Abstract
Many of the chemicals in the environment are naturally derived from compounds in plants, petroleum oils, or minerals in rocks. However, their chemical composition, concentration, and distribution through the environment have been altered by humans, usually as a result of an economic incentive (e.g., mining).
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3.1 Introduction
Many of the chemicals in the environment are naturally derived from compounds in plants, petroleum oils, or minerals in rocks. However, their chemical composition, concentration, and distribution through the environment have been altered by humans, usually as a result of an economic incentive (e.g. mining). Other chemicals are synthetic, produced in laboratories, and manufactured for specific uses. These manufactured chemicals are known as xenobiotics and include some fertilisers, pesticides, dyes, manufactured petroleum products, personal care products, and pharmaceuticals. What is common to all natural and Synthetic chemicals is that they are potentially toxic and likely to have come from a small geographic area or a limited number of sources. The chemicals are then redistributed in the environment through natural and anthropogenic activities, where organisms can intentionally or unintentionally be exposed to them. Some exposed species, and indeed some individuals, will be more sensitive than others, which can lead to adverse effects at the population level. When sensitive species are keystone or foundation species for a particular ecosystem, or enough species are affected, this can alter the structure and function of the exposed communities, having flow-on effects at the ecosystem level (Figure 3.1).
3.2 Ecotoxicology
Toxicity testing of organisms has been developing since the 1940s (Cairns Jr and Niederlehner 1994) because of the need to understand the effects of chemicals on organisms. Its application in environmental monitoring has grown rapidly (e.g. Auffan et al. 2014). In fact, the term ecotoxicology, which the field of study is now referred to as, was first used in 1969 by René Truhaut, defining it
“as a science describing the toxic effects of various agents on living organisms, especially on populations and communities within ecosystems”
Ecotoxicology is a multidisciplinary science that combines chemistry, biology, ecology, pharmacology, epidemiology, and of course toxicology. It seeks to understand and predict the effects of chemicals on organisms and ecosystems and is constantly evolving as a discipline area (Sánchez-Bayo et al. 2011). Pollution studies use ecotoxicology as a tool to document the effects of pollutants at known concentrations on living organisms (Phillips 1977; Chapman and Long 1983) and to supplement conventional pollutant concentration data.
Ecotoxicology is used in a multiple lines of evidence (LOE) approach to risk assessment. This means that you use more than one source of information to support and understand the risk. Ecotoxicological experiments are most often conducted in laboratories under controlled conditions, and this chapter provides a guide to how these experiments are performed. While the resulting information is limited by its lack of relevance to conditions in the environment, it provides important standard approaches for comparative assessment to help understand the relative sensitivity of different species and the relative toxicity of chemicals. Standard approaches to toxicity testing are explained in Section 3.2.3. Non-laboratory approaches for assessment under more relevant environmental conditions are discussed in Section 3.6.1.
The expansion of ecotoxicology to develop species sensitivity distribution (SSD) using toxicity data for a range of species and taxonomic groups (for detail see Section 3.4) increases the ecological relevance of laboratory-based toxicity results. The greater the number and diversity of species used in an SSD the greater the confidence in predicting concentrations that should protect any chosen percentage of species (see Section 3.4). Another means of gaining a better understanding of ecological interactions through ecotoxicological assessment is by using microcosm or mesocosm level studies in the field or laboratory (Box 3.1).
3.3 Current Status of Marine Ecotoxicology
The vast majority of aquatic ecotoxicological data is for freshwater species because humans have been aware of contamination of fresh water for much longer. We have had a much greater investment in fresh water, and hence pollution of freshwater ecosystems has historically been more relevant to us and more noticeable. As discussed in Chapter 1, it is only relatively recently in human history that we have become aware of the effects of pollution in marine ecosystems, and the vast majority of marine data is for temperate, Northern Hemisphere species, because that is where most ecotoxicology has been conducted (Lacher and Goldstein 1997).
3.3.1 Temperate Marine Ecotoxicology
As you can see from Table 3.1, most species used in sediment toxicity testing are temperate, and this is the same for aquatic toxicity tests. Because a lot of toxicity data exists for temperate marine species, and ecotoxicological risk assessment is much more relevant when a larger amount of data are available, considerable effort has been directed towards understanding whether temperate data can be applied for the ecosystem protection in other climatic regions. Research has illustrated that there are no predictable patterns in toxicity between temperate and tropical, or temperate and polar species (Chapman et al. 2006; Wang et al. 2014). Rather, it is evident that the relative toxicity depends on the contaminant (i.e. some metals are more toxic to tropical species than temperate species, and vice versa) (Kwok et al. 2007).
3.3.2 Polar Marine Ecotoxicology
Despite the remoteness and isolation of polar regions from the centres of anthropogenic activity, contamination is increasingly being identified in these regions, including in deep ocean sediments (e.g. Isla et al. 2018), benthic organisms, and in the tissues of organisms high in the food chain (e.g. polar bears and other mammals, large seabirds [e.g. Eckbo et al. 2019], and sharks [Ademollo et al. 2018]). While there are a few isolated point sources of contaminants (e.g. sewage and other waste from research stations, fuel, and oil), the primary concern and challenge is that contaminants are being transported to polar regions in ocean currents, in the atmosphere, (refer to Chapter 7 for more detail), and by trophic transfer. These dispersed contaminants include a wide range of persistent organic pollutants (POPs) (see Chapters 7 and 8), plastics (including micro- and nano-plastics, microfibres, and their degradation products) (e.g. Mishra et al. 2021) (see Chapter 9), and pesticides (see Chapter 7).
Unfortunately, to date, ecotoxicological risk assessments for polar environments are constrained by the very limited amount of regionally relevant toxicity data. Consequently, they are mostly derived from extrapolations of temperate and tropical toxicity data. There is a valid argument that extrapolation of data from other regions is better than no data at all; however, taxonomic compositions, chemical toxicity, and organism physiology are extremely different in the consistently low temperatures experienced in polar regions (e.g. Kefford et al. 2019). Obtaining the necessary toxicological data to enable the development of relevant water quality guidelines for these ecosystems is currently the subject of dedicated research effort (e.g. King et al. 2006; Gissi et al. 2015; Alexander et al. 2017; Koppel et al. 2017; Kefford et al. 2019; van Dorst et al. 2020).
Generally, polar species are more sensitive to long-term exposure to contaminants than tropical or temperate species. Chapman and Riddle (2005) suggest the following possible reasons for this:
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many species have relatively long lifespans and long development times (and so have a long time to accumulate contaminants);
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many species are relatively large, exhibiting gigantism, which may influence the response time (and so have a slower uptake of contaminants due to the low surface-area-to-volume ratio);
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slower metabolic rates and slow uptake kinetics (resulting in slower accumulation of contaminants, but also slower detoxification/depuration);
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less energy consumed (so less energy is available for detoxification/depuration); and/or
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high lipid content (so accumulate higher concentrations of lipophilic contaminants).
Some of these factors also affect the way that toxicity tests need to be conducted. For example, toxicity tests may need to be continued for longer time periods to account for slower metabolism and slower transition through different life stages, and different endpoints or assessments may be needed for lipophilic substances (Chapman and Riddle 2005).
3.3.3 Tropical Marine Ecotoxicology
Tropical marine ecosystems have a very different taxonomic composition, biodiversity, and physiology of organisms compared to temperate marine ecosystems. Some tropical marine ecosystems have extremely high levels of biological complexity, organisation, and diversity. For example, the tropical area known as the Coral Triangle (a marine region that spans parts of Indonesia, Malaysia, Papua New Guinea, the Philippines, the Solomon Islands, and Timor-Leste) is recognised as a global hotspot of biodiversity for corals and reef fishes (Allen 2007). Hence, there are more species that are susceptible to being exposed to contaminants, as well as more complex ecological interactions which further complicates risk assessment. Many tropical marine waters are oligotrophic (see Chapter 4), which provides less opportunities for contaminants to form complexes and so can result in higher bioavailability. Physiologically, organisms generally have a higher metabolism in warmer temperatures, which can increase either or both uptake and detoxification of contaminants. Degradation (biological and abiotic) might be expected to be enhanced in tropical marine systems compared to temperate and polar marine systems; however, research by Mercurio et al. (2015) found that five herbicides had half-lives of greater than 1 year in tropical marine water compared to earlier studies in temperate laboratories that reported half-lives of months.
Ecotoxicology in tropical marine environments is limited and there is a dearth of data on the dose–response characterisations of pollutants, particularly for early life stages. Lacher and Goldstein (1997) discussed the rapid increase in agricultural, urban, and industrial development in tropical regions. Peters et al. (1997) stressed that managers of tropical marine ecosystems have few tools to aid in decision-making and policy implementation and presented conceptual models as a future tool for the problem formulation phase of ecological risk assessment. Measurable responses to stressors, such as the concentrations of chemicals (i.e. ecotoxicological studies), are used within these models and are pertinent to the decisions that may be made to protect the environment (Peters et al. 1997). Since the study by Peters et al. (1997), some progress has been made in developing an understanding of the impacts of trace metals on tropical species (Chapman et al. 2006). However, there is a paucity of fully developed regionally relevant marine toxicity testing methods for tropical marine systems (e.g. van Dam et al. 2008). Fortunately, research effort is growing in tropical marine ecotoxicology.
3.4 Using Ecotoxicological Data to Set Guideline Values
As the preceding text has shown, chemicals, if present at sufficiently high concentrations, can cause a diverse range of harmful effects. Largely as the result of some particularly disturbing pollution events in the United States of America, the United States Environmental Protection Agency (US EPA) started to develop maximum concentrations of chemicals in the water that are safe or provide a high degree of protection to aquatic ecosystems. Subsequently, numerous countries, states, and provinces have developed similar limits for chemicals in water, soil, sediment, and animal tissue in order to protect ecosystems. These limits are called guidelines, criteria, standards, or objectives, depending on the legal framework of the jurisdiction developing the limits. Although these terms are often used interchangeably, they have different meanings. Criteria and standards generally have some legal standing and if they are exceeded, this can lead to prosecution in courts of law. Guidelines do not have any legal standing but rather provide guidance on what is a safe concentration. Typically, if environmental concentrations are greater than a guideline concentration, then further work is required. This work can take several forms such as the development of management actions to decrease the concentration, amount, or type of chemicals released, or investigations to determine if the guideline is appropriate or if there are special conditions at the site that may increase or decrease the degree of protection provided. Criteria, standards, and guidelines are all based on the available scientific information, but scientific information is only one of the multiple factors that may be considered in deriving objectives. Other potentially relevant factors include costs and benefits, commercial considerations, and religious and cultural values. Objectives, criteria, standards, and guidelines are limits that reflect the management goals for a particular part of the ecosystem. For the remainder of this section, the term limit will be used generically to mean guidelines, objectives, criteria, and standards.
3.4.1 Deriving Limits
There are three main methods for deriving limits: background concentrations, assessment or safety factors (AF or SF), and SSDs.
Background Concentration Method
This method determines a fixed percentile (e.g. the median or 90th percentile) of the background concentration of a chemical and adopts that as the limit. While this is conceptually straightforward, it is often quite complex to obtain background concentrations, particularly in areas with a long history of human activity (e.g. in-shore regions near major urban developments). However, while they may not be relevant for particular sites, publications or databases of background concentrations are often available.
Assessment Factor Method
The assessment factor (AF) method requires a literature search for available data on the responses of marine organisms to toxicants. The data are then screened and assessed for quality, and inappropriate and/or low-quality data are removed. Then the lowest toxicity value is identified and divided by an AF to derive the limit. The magnitude of the AF depends on the amount and type of toxicity data that are available (Table 3.3). Basically, an AF of 10 is applied to account for the following: a lack of data, the difference between toxicity values from acute (short-term) and chronic (long-term) exposures, and differences in toxicity data from laboratory-based and field-based experiments (Table 3.3). This method is easy to understand, and the resulting limit will prevent any of the toxic effects reported in the literature, but it can lead to very low limits.
A key criticism of the AF approach is that there is little scientific justification for the magnitude of the AFs. A critical assessment of the strengths and weaknesses of the AF methods is provided in Warne (1998). Experimentally determined AFs termed acute to chronic ratios (ACRs) have been developed to convert acute toxicity data to chronic data. However, while these are better than the above default AF of 10, they also have limitations (Warne 1998).
Species Sensitivity Distribution Method
The newer and currently preferred method for deriving limits is the SSD approach. This approach was developed in 1985 by Stephan and colleagues (Stephan et al. 1985) and has subsequently been extensively improved. All these SSD methods require a thorough search of the literature, followed by screening and assessing the quality of the toxicity data. The data that pass the screening and quality assurance process are then manipulated to obtain a single value to represent each species for which are data available (e.g. Saili et al. 2021). The data are ordered from highest to lowest toxicity (i.e. lowest to highest concentration at which toxic effects occur) and then given a ranking increasing from one. The cumulative frequency (a percent value) for each species is then calculated by:
Cumulative frequency = rank/(n + 1) × 100.
where n is the number of species for which toxicity data are available (i.e. the highest rank number). Thus, if there are toxicity data for 10 species, the cumulative frequency values for the first three species would be 9.09% (1/11×100), 18.18% (2/11×100), and 27.27% (3/11×100). The cumulative frequency values for each species are then plotted against the toxicity value representing the species and a statistical distribution is fitted to the data. Once the statistical distribution that best fits the data has been identified, that distribution is used to calculate the concentration that corresponds to protecting any selected percentage of species (conversely, the concentration that will permit a certain percentage of species to experience adverse effects). An example SSD for a hypothetical toxicant to marine species is presented in Figure 3.6.
The usual percentages of species selected to be protected for toxicant limits are 99 and 95%, and the usual limits of species that are permitted to be harmed are 1 and 5%. The concentrations that correspond to these levels of protection are termed the protective concentrations for 99 and 95% of species (i.e. PC99 and PC95, respectively) or the harmful concentrations for 1 and 5% of species (i.e. HC1 and HC5, respectively). While these are the most commonly used levels of protection, it is possible to calculate the concentration that corresponds to any percentage of species desired to be protected or harmed. Examples of the protocols for deriving limits are those used in Canada (CCME 2007) and Australia and New Zealand (Warne et al. 2018), and the software packages used to generate SSDs include ssdtools (Thorley and Schwarz (2018) and Burrlioz (CSIRO 2016). A critical assessment of the strengths and weaknesses of the SSD approach and the validity of its assumptions is presented in Warne (1998).
3.5 Limitations of Species Toxicity Studies
There are many limitations with toxicity studies and these are principally related to the fact that they are conducted in laboratories. Some of the limitations are as follows:
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The experimental conditions are highly standardised and controlled to minimise variation (i.e. not like the real world). For example, toxicity tests try to maintain the concentration of the test chemical for the duration of the test. This makes calculations of the toxicity simpler, but organisms in the environment are exposed to concentrations that change over time.
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The experimental conditions are usually optimal for the species, whereas that is often not the case in the environment and could lead to an underestimation of toxicity.
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The species used in toxicity tests have often been chosen because of their ease of being cultured in the laboratory or other pragmatic considerations. Such organisms may not be the most sensitive to chemicals. Rare and endangered species are seldom used in toxicity tests—yet these might be organisms that are important to protect.
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Toxicity test methods have only been developed for some species and are therefore biassed with many important organism types not being included, or there is a marked bias in the proportion of organism types with toxicity data.
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The duration of short-term (acute) tests is based on pragmatic considerations rather than biological reasons. For example, many acute tests are of 96 h duration to permit a toxicity test to be established and completed in a working week.
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Most toxicity tests only expose individuals of a single species and can therefore only measure the direct effects of chemicals on the test organism. Whereas, in the real-world, multiple species will be simultaneously exposed and both direct and indirect effects of chemicals on the test organisms can occur.
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Most toxicity tests only expose the test organism to a single chemical, whereas in reality, organisms are usually exposed to mixtures of chemicals (e.g. Warne et al. 2020). This can lead to underestimation or overestimation of the harmful effects caused by chemicals.
3.6 Assessing Responses from Organisms at the Community Level
So far in this chapter, we have focussed on single-species ecotoxicological assays, with these generally based on the exposure to one or a small number of toxicants under stable laboratory conditions. Ideally, these are designed to provide guidance about the exposure and the concentrations of contaminants at which toxicity commences which can then be used to derive chemical limits and protect marine environments. However, as discussed above (Section 3.5), these approaches are not without limitations, and extrapolating laboratory-derived predictions about the adverse effects of contaminants at specific concentrations is fraught with ambiguity. This is not only because organisms are generally exposed to multiple stressors in the field (see Chapter 14) but also because environmental protection focuses on communities rather than individual species.
Even without pollutants, marine communities are complex and dynamic systems. Species migrate and immigrate, and interact with each other, sometimes favourably (e.g. symbiotic or mutualistic relationships), other times, less favourably (e.g. predation, parasitism, disease, and competitive displacement). Overlaying these complex interactions are a myriad of abiotic conditions (e.g. seasonality, substrate differences, temperature, depth, salinity, etc.). In many cases, these variables can also be stressors, albeit natural stressors. For example, while many organisms are accustomed to residing in relatively stable marine waters, living in estuaries is far more challenging, with marked tidal changes in salinity, temperature, pH, etc., and only a relatively small number of species are physiologically equipped to deal with such conditions. Marine communities change over space and time, and the challenge for scientists is being able to distinguish the effects of any contaminants over the natural variation. This will assist in determining whether the ecological impacts of the contaminants are significant and, therefore, whether the contamination is deemed pollution. Ideally, the assessment would identify which contaminant(s) are driving any observed changes. Here, we will discuss three approaches for examining the effects of contaminants on marine communities: in situ (field) surveys; experimental in situ studies; and community-level laboratory studies.
3.6.1 In situ Studies
Logically, the most common approach for examining the potential effects of contaminants on marine communities is by in situ surveys, also known as field studies. Given that marine ecosystems can encompass many different types of environments (e.g. seawalls, pelagic, coral, soft-substrate, intertidal, abyssal, etc.), each with its own range of communities, it is imperative to first establish which communities or assemblages (group of taxonomically related species, e.g. fish) should be targeted. This decision should be driven by a number of factors, including the following:
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whether a particular community or assemblage has a high conservation, socio-economic, or other values (e.g. key diet species of local people);
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the type of contaminant and its primary exposure pathway;
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accessibility, time, and cost to collect and process samples;
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relevant taxonomic and ecological expertise; and
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whether sufficient and representative community-level samples can be obtained.
Hence, it is essential that the targeted community is of ecological and ecotoxicological relevance to the potential pollutant(s).
While the types of marine communities captured in ecotoxicological studies are highly varied and can include rocky tidal platform communities, fish, and even the microbiome associated with particular host species (e.g. sponges [Glasl et al. 2017]), the most common approach is to examine the macrobenthic communities associated with soft-bottom sediments. This includes taxa such as polychaetes, amphipods, bivalves, and gastropods. This is because, firstly, sediments often contain far greater concentrations of contaminants than the water column and, thus, can pose a significant risk to the whole ecosystem. Secondly, macrobenthos interact with the sediment via consumption or residing within it, and therefore may experience multiple exposure pathways. Thirdly, macrobenthos are numerous and diverse, and therefore not only capture a wide variety of life strategies and sensitivities, but also are generally in numbers sufficient for robust statistical analysis. Fourthly, because of their size and historic use, they tend to be relatively easy to identify at the family level of taxonomic rank and higher. Fifthly, macrobenthic invertebrates are also generally relatively sessile, and therefore, their composition reflects the condition of the environment they were sampled in. Finally, and importantly, benthic communities are a crucial part of near-shore food webs, and consequently, changes in their composition may have cascading effects on other components of the system (e.g. fish) (Antrill and Depledge 1997; Fleeger et al. 2003).
While macrobenthic communities are typically the focus of in situ field studies, it is important to reiterate that the choice of targeted community in any in situ study is dependent on several factors and is by no means limited to macrobenthic invertebrates. In fact, there is an increasing trend to use DNA-based approaches such as metabarcoding to capture a far wider range of taxa than can be obtained using traditional means (Chariton et al. 2010b; Cordier et al. 2020; DiBattista et al. 2020).
Broadly speaking, two different approaches are most commonly used in in situ community surveys: reference/condition and gradient (Quinn and Keough 2002; Chariton et al. 2016). The first approach is a comparison of reference and condition sites. For example, the composition of the targeted communities from several relatively unmodified reference sites is compared to those from several sites exposed to the contaminant(s) of interest (e.g. sites with elevated copper derived from mine tailings). It is emphasised that multiple sites are required for each treatment, with this approach being founded on a factorial design, like those where an Analysis of Variance (ANOVA) can be applied. One of the biggest challenges of this approach is finding replicated reference locations, which is becoming increasingly difficult with the increasing loss of natural marine habitats. Careful consideration is required when choosing reference sites, as it is possible that other variables (e.g. grain size, seagrass cover, and unmeasured contaminants), and not the stressor of interest, may be driving any potential differences between the reference and impacted sites. This issue was highlighted in a comprehensive survey of North Carolina estuaries (Hyland et al. 2000) where the authors found that benthic assemblages were impaired in approximately one-quarter of sites (27%) even though no significant concentrations of contaminants were observed. This suggests that in these cases the communities were either being modified by natural variables, unmeasured contaminants, or a combination of the two.
In contrast to the factorial design which underpins the reference/condition sites approach, gradient studies aim to detect variability along a dominant pollution gradient. For example, sites are sampled at increasing distances from the deposition point for deep-sea tailings. Indeed, the approach can be used to capture multiple gradients, both natural and anthropogenic, with an increasing number of statistical tools becoming available which enable scientists to identify the proportion of variation in the community data which can be explained by the measured contaminants and other environmental variables (Chariton et al. 2010b). Gradient studies can be expensive and time-consuming, requiring sufficient environmental data (e.g. metals, pesticides, and natural stressors) and community data to capture the correlative patterns which underpin the gradient(s). However, if designed and implemented properly, they can provide key insights into how communities may be being shaped by both natural and anthropogenic variables, as well as their interactions. Preliminary or investigative studies that provide a reasonable understanding of the factors that may be at play will help in designing gradient studies.
At first glance, it would be logical to assume that either in situ approaches would result in determining whether a contaminant is causing the observed negative impairments to the marine community. However, this is not the case, as in situ studies are correlative and consequently cannot be used to state causality. For example, even if there was a very strong correlation between sediment copper concentrations and benthic diatom communities, there may be other reasons for the observed trend, such as another unmeasured contaminant or natural variation. What in situ studies can tell us is that there is evidence that the contaminant is causing an adverse effect and is acting as a pollutant. Very much like a legal court case where no single LOE can be used to make a verdict, additional LOEs are required to state with a high level of certainty that the contaminant is causing an effect. This evidence may be obtained from ecotoxicological data such as described earlier in this chapter, bioaccumulation and biomarker studies, or via the use of manipulative experiments.
To reiterate, in situ studies are an important tool for helping to determine whether contaminants may be negatively impacting a marine community and therefore causing pollution. However, they cannot determine causality, and thus their findings must be used within the context that they are a correlative LOE. Additional information on the fundamental designs and analyses associated with in situ studies can be found in Underwood (1994), Quinn and Keough (2002), and Chariton et al. (2016).
3.6.2 Experimental In situ Studies
In order to validate correlative studies and to increase our understanding of how contaminants affect marine communities, the testing of models founded on cause and effect is essential. One way to do this is via in situ experiments. In marine community ecotoxicological studies, the two most common approaches are spike and translocation studies. Spiked studies are predominately sediment-based experiments that involve dosing a sediment with the contaminant of interest. Non-dosed (control) and dosed sediments are then transferred into containers and placed into the substrate unimpacted site(s) (e.g. Lu and Wu 2006; Birrer et al. 2018). The containers then remain in the sediment for sufficient time for them to be recolonized by the native biota, and comparisons between the compositions of the recolonized communities are used to determine whether the spiked sediment altered the species composition, and if so, at what concentrations effects were observed.
One of the challenges of spiked studies is ensuring that the contaminant remains bound to the sediment and does not alter the physicochemical properties of the sediments, minimising any differences between the controls and spiked sediments, apart from the toxicant of interest. Other additional limitations associated with this approach are that the endpoint is based on recolonized communities, and these may behave differently to established fauna; this approach is also limited to contaminants that bind to sediments, as hydrophilic contaminants will be released into the water column and be no longer present in the sediment.
Importantly, considerable resources are required to perform such studies, especially when you consider the quantity of sediment that may be required to fill, for example, 50 × 5-L containers. Furthermore, upscaling the approach to capture multiple contaminants at multiple concentrations is challenging. Imagine if you needed five containers of each treatment (replicates). If you just had a control and one contaminant at one concentration, 10 containers would be required. If you had one control and four concentrations (contaminant A, e.g. copper) this would require 25 containers (Figure 3.7a). If you added another stressor (contaminant B, e.g. endosulfan) even at one concentration, you would need 5 control containers, 20 containers of contaminant A, 5 containers of contaminant B by itself, and 20 containers to capture each concentration of contaminant A plus contaminant B. That’s 50 × 5-L containers (Figure 3.7b), and as sediment often weighs around 2 kg/L, that is roughly 500 kg of sediment which needs to be manipulated. As you can see, if you wanted multiple concentrations of contaminant B, the experiment would get rather big and complex, not only requiring a lot of sediment to be spiked, but also an extraordinary large amount of effort to install, recover, and process the samples. Consequently, spiked studies are generally restricted to a limited number of treatments.
Translocation studies are similar to spiked studies in that they involve placing containers of sediments into a reference site; however, in this case, the sediments are sourced from the locations sampled in the field study. The aim is not specifically to identify if a specific contaminant is causing an effect, but rather to test whether there is something about the sediments per se which is causing the effect. That is, translocation studies are designed to remove the effect of location by translocating all the sediments to a single location, enabling a direct comparison between sediments obtained from multiple locations. One of the challenges of this process is keeping the sediments intact, including their contaminants, while simultaneously removing the biota. This can be done either by freezing or anoxia (Chariton et al. 2011; O’Brien and Keough 2013) and is essential, given that recolonisation is the endpoint and all sediments must start with the same de-faunated state (i.e. no organisms).
While by no means routine, both spike and translocation studies can provide an additional line of community-level information to complement in situ field studies and laboratory-based toxicity tests. Spiking experiments aim to provide experimental evidence of whether a specific contaminant has the capacity to alter community composition, as well as some insight about at what concentration this may occur. Translocation studies, on the other hand, remove the potential confounding influence of location. Both approaches are resource-intensive and thereby place constraints on the experimental design, often limiting their statistical power (Chariton et al. 2011).
3.6.3 Laboratory Studies
While in situ experiments can provide community-level responses under environmentally relevant conditions, they are not without limitations. Most notably, they are very much sediment focussed and not easily amendable to hydrophilic chemicals, chemicals with short-half-lives (e.g. some herbicides), or for exploring the toxicity of contaminants within the water column. In such cases, it may be more appropriate to expose whole communities to a contaminant of interest under laboratory conditions (e.g. in replicated aquaria). Under such conditions, the physicochemical properties of the water column as well as the concentration of the stressor can be controlled. Furthermore, the overlying waters can be continually renewed, ensuring that metabolic waste such as ammonia is removed and not impairing the health of the exposed communities.
The power of laboratory community assays was demonstrated by Gissi et al. (2019) who exposed the coral Acropora muricata to a range of dissolved nickel and copper doses. While the coral itself is a species, each fragment contains its own microbial assemblages. Consequently, in this study, the authors were able to gain experimental laboratory species-level data from the host (A. muricata) as well as community-level information by examining the host’s external microbiomes. In the case of Gissi et al. (2019), the corals were wild-caught and allowed to acclimate for many weeks prior to exposure to the copper and nickel. While this helps ensure that the microbial communities are similar across all individuals, it does not infer that the microbial communities are the same as those which naturally reside on the corals at the site of collection, with other studies showing that marked differences in community structure can occur when transferring communities from the field to the laboratory (Ho et al. 2013; Chariton et al. 2014).
In a novel study by Ho et al. (2013), the authors examined the effects of the antibacterial agent triclosan on marine meiobenthic and macrobenthic communities. Their approach was to collect whole communities, including their residing sediment, and allow the whole sediment communities to acclimatise within a facility under a continual flow system that also supplied food. Instead of dosing the sediments, the authors placed a layer of the toxicant in a slurry on top of the community, and then 2 weeks later applied a clean sediment on top of this. The authors hypothesised that those animals which were alive would migrate through the clean sediment enabling them to be sampled, thereby providing a different community-level endpoint to the recolonized fauna obtained in in situ spiked and translocation experiments.
As in the case of in situ field experiments, laboratory-based community experiments can be logistically challenging and require significant resources and expertise. As a rule of thumb, replication is generally kept to a minimum, and designs incorporating the interactions between multiple stressors are generally avoided.
The data generated by both in situ field and laboratory-based community experiments can be used in SSDs to derive limits. The data could be used by itself or by combining it with more traditional single-species laboratory-based data (e.g. Leung et al. 2005).
3.7 Summary
Obtaining a comprehensive understanding of the effects of contaminants on marine organisms and communities requires a combination of both laboratory and field studies. Ecotoxicology is one of the LOEs that can be used in risk assessment. Experiments are most often conducted in laboratories under controlled conditions. As with all studies, ecotoxicology requires strict adherence to protocols including replication, quality control, statistical analyses, and interpretation of results. Physicochemical conditions need to be standardised throughout experiments and measured to ensure experimental conditions are suitable for organism survival. While this information is limited by its lack of relevance to changing conditions in the environment, it provides some important criteria for comparative assessment and is used in a multiple LOEs approach to develop guideline values for water and sediment quality. Studies of the effects of contaminants on community structure and function are more often based in situ. Due to the complex and dynamic nature of marine communities, thought must be given to the many variables that will influence toxicity.
3.8 Study Questions and Activities
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1.
What are the benefits of using sub-lethal endpoints to assess toxicity, as opposed to lethal endpoints?
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2.
What considerations must be given to chronic toxicity test procedures?
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3.
Describe what a species sensitivity distribution curve is and how it is used.
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4.
Explain why range finder experiments are used in ecotoxicology.
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5.
Design a laboratory toxicity test to assess the effects of the pesticide imidacloprid on a marine species. Consider the species of interest to you, the experimental design, the duration of the exposure, what endpoint you will use, the concentration range you will use, how and when you will measure the test conditions (including the imidacloprid concentrations), and how you will interpret the results.
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6.
What are the advantages and disadvantages of in situ experiments?
Abbreviations
- ANOVA:
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Analysis of variance
- AF:
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Assessment factor(s)
- EC50:
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Concentration of a toxicant that causes a measured negative effect to 50% of a test population
- EC10:
-
Concentration of a toxicant that causes a measured negative effect to 10% of a test population
- EDA:
-
Effects-directed analysis
- HC1:
-
Harmful concentration for 1% of species. Equivalent to the PC99
- HC5:
-
Harmful concentration for 5% of species. Equivalent to the PC95
- LC50:
-
Concentration of a toxicant that causes a 50% mortality to a test population
- LOE:
-
Line of evidence
- NOEC:
-
No observed effect concentration
- PC99:
-
The protective concentration for 99% of species
- PC95:
-
The protective concentration for 95% of species
- POPs:
-
Persistent organic pollutants
- QSAR:
-
Quantitative structure–activity relationship
- SF:
-
Safety factor(s)
- SSD:
-
Species sensitivity distribution
- TIE:
-
Toxicity identification evaluation
- US EPA:
-
United States Environmental Protection Agency
- WET:
-
Whole effluent toxicity test
- WOE:
-
Weight of evidence
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Reichelt-Brushett, A., Howe, P.L., Chariton, A.A., Warne, M.S.J. (2023). Assessing Organism and Community Responses. In: Reichelt-Brushett, A. (eds) Marine Pollution – Monitoring, Management and Mitigation . Springer Textbooks in Earth Sciences, Geography and Environment. Springer, Cham. https://doi.org/10.1007/978-3-031-10127-4_3
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