Keywords

1 General Introduction

Among the long list of compounds included in the Priority Pollutants (PP) list (European Water Framework Directive 2000/60/EC [1]) and those not included but considered also of environmental concern, few studies demonstrate their real impact on fluvial ecosystems. Discriminating the influence of environmental variability from chemical pressure on the biota is a challenging issue. It is necessary to detect joint and specific effects of simultaneously co-occurring stressors. This difficult task requires an interdisciplinary approach including environmental chemistry, toxicology, ecotoxicology, ecology, etc., and also a multiscale perspective from field investigations to laboratory experiments, microcosm, mesocosm studies, and back to the field in order to establish cause-and-effect relationships.

The change of scale is easily achieved working with fluvial communities. Exposure might be done under controlled experimental conditions in microcosms and mesocosms or in the field where the response can be evaluated under real exposure conditions. It allows defining non-effect concentrations, and also assessing the effects of different types of stress alone or in conjunction (multiple stressors). In contrast to single-species tests, ecotoxicological studies with communities involve a higher degree of ecological realism, allowing the simultaneous exposure of many species. Furthermore, it is possible to investigate community responses after both acute and chronic exposure, including the evaluation of direct and indirect toxic effects on entire communities.

In particular, fluvial biofilm and benthic macroinvertebrate communities fulfill these conditions and have been used to investigate single- and multiple-stress situations at different spatial and temporal scales.

Fluvial biofilms (also known as phytobenthos or periphyton) are attached communities consisting of cyanobacteria, algae, bacteria, protozoa, and fungi embedded within a polysaccharide matrix [2]. In rivers, these communities are the first to interact with dissolved substances such as nutrients, organic matter, and toxicants. Biofilms can actively influence the sorption, desorption, and decomposition of pollutants. Fluvial biofilms are relatively simple and easy to investigate compared to other communities (i.e., macroinvertebrate or fish communities).

Biological monitoring programs employed by state and federal agencies to assess effects of contaminants have routinely focused on benthic macroinvertebrate communities. Benthic macroinvertebrates are exposed to contaminants in water, sediment, and biofilm, providing a direct pathway to higher trophic levels. Because of considerable variation in sensitivity among species, community composition and the distribution and abundance of benthic macroinvertebrates are useful measures of ecological integrity.

The main aim of this chapter is to present key studies dealing with priority and emerging pollutants that provided clues to link field observations with causality following a community ecotoxicology approach.

While not being an exhaustive review, 100 different investigations have been analyzed, including field and laboratory investigations, most of them dealing with metals and pesticides (90), but several investigations focused on emerging compounds (10) are also provided (Fig. 1).

Fig. 1
figure 1

Summary of field and experimental investigations addressing the effect of toxic exposure on biofilm and macroinvertebrate communities. Studies are grouped on the basis of the type of chemical investigated: metals, pesticides, and emerging compounds

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2 Field Investigations Used to Generate Hypotheses

Field investigations, although being the basis of ecotoxicological research, have a high uncertainty linked to environmental variability. Fluvial ecosystems are very dynamic and complex systems, requiring the compilation and statistical analysis of extensive sets of samples and environmental data to characterize the exposure and response of their living organisms. This holistic approach, commonly used in ecology, is less extended in ecotoxicology (see [3] for more details).

Uncertainty might be reduced depending on the scale, type of pollution, and set of response variables investigated (Fig. 2). Large-scale studies are influenced by regional scale variability that may conceal the effects of human impacts. Therefore, human impacts may be better detected in small-scale studies, including reference and impacted situations, within a restricted range of environmental variability [124]. The presence of a specific or dominating type of pollution may also contribute to link causes and effects. The effects of toxic exposure, although being present, are more difficult to detect under multiple-stress situations with many chemical compounds occurring at low concentration and environmental stress factors affecting the biota. Choosing an appropriate set of response variables may allow identifying different types of stress and their ecological consequences (Fig. 2).

Fig. 2
figure 2

Scheme showing the main factors influencing the degree of uncertainty in field studies aiming to assess the impact of chemical pollution on natural communities: scale, type of pollution, and set of response variables investigated. Reducing uncertainty will allow to distinguish biological effects caused by toxic exposure from other confounding factors caused by environmental variability, chemical complexity, biological interactions, or co-tolerance

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The link between chemical pollution and ecosystem damage has been addressed in mining areas (e.g., [4, 5]). Metal pollution in other environments is less documented. Influences of pesticide pollution or other contaminants received less attention. Emerging pollutants (a group where pharmaceuticals, nanomaterials, the family of perfluorinated compounds, hormones, endocrine disrupting compounds, pesticide degradation products, and newly synthesized pesticides and others are included) are unregulated pollutants, but may be candidates for future regulation depending on research on their potential health effects and monitoring data regarding their occurrence in ecosystems [6]. Their occurrence in aquatic ecosystems has been recently detected and/or their effects on biota recently investigated [7–9].

Many field investigations on metal, pesticides, herbicides, and emerging compounds’s toxicity were focused on biofilms, the basis of the food chain, and consumers, mainly macroinvertebrate communities.

2.1 Biofilms

Different types of field studies have been performed to assess the effects of metal and organic contamination on fluvial biofilms (Table 1) such as seasonal monitoring [10], biofilm translocation experiments [11, 12], sampling before and after impact [13, 14], and the study of tolerance induction [15–20].

Table 1 Summary of field studies approaching the effects of chemical pollution on fluvial biofilms
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A key aspect in many ecotoxicological studies is to link water concentration with real exposure conditions due to the variety of factors influencing bioavailabilty and/or the stability of a compound. In the case of biofilms, the analysis of the chemical concentration within the biofilm matrix may contribute to partially overcome this constrain. Evidence of the link between metal exposure (water concentration) and metal contents in biofilms is provided in Behra et al. [21]; Meylan et al. [22]; Le Faucher et al. [23], and also between sediment and biofilm metal concentrations [24] (Table 1), highlighting their possible effects through the trophic chain [25]. Internal concentrations are more difficult to assess and less reported for organic toxic compounds [26].

The link between metal pollution and community changes has been shown in many field studies (see [27] for more details). In particular, the effects of metals on diatom communities have been well documented in mining areas. Hill et al. [13] evaluated the effects of elevated concentrations of metals (Cd, Cu, and Zn) on stream periphyton in the Eagle River, which is a mining impacted river in central Colorado. They found differences in diatom taxa richness, community similarity, biomass, chlorophyll a (chl-a), and the autotrophic index (AI) that were able to separate metal-contaminated sites from reference or less impacted sites. Sabater [14] evaluated the impact of a mine tailings spill on the diatom communities of the Guadiamar River (SW Spain) showing a clear decrease of diversity (Shannon–Wiener index) and water quality diatom indices. Other variables such as the size of individuals were also attributed to metal pollution in the Riou-Mort [5, 28].

Biomass reduction due to metal toxicity was suggested by Hill et al. [13]; however, these effects were not confirmed in other highly metal-polluted sites such as the Riou-Mort where metal pollution co-occurs with high levels of nutrients and organic matter (Table 1).

Some investigations reported the induction of metal tolerance after chronic metal exposure [15] using the Pollution Induced Community Tolerance (PICT) approach. Based on the PICT concept, natural communities colonizing different river sites can be used to investigate their exposure history by comparing their physiological responses to a sudden exposure (short-term toxicity tests referred to as PICT tests in this paper) to high doses of the toxicant investigated [29].

Other publications [16] highlight the presence of many selective factors in addition to metal exposure. It may explain the lack of relationship between metal exposure and community tolerance in large-scale studies [30].

Obtaining clear causal relations is difficult in situations of low but chronic metal pollution due to the co-occurrence of many stress factors (Table 1). A low portion of the diatom species composition variance was explained by biofilm metal contents in a small-scale study [31]. Le Faucher et al. [23] found phytochelatines’ (PC) production as a response to low metal exposure, but could not identify the specific metal responsible for that.

Most field studies evaluating the effects of pesticides on biofilms were focused on herbicides (atrazine and its residues, diuron and its residues, prometryn or isoproturon) targeting phototrophic organisms (Table 1). Some investigations (Table 1) evaluated the influence of environmental variability (e.g., light conditions, nutrients, maturity, seasonality, temperature, and hydrology) on biofilm’s tolerance to pesticides (e.g., [17–20, 32–36]). Other studies investigated the capacity of biofilms to be adapted to sites presenting pesticides’ contamination or their capacity to recover from chemical exposure [11, 12]. Monitoring studies aiming to discriminate the effects of pesticides from other stressors are also reported [10, 19]. In the Llobregat River, organophosphates and phenylureas were related with a significant but low (6%) fraction of variance of several biofilm metrics including algal biomass and the photosynthetic efficiency of the community [10]. In the same case study, a significant correlation between the taxonomic distinctness index of diatoms and diuron concentration was obtained [37]. In many other field studies (Table 1), discerning the specific effects of pesticides toxicity on biofilms was complicated by the co-occurrence of pesticide pollution with other types of pollution such as nutrients, organic pollution, or metals (e.g., [12, 19, 38]). On the other hand, the sensitivity of biofilms to pesticides has also been shown to be influenced by biofilm’s age [11, 18] or light conditions [32].

In the case of organic pesticides acting as PSII inhibitors, the PICT approach has been demonstrated to be a valuable tool to assess the effects of chronic exposure [29, 39]. For example, Dorigo et al. [34] showed that the response of algal communities is likely to reflect past selection pressures and suggest that the function and structure of a community could be modified by the persistent or repeated presence of atrazine and isoproturon in the natural environment.

Other pesticides or PPs and emerging contaminants have been poorly investigated (Table 1), probably due to their occurrence at low dose in complex mixtures and the lack of sensitive methods applicable to complex field samples [40].

Overall most ecotoxicological field studies have been focused on structural changes, clearly demonstrated under high pollution conditions. Effects on functional attributes (e.g., photosynthesis) are less clear, probably due to functional redundancy between non-impacted communities and those adapted to chemical exposure (see [41] for more details). Besides, the link between moderate toxic exposure and biological damage is less documented. In these cases, a clear increase of chemical concentration in the biofilm matrix is commonly described, mainly in cases of metal pollution; however, the link with persistent effects on the community is, in most cases, less evident. Moreover, almost all the field studies reported point out the influence of environmental variables on biofilm toxicity. In particular biofilm age, nutrients’ availability, and light conditions have been shown to influence biofilms’ tolerance to various contaminants. This observation highlights the importance of an extensive monitoring, including a multidisciplinary approach to better assess the effects of pollution in complex fluvial ecosystems.

2.2 Macroinvertebrate Communities

Biomonitoring studies conducted with macroinvertebrate communities have been performed to assess effects of metals and organic contaminants on streams (Table 2). Some of these studies have been conducted at relatively large spatial scales [42, 43], allowing investigators to quantify effects of landscape-level characteristics on responses to stressors. These investigators have also developed new approaches based on the biotic ligand model (BLM) to quantify metal bioavailability and effects in the field [43, 44]. Other studies have employed a more traditional longitudinal study design, in which upstream reference sites are compared to downstream contaminated sites [45–47]. Although these single watershed studies are more restricted spatially, they often include investigations that provide a long-term perspective. For example, Clements et al. [47] documented changes in macroinvertebrates and water quality over a 17-year period and related changes in community composition to long-term improvements in water quality.

Table 2 Summary of field studies approaching the effects of chemical pollution on macroinvertebrates
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One of the most consistent observations from these descriptive studies of metals and organic pollution is that certain groups of macroinvertebrates, especially mayflies (Ephemeroptera), are highly sensitive to chemical stressors [42, 45, 48–50]. Interspecific variation in sensitivity to organic and inorganic contaminants forms the basis for the development of new community-level measures to quantify toxicological effects, such as the species at risk (SPEAR) model developed for pesticides and other organics [46, 51]; see also [3]. Recent studies conducted with macroinvertebrates have identified specific morphological and physiological characteristics that are likely responsible for interspecific variation in sensitivity to toxic chemicals [52].

The study of the effects of organic pollutants such as PCBs (polychlorinated biphenyls) and PAHs (polycyclic aromatic hydrocarbons) on macroinvertebrate communities is scarce in the literature, as has been reviewed by Heiskanen and Solimini [53]. Archaimbault et al. [54] used invertebrates’ species traits in order to relate the structure of the community to the presence of PCBs, PAHs, and metals and predict impacted sites by these compounds. As in the case of the SPEAR approach the use of species traits was also useful in order to predict impacted zones. Beasley and Kneale [55] studied the effects of PAHs and heavy metals by multivariate statistical analysis and they found a clear response of the community to the presence of these pollutants with a decrease in diversity.

The effects of pharmaceuticals on aquatic biota have been recently taken into account and have generated concern [56–58]. Despite the existence of experimental evidence of their effects on aquatic invertebrates, the potential effects of these substances on invertebrates’ communities have been poorly studied in the field. Muñoz et al. [59] indicated possible relationships between the presence of some families of pharmaceuticals and the shift in the abundance and biomass of different macroinvertebrate species in freshwater communities.

2.3 Summary of the Hypotheses Generated

Differing in the degree of uncertainty, several hypotheses can be derived from the set of community ecotoxicology field investigations reviewed (Table 3).

Table 3 List of hypotheses derived from field community ecotoxicology investigations
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3 Experimental Studies Searching for Causality

Although biomonitoring studies and other field approaches provide important insights into the effects of contaminants on aquatic ecosystems, these descriptive studies are limited because of their inability to demonstrate cause-and-effect relationships between stressors and ecological responses. Establishing cause-and-effect relationships between stressors and responses is widely regarded as one of the most challenging problems in applied ecology [60]. Because biological assessments of water quality rely almost exclusively on observational data, causal inferences are necessarily weak [61]. In addition, most descriptive studies are unable to identify underlying mechanisms responsible for changes in aquatic communities. Because contaminants often exert complex, indirect effects on aquatic communities, an understanding of underlying mechanisms is often critical. In general, descriptive approaches such as biomonitoring studies provide support for hypotheses rather than direct tests of hypotheses. Equivocal results of biomonitoring studies result from the lack of adequate controls, nonrandom assignment of treatments, and inadequate replication [62].

Various approaches have been developed to address the lack of causal evidence in ecotoxicological investigations. Several investigators have provided useful advice on how to strengthen causal relationships in descriptive studies [63–65]. Weight of evidence approaches, such as the sediment quality triad [66], combines chemical analyses with laboratory toxicity tests and field assessments. These integrated approaches can provide support for the hypothesis that sediment contaminants are responsible for alterations in community structure. Recent approaches developed by the U.S. Environmental Protection Agency (U.S. EPA) employ formal methods of stressor identification, analogous to those used in human epidemiological studies [63], to determine causes of biological impairment in aquatic systems (http://cfpub.epa.gov/caddis/). Hill’s [63] nine criteria and modifications of these guidelines [64, 65, 67] have been employed to strengthen causal relationships between stressors and ecological responses:

  • Strength of the association between stressors and responses

  • Consistency of the association between stressors and responses

  • Specificity of responses to contaminants

  • Temporal association between stressors and responses

  • Plausibility of an underlying mechanistic explanation

  • Coherence with our fundamental understanding of stressor characteristics

  • Analogous responses are observed to similar classes of stressors

  • A gradient of ecological responses are observed as stressor levels increase

  • Experimentation support for the relationship between stressor and responses

Perhaps the most important of these nine criteria is the availability of experimental data to support a relationship between stressors and responses. Regardless of the established strength, consistency, specificity, coherence, plausibility, etc., of the relationship between stressors and responses, there is no substitute for experimentation to demonstrate causality. Experimental approaches provide an opportunity to demonstrate causal relationships between stressors and ecologically relevant responses across levels of biological organization. These experiments are a practical alternative to single-species toxicity tests and address the statistical problems associated with field biomonitoring studies. The maturity of a science such as ecotoxicology is often defined by the transition from purely descriptive to manipulative approaches. The ability to test hypotheses with ecologically realistic experiments represents a major shift in the quality of the questions that can be addressed [68].

Microcosm and mesocosm experiments are often designed to manipulate single or multiple environmental variables, providing opportunities to quantify stressor interactions and identify underlying mechanisms. It is the ability of an investigator to manipulate and isolate individual factors that makes the application of microcosm and mesocosm experiments particularly powerful in ecotoxicological research. Microcosm and mesocosm experiments address two of the key limitations associated with environmental assessment of contaminant effects: the lack of ecological realism associated with traditional laboratory experiments and the inability to demonstrate causation using biomonitoring. In addition, these experimental approaches provide the opportunity to investigate potential interactions among stressors.

Because of the limited spatiotemporal scale, measuring certain responses in microcosm and mesocosm experiments presents significant challenges. The duration of microcosm and mesocosm experiments is of critical importance when assessing effects of contaminants. An important consideration in the development of microcosm and mesocosm approaches is whether greater statistical power and ability to demonstrate causation outweigh their limited spatiotemporal scale. Thus, combining these experimental approaches with results of field studies conducted at a larger spatiotemporal scale is the most reliable way to demonstrate causation.

3.1 Experimental Studies Addressing Cause-and-Effect Relations Between Toxicant Exposure and Biofilm Communities Responses

Different experimental designs have been used to investigate, under controlled conditions, the effects of toxic substances on biofilm communities (Table 4). The scale of exposure ranges from hours to weeks, either continuous (the most common experimental design) or pulsed. After controlled toxic exposure, the PICT approach is often used to assess differences in sensitivity at community level with the same procedure applied to natural biofilms (e.g., see Sect. 2.1).

Table 4 Experimental studies addressing cause–effect relations between toxic exposure and their effects on biofilms
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Many authors have investigated metal and pesticides’ effects in microcosms and mesocosms by adding toxicants to natural water (Table 4). This experimental approach was used to investigate the single effect of a toxicant, the effect of mixtures, the bioaccumulation (only for metals), the tolerance and co-tolerance induction (using the PICT approach), and also the interaction between toxicity and environmental and biological factors such as nutrients, light conditions (including UV radiation), or grazing (Table 4). Effects of the most commonly reported metals in the environment: Cu, Zn, Ni, Ag, Cu, Pb, and Cd were investigated at environmental realistic concentrations. As in the field studies reported above, most of the pesticides studied were herbicides (atrazine, diuron and its residues, isoproturon, prometryn) targeting photosynthetic processes and hence the sensitivity of the phototrophic component of biofilms to these compounds (Table 4).

Investigating pesticides’ effects at different temporal scales from hours, days [69, 70], to weeks [39, 71–74] or months [75] allowed to assess the physiological and structural biofilm responses. Herbicides’ indirect effects on invertebrates [71, 74] or bacterial communities [73] were also evaluated. Among the long list of the so-called emerging pollutants that are commonly found in environmental samples, few pharmaceuticals and personal care products were selected to investigate their potential effects on biofilms (Table 4), the bactericide triclosan (TCS) being the most investigated compound [76–79]. These studies addressed the protective role of biofilms and the effects of TCS on structural and functional attributes of their heterotrophic and autotrophic component. Short-term effects of three similarly acting β-blockers were investigated by Bonnineau et al [80] using a multibiomarker approach (Table 4).

Overall, the experiments reported tested an important number of hypotheses on toxicant availability, their effects on biofilm communities, and the potential modulating factors (Fig. 3). It was shown that metal bioaccumulation was driven by speciation and the presence of ligands [81]. Negative effects on growth were also reported (Table 3) as a reduction in biomass production (i.e., [82]), diatoms density [75, 83], and in diatoms biovolume [73]. It was also shown that prolonged exposure to a toxic compound alters the community composition causing a selection of tolerant species in most cases (e.g., [83, 84]), an increase in the tolerance of the community in many cases (e.g., [85–87]), and the production of EPS in some occasions [88]. Other investigations also demonstrated that these community responses were modulated by different environmental and biological factors (Fig. 3).

Fig. 3
figure 3

Figure summarizing experimentally tested hypotheses on (1) toxicants’ availability; (2) the biological and environmental factors modulating toxicity; and (3) the toxic effects after prolonged exposure of biofilm communities to different types of toxicants. M indicates hypotheses mainly tested with toxic metals; H for those tested with herbicides. No label for those tested for at least a metal, an herbicide, and an emerging compound

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That biofilm’s maturity and/or thickness may reduce toxicity is a general rule demonstrated for metals (i.e., [83]), herbicides (e.g., [87]), and emerging compounds such as TCS [76]. On the other hand, increased herbicide toxicity was demonstrated under grazing pressure [71], linked with the observation that grazing was responsible for maintaining the biofilm younger (Fig. 3).

Nutrient concentration may also influence the effects of toxic exposure (Fig. 3). It was shown for toxic compounds like Zn or Cu directly affecting nutrients availability (e.g., [89]).

Light history as well as light conditions during toxicant exposure affected toxicity. It was shown by the increase in sensitivity to Cd observed in biofilms adapted to UVR [90] or the increase in toxicity of isoproturon under a dynamic light regime [70]. In the case of Cu, co-tolerance to other metals was also demonstrated [85].

The toxicity of three similarly acting β-blockers was confirmed, but at very high concentration. In this study, differences in toxicity between the three compounds highlight the need to increase ecological realisms in toxicity testing used to derive environmental quality standards [80].

3.2 Experimental Studies Addressing Cause-and-Effect Relations Between Toxicant Exposure and Macroinvertebrates

Experimental studies have been conducted in aquatic microcosms and mesocosms to examine the effects of heavy metals, pesticides, and other organic chemicals on the structure and function of macroinvertebrate communities (Table 5). Although the duration of most studies was relatively short (1–4 weeks), the duration of several mesocosm and field experiments was considerable longer [91–93]. Endpoints measured in these studies included responses across several levels of organization, from physiological effects to alterations in community structure and ecosystem function. Some of these studies were conducted to provide additional support for field studies, thereby strengthening arguments for causation [60], while others were conducted to verify results of laboratory toxicity tests [46, 94]. Several studies were conducted to identify direct sources of toxicological effects in systems receiving multiple stressors [95, 96] or to quantify the influence of natural habitat characteristics on contaminant effects [97, 98]. Also experiments to examine competition/predation interactions between populations after exposure to pollutants have been developed [99]. A common goal of many experiments was to examine interactions among stressors [91] or to quantify the potential cost of tolerance associated with contaminant acclimation or adaptation [100–102]. A consistent justification for the application of these ecologically realistic approaches in risk assessment was to integrate responses across multiple levels of biological organization [93, 103].

Table 5 Experimental studies addressing cause–effect relations between toxic exposure and their effects on macroinvertebrate communities
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3.3 Integrating Descriptive and Experimental Approaches to Demonstrate Causation in Stream Bioassessment Studies: Case Study of a Metal-Polluted Stream

To demonstrate how descriptive and experimental approaches can be integrated to demonstrate causation, we present results of a case study conducted in a metal-polluted stream in Colorado, USA. A long-term monitoring program of water quality and benthic macroinvertebrates was initiated in the Arkansas River in 1989 [47]. Concentrations of heavy metals (Cd, Cu, and Zn) are greatly elevated downstream, and often exceed acutely toxic levels. Over the past 17 years, we measured physicochemical characteristics, habitat quality, heavy metal concentrations, and macroinvertebrate community structure seasonally (spring and fall) at locations upstream and downstream from several sources of metal contamination. Three years after we began this research, a large-scale restoration program was initiated to reduce metal concentrations and reestablish trout populations in the upper Arkansas River basin. Because these data were collected before and after remediation, this long-term research provided a unique opportunity to quantify ecological responses to improvements in water quality. For the purposes of this case study, we present macroinvertebrate and metals data collected from one upstream and one downstream station collected from 1989 to 2006.

To support this descriptive study, microcosm experiments were conducted to develop concentration–response relationships between heavy metals and measures of macroinvertebrate community structure. Benthic macroinvertebrate communities for these experiments were obtained from an uncontaminated reference stream with no history of metal contamination using a technique previously described [104]. Communities were transferred to stream microcosms and exposed to combinations of Cu and Zn at concentrations that bracketed those measured at metal-contaminated sites in the field. Because these experiments involved a mixture of heavy metals, an additive measure of toxicity was used to express metal concentrations relative to the U.S. EPA chronic criterion values. The cumulative criterion unit (CCU) was defined as the ratio of the measured metal concentration to the hardness-adjusted criterion value and summed for each metal.

Significant concentration–response relationships were developed for total abundance and species richness in stream microcosms (Fig. 4). Total macroinvertebrate abundance was more sensitive to metals than species richness, a finding previously reported from stream microcosm experiments [105]. The LC20 concentrations for macroinvertebrate abundance and species richness, defined as the CCU levels that caused a 20% reduction, were approximately 2.3 and 9.0, respectively. These experimental data support the hypothesis that macroinvertebrate communities were highly sensitive to heavy metals and that relatively low concentrations resulted in significant alterations in community composition.

Fig. 4
figure 4

Results of microcosm experiments showing concentration–response relationships between metal concentration (CCU), total macroinvertebrate abundance, and species richness. Dashed lines show the estimated LC20 concentrations for each metric

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Heavy metal concentrations in the Arkansas River were seasonally variable, but decreased significantly after the remediation program was initiated (Fig. 5). Metal concentrations at station EF5 rapidly decreased below the estimated EC20 values in 1992. In contrast, metal concentrations at station AR3 remained elevated until about 1999, fluctuated between the estimated EC20 values for several years, and then decreased below these levels in 2003.

Fig. 5
figure 5

Long-term changes in metal concentration (CCU), abundance, and species richness at Arkansas River stations EF5 and AR3. Horizontal lines correspond to the EC20 values for species richness (dashed) and abundance (dotted) estimated from microcosm experiments. Alternating solid and open symbols refer to samples collected in spring and fall, respectively. Macroinvertebrate data are means ±s.e.

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Macroinvertebrate communities in the Arkansas River responded to these improvements in water quality after remediation (Fig. 5). Macroinvertebrate communities at station EF5 quickly recovered after metal levels were reduced below the EC20 values. Abundance and species richness recovered more slowly at station AR3, consistent with expectations based on long-term changes in metal concentrations and estimated EC20 values for these metrics. Total macroinvertebrate abundance recovered in approximately 2003, whereas species richness recovered in 2000.

Although the biomonitoring results from the Arkansas River were consistent with the hypothesis that heavy metals are responsible for reduced abundance and richness, these data are not sufficient to demonstrate causation. Long-term improvements in water quality and the associated increases in abundance and species richness after remediation represent a natural experiment that allowed a more rigorous test of this hypothesis. Finally, highly significant concentration–response relationships between macroinvertebrate community metrics and heavy metal concentration from stream microcosm experiments provided estimates of metal levels that likely impact benthic communities. The consistency of these LC20 estimates with concentrations measured in the field where abundance and richness recovered greatly strengthened the argument that heavy metals were responsible for alterations in benthic communities. These results demonstrate the importance of employing ecologically realistic experimental techniques to support descriptive studies for developing causal arguments.

4 General Discussion and Prospects

The papers found in the literature show how field and laboratory investigations evolved from the end of the 1990s, when community ecotoxicology studies focused on metals and pesticides began, to present days, when the first emerging compounds investigations have appeared (Fig. 1). To date, quite a lot of information has been generated about the effects of emerging pollutants (mixtures or simple compounds) on single organisms but community approaches are very scarce, highlighting the existence of a missing gap requiring future investigations.

The set of studies presented in this chapter illustrate how field and laboratory investigations complement to provide causality between toxicant exposure in running waters and benthic communities’ responses. As a general scheme (Fig. 6), hypotheses based on field observations cannot be confirmed without experimentation and thus needed to provide causal evidence. If the formulated hypothesis is not confirmed, new field observations may be required and the results should be analyzed again in order to evaluate the role of different environmental factors (confounding factors) influencing the biological responses observed. The newly generated hypothesis should also be experimentally tested including possible interactions by simulating multiple-stress situations (Fig. 6).

Fig. 6
figure 6

Graph illustrating the steps required to derive causality. Hypothesis formulated on the basis of field observations should be experimentally tested for confirmation. If the formulated hypothesis is not confirmed, new field observations may be required and the results analyzed again in order to evaluate the role of different environmental factors (confounding factors) influencing the biological responses observed. The newly generated hypothesis should also be experimentally tested including possible interactions simulating multiple-stress situations

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Many of the hypotheses derived from field investigations have been validated following this general scheme. The Arkansas River case study is a good example showing how field observations, together with long-term natural experiments and microcosm experiments, provide consistent arguments and evidence of metals’ effects on the biota. More precisely, metal pollution levels were responsible for reduced abundance and taxa richness of the macroinvertebrate community. Long-term monitoring studies following temporal trends in both chemical pollution and the biological responses are very informative, but unfortunately very scarce. In the case of biofilms, microcosm and mesocosm experiments confirmed that metals and pesticides are responsible for the loss of sensitive species in the community, and that this influence is modulated by several biological and environmental factors including the successional stage (biofilms age), the trophic status of the river or stream (nutrient concentration), and the light regime (whether is it an open canopy or shaded reach). These concluding findings should be included, in the future, in risk assessment models in order to account for the influence that the environment exerts on the effects of chemicals on the biota.

In contrast to the results obtained with metals and pesticides already included in the list of priority pollutants, emerging pollutants and especially pharmaceuticals (molecules designed to be biologically actives) are not expected to cause changes on natural communities easily to detect. Given that the levels reported in rivers for these substances in waters and sediments are generally low, no lethal effects on the species are expected at concentrations found in the environment [56, 106]; hence, new approaches should be used in order to know the effects of these substances on natural communities. Experimental investigations on communities are difficult because of the long-term studies required and because of the inconspicuous endpoints that need to be studied, whereas in field studies the difficulties in predicting the effects of emerging pollutants and changes in community arise from the nature of the effects caused by these substances, such as feminization or changes in behavior or emergence time, not studied enough in invertebrates in wildlife. Effects in communities can be detected by examining other mechanisms than the direct effect in species’ density (like in the case of pesticides) such as sublethal or long-term effects on physiology, on reproductive traits or hormone-mediated processes (Endocrine Disrupting Compounds as explained in Lagadic et al. [107] and Soin and Smagghe [108]), or on other less obvious traits such as behavior [109].

Overall, the examples provided in this chapter, together with the recommendations given, are proposed as a general guide for studies aiming to link chemical pollution with ecological alterations. The proposed approach, although being complex and probably expensive in terms of dedication, is strongly recommended for investigative monitoring, situations where routine monitoring may fail in the detection of the causes accounting for a poor ecological status.