “soluble” when the researchers used salt solutions or water, or as “extractable” when chelating agents or acids were used.

We reviewed studies that reported the values of effective concentrations at 10% and 50% (EC₁₀ and EC₅₀) of soil metals (either total concentrations or bioavailable concentrations) for soil microbiological response in real-world contaminated soils (Supplementary Tables 2–9). We were able to identify studies reporting toxicity thresholds using microbiological responses in real-world contaminated soils for the following elements: arsenic, cadmium, copper, lead, nickel, and zinc, whereas data for other metals were absent. Thus, we decided to focus our review on these six elements.

A single effective concentration value for a specific biological endpoint (e.g., growth, reproduction, survival, etc.) is insufficient for decision making in agricultural or ecological contexts (ISO 17616, 2019). As an alternative, Checkai et al. (2014) recommend averaging effective concentration values across different biological endpoints. Adopting this approach, we summarized available metal toxicity thresholds for soil microorganisms (Table 1). Few studies have established toxicity thresholds of soluble or extractable soil metal fractions for microorganisms using real-world contaminated soils. We were able to identify studies only for copper (Supplementary Table 8) and zinc (Supplementary Table 9), whereas data for other metals were absent. Supplementary Tables 10–15 summarize studies reporting correlations between various metal pools in soil and microbiological responses.

Our second criterion for the selection of studies was to include studies in which metal concentrations (either total concentrations or bioavailable concentrations) were above the effective concentration at 50% (EC₅₀) for microorganisms (Table 1) for at least one metal. Supplementary Tables 16–19 present metal concentrations reported in each study, highlighting metal concentrations above the EC₅₀ values for microorganisms. Therefore, in the studies quantified in Table 2, metal concentrations exceeded mean reported toxicity thresholds and metal toxicity was expected for soil microorganisms. The specific authors corresponding to each study counted in the table are listed in Supplementary Tables 20–24.

On the other hand, we noted studies reporting statistically significant negative correlations between microbiological responses and soil total copper concentrations, which were below the reported effective concentration at 10% (EC₁₀) for copper using microorganisms’ responses as a bioindicator (Table 1). In other words, these studies report copper toxicity to soil microorganisms at low copper concentrations, at which toxicity is not expected. These studies are summarized in Supplementary Table 25. Conversely, we also found studies reporting statistically significant positive correlations between microbial responses with rising soil copper concentrations, even though none of the metal concentrations exceeded the mean reported EC₁₀ threshold for microbial bioindicators. These findings are presented in Supplementary Table 26.

We arranged the reviewed information according to the well-established classification of microbiological responses used in soil quality assessment (Benedetti and Dilly, 2006): (1) soil microbial biomass and number, (2) soil microbial activity, (3) soil microbial diversity and community structure, and (4) plant-microbe interactions (Table 2). Additionally, Shaw and Burns (2006) posited soil enzymatic activity as a proxy of soil microbial activity. To this end, we also included studies on soil enzymatic activity in our review (Table 2). In the following discussion, the terms “microbiological responses” or “microorganism responses” refer to at least one of the five types of responses listed above.

This review demonstrates that the effects of soil metals on microbiological responses range from toxic (negative) in soils with metal concentrations below the mean reported EC₁₀ values to stimulatory (positive) in soils with metal concentrations above the mean reported EC₅₀ values. Hence, in some cases, microorganisms’ responses indicate metal toxicity at low soil metal concentrations, at which toxicity is not expected. In contrast, in other cases, microorganisms are stimulated by metals at high soil metal concentrations, at which stimulatory responses are not expected. Further, soil microbiological responses can be influenced by soil physicochemical properties rather than soil metals concentrations even at metal concentrations above the mean reported EC₅₀ values, at which metal toxicity for soil microorganisms is expected. In summary, the unpredictable nature of microbiological responses to metals makes them unreliable indicators of metal toxicity in real-world contaminated soils.

As discussed above, individual-level responses of plants and invertebrates are known to be more sensitive to metal toxicity than are the community-level responses of microorganisms. Indeed, the toxicity responses of plants and earthworms are largely predictable based on reported EC₁₀ or EC₅₀ values (Neaman et al., 2025; Schoffer et al., 2024; Tapia-Pizarro et al., 2025). Thus, we propose using plants and invertebrates, rather than microorganisms, as bioindicators of metal toxicity in real-world contaminated soils.

The following hypotheses can be tested in future research:

Metal uptake and retention by microbial cells in soils is regulated by the capacity of the total soil metal pool (dominated by the solid phase) to supply metal ions to the soil solution at the time when microbial cells are assimilating metal ions.

To test this hypothesis, the methodology of the study of Khan et al. (2009) can be used. Specifically, a salt solution can be used to extract the soluble metal pool from soils following chloroform fumigation to estimate the soil metal fraction retained by the cells of microorganisms. This approach will help discern the relationship between cellular metal concentrations, the concentrations of metals in the soil solution, and the total metal concentrations in soils.

2. Fungi have greater tolerance of metals than bacteria do.

To test this hypothesis, the methodology of the study of Naveed et al. (2014) can be used. Specifically, DNA analysis can be used to quantify bacteria and fungi richness and diversity in soils with an increasing gradient of metal contents. Then, effective concentrations at 10% and 50% (EC₁₀ and EC₅₀) can be derived for fungal and bacterial responses, allowing the comparison of the sensitivity of fungi and bacteria to metal stress.