Most ecotoxicological research of the effect of metals on soil organisms focuses on the toxicokinetics of these elements to predict their toxicity to soil invertebrates or microorganisms. Typically, it is thought that free ions cross the cell membrane, are distributed to the site of toxic action, and cause deleterious effects (Peijnenburg and Jager 2003). In this paradigm, metal toxicity is dependent on the availability of metals present in soil porewater. In turn, metal porewater availability is linked to the partitioning of metals between soil and water, the availability of porewater itself (soil moisture), as well as organism traits that influence organism exposure to it. Despite considerable research on this topic, there is still no clear and consistent correlation between metal availability, soil properties, and toxicity. For instance, soil pH is a good predictor of metal availability across soils but not of toxicity for many metals (Smolders et al. 2009). The mismatch between availability and toxicity may arise if porewater ingestion or dermal exposure is not the dominant exposure pathway. It is possible that for invertebrates soil ingestion and subsequent changes of metal availability in the gut explain why metal bioavailability is not always linked to toxicity. Annelids, including both earthworms and enchytraeids, actively ingest soil as a result of their burrowing behavior; but for other invertebrates (oribatid mites and collembolans) it is unclear if or how much soil is ingested. For earthworms only one experiment has tried to distinguish dermal from oral exposure to metals, but more explorations under different conditions/metals/soils/species are needed (Vijver et al. 2003). In this Points of Reference, we argue that the paradigm where toxicokinetics is the sole driver of metal toxicity to soil organisms should be abandoned and that we should look also at toxicodynamic drivers.

Toxicodynamics is the dynamic interaction of a toxicant with a site of toxic action and its subsequent biological effects. Typically for metals, toxicodynamic research focuses solely on the site of toxic action, such as reactive oxygen species generation or calcium homeostasis disruption. In the soil environment there are multiple factors that influence soil organism health, such as texture, organic matter composition, pH, and cation exchange capacity. In the past, these parameters have largely been viewed through the lens of toxicokinetics. For example, studies on cation exchange capacity or organic matter focus on how these factors influence metal bioavailability. We argue that we need a different view, in which we explicitly recognize that soil factors not only are key for toxicokinetics but in fact drive toxicodynamic behavior in the organism (Figure 1). Energy is one example by which soil properties drive toxicodynamics of metals within an organism. For example, Oppia nitens has an increased tolerance to metals in high–habitat quality soils compared to those with low habitat quality (Jegede et al. 2019), despite similar Zn bioavailability. An organism's ecological strategies can also affect its responses to metal contamination; for example, some species could reduce reproductive output as a strategy to increase energy allocation for metal resistance and survival (Van Gestel and Hoogerwerf 2001). Although this hypothesis has been suggested (Van Gestel and Hoogerwerf 2001), it has not been experimentally demonstrated, and the mechanisms behind these strategies are not well known. Finally, climate, which has been a major focus in recent years as a result of climate change, is known to affect the response of organisms to contamination through temperature and is expected to function as an added stressor to the organism's biology. Although experimentation has been performed at different temperatures, few studies, if any, report the toxicodynamic mechanisms of how temperature affects organism response to metals.

The current risk assessment of metals in soil is broken. Typically, field observations and laboratory toxicity tests show no correspondence. Further, in the real world, metals exist as mixtures; and our existing methods to account for this are theoretically inadequate and empirically not predictive. Yet, field practitioners recognize that increasing the quality of the environment, via amendments or eco‐restorative practices, can dramatically increase the abundance of organisms at impacted sites. Our current risk‐assessment framework is unable to account for this because it can only account for toxicokinetic changes in metals.

A new risk‐assessment framework needs a different platform of ecotoxicological research, one that focuses on the key toxicodynamic questions that determine organism response to pollutants. We can build off of the existing literature for how metals interact at the site of toxic action, but additional research is needed on how habitat quality, organism behavior, and organism traits interact during metal impacts. Once built, we can readily adapt existing risk‐assessment frameworks to modify predicted risk based on toxicodynamic modifiers. In doing so, we can combine the best of toxicokinetic research with toxicodynamic research and thereby better predict the real risk in our environment.

关于金属对土壤生物的影响的大多数生态毒理学研究都集中在这些元素的毒代动力学上，以预测其对土壤无脊椎动物或微生物的毒性。通常，人们认为自由离子穿过细胞膜，分布到毒性作用部位，并造成有害影响（Peijnenburg和Jager 2003）。在这种范式中，金属毒性取决于土壤孔隙水中存在的金属的有效性。反过来，金属孔隙水的可利用性与土壤和水之间金属的分配，孔隙水本身（土壤水分）的可利用性以及影响生物体对其暴露的生物特征有关。尽管对此主题进行了大量研究，但金属可用性，土壤性质和毒性之间仍然没有明确且一致的相关性。例如，土壤pH值是跨越土壤金属可用性的良好预测但不是毒性许多金属（的闷烧等人 2009年）。如果毛孔摄入或皮肤暴露不是主要的暴露途径，则可用性和毒性之间可能会发生不匹配。对于无脊椎动物来说，消化道中肠道金属利用率的随后变化可能解释了为什么金属生物利用率并不总是与毒性相关。lid，包括earth和肠线虫，由于其穴居行为而积极地摄取土壤；但对于其他无脊椎动物（螨类螨和collembolans），尚不清楚是否摄入了多少土壤。对于蚯蚓只有一个实验试图从口服暴露于金属区分皮肤，但在不同的条件/金属/污垢/物种更探索需要（•维杰威等人 2003）。在本参考点中，我们认为应该放弃毒物动力学是金属对土壤生物毒性的唯一驱动因素的范式，而我们也应该关注毒物动力学驱动因素。

毒物动力学是毒物与毒作用位点及其后续生物学效应的动态相互作用。通常，对于金属，毒物动力学研究仅集中于毒性作用的部位，例如活性氧的生成或钙稳态的破坏。在土壤环境中，有多种因素会影响土壤生物的健康，例如质地，有机物组成，pH和阳离子交换能力。过去，这些参数很大程度上是通过毒物动力学的角度来观察的。例如，有关阳离子交换容量或有机物的研究集中于这些因素如何影响金属的生物利用度。我们认为我们需要不同的看法，在其中我们明确认识到土壤因素不仅是毒物动力学的关键，而且实际上还可以驱动生物体内的毒理动力学行为（图1）。能量是土壤属性驱动生物体内金属毒性动力学的一个例子。例如，尽管锌的生物利用度相似，但与低栖息地质量的土壤相比，乌鸦纲对高栖息地质量的土壤具有更高的金属耐受性（Jegede等，  2019）。生物体的生态策略也可能影响其对金属污染的反应。例如，某些物种可能会减少生殖产出，以此作为增加金属抵抗力和生存的能源分配的策略（Van Gestel and Hoogerwerf  2001）。尽管已经提出了这一假设（Van Gestel and Hoogerwerf 2001），但尚未通过实验证明，并且这些策略背后的机制尚不清楚。最后，气候是近年来由于气候变化而引起的主要关注，众所周知，气候会通过温度影响生物体对污染的响应，并有望作为生物体生物学的附加压力源。尽管实验是在不同的温度下进行的，但是很少有研究报告温度影响生物体对金属反应的毒动力机制。

目前土壤中金属的风险评估已被破坏。通常，现场观察和实验室毒性测试表明没有对应关系。此外，在现实世界中，金属以混合物形式存在。并且我们现有的解决方法在理论上是不足的，并且在经验上是无法预测的。但是，现场从业人员认识到，通过修订或生态修复措施来提高环境质量可以显着增加受影响地点的生物数量。我们目前的风险评估框架无法解决这一问题，因为它只能解决金属的毒物动力学变化。

一个新的风险评估框架需要一个不同的生态毒理学研究平台，重点关注决定生物对污染物的反应的主要毒理动力学问题。我们可以基于现有的文献探讨金属在毒性作用部位如何相互作用，但是还需要进一步研究金属撞击过程中栖息地质量，生物行为和生物特征如何相互作用。建立后，我们可以轻松地调整现有的风险评估框架，以基于毒物动力学修饰符修改预测的风险。通过这样做，我们可以将毒物动力学研究的最好成果与毒物动力学研究结合起来，从而更好地预测环境中的实际风险。