Trace elements in the environment

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Trace elements speciation

As already mentioned, trace elements can have variable toxicity which often depends on their chemical forms. As a consequence, the bulk concentration of the element is often not sufficient information for the assessment of its real toxicity (Chai et al., 2004). Consequently, in the last decades, researches have been focusing on the analysis of chemical species of several TE to evaluate their impact on the environment (Sánchez Uría, 1998; Leermakers et al., 2005; Semeniuk et al., 2016). Among toxic elements, arsenic and mercury are two of the most studied elements regarding chemical speciation in environmental samples.

Arsenic

Arsenic (As) is classified as metalloid as it presents characteristics of both, metallic and non-metallic element (Phillips, 1990). It is associated with igneous and sedimentary rocks, particularly with sulfidic ores as it is strongly chalcophile. Several anthropogenic activities such as metals and alloys manufacturing, petroleum refining and burning of wastes and fossil fuels, have contributed to a strong legacy of As pollution throughout the world (Melamed, 2005). Several As-containing compounds are produced by industry and have been used to manufacture products for agricultural applications such as insecticides, herbicides, and fungicides. Numerous species of As can occur in the environment and the interconversion of species from toxic to less harmful and vice versa is possible as a consequence of some physical-chemical changes such as biological activity, changes in redox potential, or pH. To determine the actual risk represented by this element in the environment, a proper study should include identifying and quantifying both the total quantity of arsenic and the specific chemical forms present in the sample (Gong, 2002).
In the environment, arsenic occurs usually under inorganic species, and rarely under organic ones (Figure 4). The main inorganic species are found under two oxidation states: arsenite As(+III) and arsenate As(+V), as the oxyacids (H3AsO3 and H3AsO4). In many environments (under neutral conditions), As(+V) is often deprotonated as an arsenate oxyanion (H2AsO4- or HAsO42-); in contrast, As(+III) remains predominantly in its neutral form as H3AsO3. Arsenate oxyanions and the neutral arsenite constitute the main targets for field analytical assays (Melamed, 2005). Inorganic As species are found in water and soils and in small quantities in living organisms where rather organoarsenic species dominate. Some biological processes (mainly due to bacterial activities) can also promote the oxidation or the reduction of inorganic As (Anderson et al., 1992; Santini et al., 2004; Silver and Phung, 2005; van den Hoven and Santini, 2004). Usually freshwaters and ground waters scarcely contain organic As species such as methylarsonate (MMA, CH3AsO3H2) and dimethylarsinate (DMA, (CH3)2AsO2H) (Bednar et al., 2004; Bohari et al., 2001; Martinez-Bravo et al., 2001). Organic As species (methylated species and arsenosugars) usually occur in biological organisms (particularly in marine ecosystems). Arsenobetaine (AsB) is usually the dominant As species occurring in fishes and seafood, but other methylated compounds are also found in marine biota (Nam et al., 2010; Grotti et al., 2010). Biological processes promoted by some microorganisms or bacteria enhance the methylation of inorganic species in organic ones, such as MMA, DMA, trimethylarsine [TMA, (CH3)3As] (Bentley and Chasteen, 2002; Cullen and Reimer, 1989; Irgolic and Stockton, 1987). Arsenic is essential for some organisms, e.g., 12 to 25 μg of total As per day is required by humans, but it is also toxic at low and high concentration (WHO, 2006). Chronic As exposure increases the risk of cancer and skin pigmentation. Arsenic has great notoriety as a poison, though there are great differences in the toxicity of different arsenic compounds. Arsenic occurs as organic and inorganic species, and it can have different oxidation states (-III, 0, +III and +V). Its toxicity also depends on its valence and speciation, inorganic species being by far the most toxic species (Table 1).
Most cases of arsenic human toxicity have been associated with exposure to inorganic arsenic. Inorganic trivalent arsenite As(+III) is 2–10 times more toxic than inorganic pentavalent arsenate As(+V) (Goyer, 2001). By binding to thiol or sulfhydryl groups on proteins, As(+III) can inactivate over 200 enzymes. This mechanisms is the likely mechanism responsible for arsenic’s widespread effects on different organ systems (Tchounwou et al., 2012). The acute toxicity of As(+V) is related to its ability to replace phosphate, which is involved in many biochemical pathways (Goyer, 2001; Hughes, 2002). Arsenic toxicity is much higher in inorganic forms than organic species of arsenic (Table 1). The major metabolic pathway for inorganic arsenic in humans is methylation. Most of the mammalian species methylate inorganic arsenic (Vahter, 1994). Inorganic arsenic is metabolized by a sequential process involving a two-electron reduction of inorganic pentavalent arsenic to inorganic trivalent arsenic, followed by oxidative methylation to pentavalent organic arsenic (Thomas et al., 2001). The reduction can occur non-enzymatically in the presence of a thiol such as glutathione (GSH) (Delnomdedieu et al., 1994; Scott et al., 1993). However in human livers, some specific enzymes have been isolated, suggesting that the reduction occurs using enzymatic reaction (Radabaugh and Aposhian, 2000; Zakharyan et al., 2001). The species produced as a result of these reactions are then expelled along with urine.

Mercury

Mercury (Hg) is non-essential and highly toxic trace element. This metal has a very different chemistry from its group and period neighbors, and it is the only metal liquid at room temperature. Mercury naturally occurs in the environment although it is considered rather rare chalcophile element. The principal mineral is cinnabar (HgS), along with the metacinnabar group of minerals; other primary Hg minerals include native mercury, corderoite Hg3S2Cl2 and livingstonite HgSb4S8. Most mercury is derived by natural sources but its biogeochemical cycles is disrupted by anthropic inputs as several human activities have contribute to increase its environmental input since the industrial revolution (Wu et al., 2006; Chen et al., 2013; Lamborg et al., 2014). Anthropogenic sources of mercury come from metal production, chlor-alkali, and pulp industries, waste handling and treatment, and coal, peat, and wood burning (Lindqvist et al., 1991).
Mercury occurs in the environment under three principal forms: Hg(0), inorganic Hg and organic forms such as methyl mercury (MeHg). Hg coming from natural and anthropogenic sources is principally emitted in the atmosphere, where its global cycle (Figure 6) is largely controlled by oxidation–reduction reactions. For instance, it can easily oxidize into soluble (inorganic) Hg2+ which in turn can deposit on water and soil surfaces. A small fraction of mercury in natural waters is converted to organic forms such as methylmercury (CH3Hg+) and dimethylmercury [(CH3)2Hg], by bacterial action in anoxic environments.
Metallic Hg is relatively inert and not readily taken up by organisms, but it is volatile and its vapor is toxic. Inorganic forms such as HgCl2 are also considered toxic especially for long-term exposure (Lohren et al., 2015). Fat-soluble organic forms (especially MeHg) accumulate in the biota via the food chain (up to several mg kg-1), resulting in human exposure through consumption of fish (Morel et al., 1998). This compound is considered the most toxic form of mercury since it causes serious central nervous system dysfunctions (Harada, 1995). Additionally, contrary to other trace elements efficiently accumulated by organisms, Hg is bio-magnified through the food chain, meaning that the highest concentrations of this element are found in the highest levels of the food chain. Hg(0) and Me2Hg are not bioaccumulated by organisms because they are not reactive and thus not retained by plankton in the first place (Morel et al., 1998). On the other hand, Hg(II) is efficiently retained by microorganisms but not efficiently transferred to other organisms through food chain in comparison to MeHg (Mason et al., 1996) as it probably happens when considering other trace elements. In 2013, a multilateral environmental agreement known as Minamata convention, has been promoted by UNEP and it addresses the adverse effects of mercury through practical actions to protect human health and the environment from anthropogenic emissions and releases of mercury and its compounds (Briant et al., 2017).

Trace elements isotopes

Many elements exist in nature with different isotopes, i.e., atoms of the same element which differ from each other only in their number of neutrons. Generally speaking, it can be stated that all elements have an isotopic composition stable in nature. However, there are some reasons why this isotopic composition may show some variation. One of this reason is the presence, among the isotopes, of radiogenic nuclides. Such nuclides are produced as a result of radioactive decay. The additional production of this radiogenic isotope can have a pronounced effect on the final isotopic composition of a given element. Other phenomena responsible for a change in isotopic composition are interaction with cosmic rays, mass-dependent and mass-independent mass fractionation and the presence of extraterrestrial material (Vanhaecke and Degryse, 2012).
As already stated, some elements show natural variation in their isotopic composition as a consequence of the presence, among its isotopes, of radiogenic nuclides (Ault et al., 1970). A radiogenic nuclide is continuously produced as long as the parent radionuclide is available, leading to an increase of its relative abundance with respect to other isotopes of the same element. The presence of one radiogenic isotope within the same element affects the general isotopic composition and the relative abundance of other isotopes. Mass fractionation is related to the fact that different isotopes can participate with different efficiency to chemical and physical processes. These differences in efficiency are related to slight differences in equilibrium for each different isotopic molecule (thermodynamic effect) or in the rate with which the isotopes participate in a process or reaction (kinetic effect) (Vanhaecke and Degryse, 2012). As a result, the isotopic composition of such element will be different at the end of the process in comparison to the isotopic composition at the beginning of it. These effects need to be considered when measurements of isotopic ratios are performed. Most of the time the phenomenon of mass fractionation shows linear dependency with the mass of the isotope: but for some elements it constitutes a hot topic of research (Vanhaecke and Degryse, 2012).
When measuring isotope ratios by ICP-MS, the effect known as mass discrimination has to be taken into consideration. As a result of differences in the efficiency of ion extraction, transmission and detection as a function of analyte mass, an isotope ratio measured with ICP-MS may show significant bias with respect to the corresponding true value (Vanhaecke et al., 2009). This effect is constant and needs to be corrected during the measurement session. For this purpose, several correction models have been developed, often involving a mathematical correction based on the measurement of a certified external isotopic standard (e.g. NIST 981 or NIST 982). Another method involves the use of an internal standard, with similar mass to the analyte, directly added in the sample to be analyzed. For example, in Pb isotopic analysis, a Tl standard is added to the sample solution and the bias between the measured 203Tl/205Tl and the corresponding true value is used to determine the mass discrimination per mass unit, which is in turn used to correct the Pb ratio data (Vanhaecke et al., 2009).
The measurement of isotopic ratios finds application for numerous elements, especially in provenance studies. Sr, Pb, Nd and U are only few examples of elements used for this application, measurable by ICP-MS (Debord et al., 2017; Kingson et al., 2017; Vio et al., 2018).

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Pb Isotopic ratio

Lead is perhaps the most popular among elements partly radiogenic accessible via ICP-MS. It has four stable isotopes, three of which are radiogenic. The decay chain of 238U (t1/2=4.468 x 109 years), 235U (t1/2=0.407 x 109 years) and 232Th (t1/2=14.010 x 109 years) finally results in 206Pb, 207Pb and 208Pb respectively. 204Pb is the only natural isotope, and its abundance is constant since the formation of the solar system (Komárek et al., 2008). As a consequence, the final Pb isotopic composition depends on the original U/Pb and U/Th elemental ratios, becoming characteristic of a particular geographical location. For this reason, Pb isotope ratios find application in provenance studies in several fields (Sjåstad et al., 2011; Nakata et al., 2015; Dudás et al., 2016; Shepherd et al., 2016; Sjåstad et al., 2016). In environmental science, the use of Pb isotope ratios is a well-known tool, helping scientist to determine the actual presence of Pb pollution and to track the source of anthropogenic inputs. There is a substantial difference between isotopic composition of crustal Pb and that in ores. Pb isotope ratios analysis provides thus an excellent tool to distinguish between local Pb and Pb pollution derived from ores (e.g., anti-knock compounds added to petrol or for industrial use) (Monna et al., 1995; Bollhöfer and Rosman, 2000; Bollhöfer and Rosman, 2001). In environmental samples such as aerosol, sediments and snow, the final isotopic composition is the result of natural and one or more anthropogenic inputs. A way to verify such a possibility is to plot the results in the so called three isotope graph (Figure 7), which is a graphical representation of two different ratios with a common term e.g 208Pb/204Pb vs 206Pb/204Pb.
If the results will fall in a mixing line (linear trend) in this diagram, it means that Pb origins from two sources, which are the two end members of the mentioned line. With this representation, the contribution of the 2 sources can be calculated (Komárek et al., 2008). If three end members contribute to the final isotopic composition, the results in the three isotopes graph will plot in a triangular field, delimited by lines connecting the ratios of the three end-members. In such and more complex case, the determination of the single contribution is rather complicated (Vanhaecke and Degryse, 2012). A popular plot in environmental studies is the 207Pb/206Pb versus 208Pb/206Pb, which often shows a linear trend that can be interpreted as a simple mixture of two Pb sources. Ellam (2010) recently demonstrated that this type of representation could lead to the erroneous conclusion since the linear trend in such representation is inevitable and thus it is not suitable in the case of multiple source mixing. A more reliable test for multiple mixing models can be obtained by plots involving 206Pb/204Pb, 207Pb/204Pb, and 208Pb/204Pb, which highlights the importance of 204Pb isotope. The limitation of this approach arises from the difficult determination of 204Pb using ICP-MS, as it is the least abundant isotope (about 1.4% natural abundance) and it is affected by 204Hg isobaric interference.

TRACE ELEMENTS MONITORING PROGRAMS

Environmental monitoring describes the processes and activities that are put in place in order to assess the quality of the environment. All monitoring strategies and programs have reasons and justifications which are often designed to establish the current status of an environment or to establish trends in environmental parameters. The ecological, economic and social importance of marine ecosystems being undeniable, a well-planned approach of managing the marine space is essential to achieve sustainability (Jonathan and Gobert, 2016). In parallel with growing concern for environmental conditions, several ocean monitoring programs have been introduced, at national and international levels (Kennish, 1994). Environmental assessment programs have a crucial role in the management of pollution sources and the preservation of marine habitats. The piece of coasts to be managed may be under actual pressure, such as eutrophication and pollution, or it may be under threat of pressure from some proposed development. Historically, most of the water quality investigations have attempted to assess trace elements in aquatic systems by direct analysis of water samples. Many studies have dealt with the determination of total and dissolved TE concentrations through the analysis of filtered and non-filtered water (Cidu and Frau, 2009; Bu et al., 2017). Concentration associated with suspended particulate matters (SPM) is usually indirectly determined as the difference between results obtained by non-filtered water and filtered one. In most of aquatic systems, TE concentrations in SPM are much higher (several order of magnitude) than those in the water samples. Some of the most toxic trace elements (e.g., Hg, As, Cd, Pb and Zn) are strongly associated with sediments and plankton which means that their distribution, mobility and bioavailability in the aquatic ecosystem cannot be solely evaluated by their determination in water (Horowitz, 1991). In Europe, the recent Environmental Quality Standard Directive 2008/105/EC (EQSD, 2008) marked an important step in the use of sediments and biota in environmental monitoring programs, in the frame of Water Framework Directive 2000/60/EC (WFD). Analysis of TE in sediments and biota is widely recognized as an effective approach in water-quality monitoring to provide baseline regional values for local monitoring, for describing the actual contamination levels as well as to identify areas of particular concern (Carere et al., 2012). Most of the toxic substances listed in the principal environmental ranking, accumulate in both sediments and biota.
Is therefore of crucial importance to include sediments, suspended or bed sediments, and biota in a proper environmental monitoring program.

Trace elements in sediments

Since sediments can act as reservoirs for many trace elements, they are often included in the planning and design of environmental monitoring (Carere et al., 2012; Quevauviller, 2016). One of the great application of sediment analysis in monitoring programs is related to the capability of this final sink to accumulate trace elements and thus providing an historical record of contamination; this property being often observed in sediment cores (Kljaković-Gašpić et al., 2008; Karbassi et al., 2005; Sun et al., 2012; Vallius, 2014). Additionally, under changing environmental and physiochemical conditions (e.g. pH, dissolved oxygen, bacterial actions), TE bound to sediments can be released in the water column, increasing their harmful effect in the environment and food chain (Le Pape et al., 2012; Tripti et al., 2013; Joung and Shiller, 2016).

Sediment components and classification

Marine sediments have very heterogeneous composition, depending on their origin and their geographical localization. The accumulation of TE in sediments depends directly on the nature of sediments. Three main types of sediments can be distinguished based on sources and components (Fütterer et al., 2006): (1) lithogenous (terrigenous or volcanogenic) sediments, transported and dissolved into the oceans as detrital particles; (2) biogenous sediments, formed directly by accumulation of skeletal parts of marine organisms; and (3) hydrogenous sediments, which precipitate directly out of solution as new formations.

Table of contents :

PART I: GENERAL INTRODUCTION
1. Trace elements in the environment
1.1. Trace elements in the aquatic system
1.2. Trace elements speciation
1.2.1. Arsenic
1.2.2. Mercury
1.3. Trace elements isotopes
1.3.1. Pb Isotopic ratios
2. Trace elements monitoring programs
2.1. Trace elements in sediments
2.1.1. Sediment components and classification
2.1.2. Factors affecting TE accumulation in sediments
2.1.3. TE monitoring in sediments
2.2. Trace elements in marine organisms
2.2.1. Bivalves
2.2.2. Sponges
3. Trace elements analysis
3.1. Atomic Absorption Spectrometry
3PART I: GENERAL INTRODUCTION
1. Trace elements in the environment
1.1. Trace elements in the aquatic system
1.2. Trace elements speciation
1.2.1. Arsenic
1.2.2. Mercury
1.3. Trace elements isotopes
1.3.1. Pb Isotopic ratios
2. Trace elements monitoring programs
2.1. Trace elements in sediments
2.1.1. Sediment components and classification
2.1.2. Factors affecting TE accumulation in sediments
2.1.3. TE monitoring in sediments
2.2. Trace elements in marine organisms
2.2.1. Bivalves
2.2.2. Sponges
3. Trace elements analysis
3.1. Atomic Absorption Spectrometry
3.2. Inductively Coupled Plasma Mass Spectrometry

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