Mercury, Hg

Abstract

In nature, mercury (Hg) occurs in the elemental form (Hg⁰), as well as in inorganic (InHg) and organic (OrgHg) compounds. It is the only heavy metal that is liquid at room temperature and easily turns into a gas. Mercury vapours can be transported with air masses for hundreds and thousands of kilometres and—after falling down—contribute to the pollution of land and waters. In aquatic environments biogeochemical processes promote the natural microbial conversion of InHg to methylmercury (MeHg), the most bioavailable form of Hg.

Human activities have increased atmospheric Hg concentrations 3–5 times over the past 150 years, mainly as a result of the combustion of fossil fuels. It is believed that all forms of Hg are toxic to endothermic animals and humans, but MeHg is particularly dangerous because of its neurotoxic and teratogenic effects as well as negative influence on reproduction. Moreover, in nature MeHg is biomagnified, and its concentration reaches the highest levels in top predators, especially piscivorous species. For several decades, there have also been reports documenting the local occurrence of dangerously high concentrations of Hg in organisms living in terrestrial ecosystems (including spiders, insects and songbirds feeding on them) in areas, which had been subject to anthropogenic Hg pollution many decades ago. Studies on inland aquatic and terrestrial ecosystems have indicated the long-term persistence of Hg introduced into the environment and the complexity of its transformations and circulation in nature. A better understanding of these processes requires further research, including the issue of bioaccumulation and biomagnification of MeHg in various ecosystems.

1 Introduction

Mercury causes many environmental and health problems. Together with lead and cadmium, it belongs to the group of particularly toxic metals, which do not have any physiological functions in warm-blooded vertebrates (including humans), and therefore even small amounts of absorbed mercury result in the disruption of biochemical processes in the body. Its elevated concentrations in birds and mammals lead to the development of many diseases (mainly in the nervous and excretory systems) and death (Clarkson and Magos 2006).

2 General Properties

Mercury (Hg from “hydrargyrum”, i.e. “liquid silver” from Greek “hydr-” for water and “argyros” for silver) is a heavy metal with a density of 13.55 g cm^–3. It is the only metal which occurs in a liquid form at room temperature; its freezing point is –38.83°C and boiling point is 356.73°C. It has good electrical conductivity and high volatility, reaching a vapour pressure of 1.22 × 10^–3 mm at 20°C (2.8 × 10^–3 mm at 30°C). Its solubility in water is 6 × 10^–6 g^–1 100 ml (25°C). In the atomic table of elements, mercury is located in the group IIB, with atomic number 80 and atomic mass 200.59. There are 33 known isotopes of Hg, of which 7 are stable. The general pool of Hg is dominated by three isotopes: ¹⁹⁹Hg, ²⁰⁰Hg and ²⁰²Hg. at 16.9%, 23.1% and 29.7%, respectively (Blum 2011). In the environment, mercury exists in the elemental form (Hg⁰) and in compounds with I (mercurous or Hg⁺) and II oxidation states (mercuric or Hg²⁺). Elemental mercury is an extremely good “solvent” for gold, silver and many other metals (except iron) via the formation of amalgams (alloys). It forms both inorganic and organic compounds, with the latter known as organomercurials. Common mercury salts contain halides (fluorine, chlorine, bromine and iodine) and sulphur (HgS). Organic compounds occur as R₂Hg and RHgX, where R represents a simple alkyl group such as methyl (CH₃ ^–) and X represents atoms or groups such as chlorine, bromine, iodine, cyanide and hydroxyl. Two of the organic compounds are monomethyl mercury CH₃HgX (methylmercury, MeHg) and dimethyl mercury (CH₃)₂Hg, the most important chemical forms of Hg with respect to environmental impact assessments (National Research Council 2000; Scoullos et al. 2001).

3 Mercury in Nature

Mercury is a natural component of the Earth’s crust, occurring in soil, water and air where it penetrates into living organisms. In the environment it is found in an elemental form or in inorganic and organic compounds with varying degrees of toxicity to plants and animals, including vertebrates.

3.1 Mercury in the Abiotic Environment

It is estimated that Hg constitutes only 0.083 × 10^–4% of the Earth’s crust and is in the 63rd position in terms of percentage share in the lithosphere. Mercury is present in the Earth’s upper crust at a mean concentration of ~0.05 mg kg^–1 (ppm). Its abundance in igneous rocks is lower than in sedimentary rocks (0.004–0.08 and 0.01–0.40 ppm, respectively) and is mainly concentrated in argillaceous sediments. As a chalcophile element, this metal exhibits high affinity for sulphur and low to oxygen. Mercury occurs mainly in minerals containing sulphides and sulpho-salts and accompanies the ores of many metals (including copper, silver, zinc and lead). Generally, Hg is considered to be a rare element and extensively dispersed in the lithosphere (Yaroshevsky 2006; Kabata-Pendias and Mukherjee 2007; Kabata-Pendias 2011).

About 90 Hg minerals have been described, including cinnabar (HgS) and calomel (Hg₂Cl₂). Various ores generally contain from 0.1 to 2.5% Hg and occasionally >7% Hg. In some parts of the world there are geological anomalies with very high accumulations of minerals rich in Hg. Geologists have described more than 2200 sites where ores not only contain significant amounts of mercury but where also the soil, deposits of coal and oil and inland waters are characterized by elevated Hg content. Most of these sites are located within three transcontinental belts, usually with significant volcanic activity. The first belt (Mediterraneo-Himalayan) runs from the Iberian Peninsula in Europe to the Himalayas in Asia, the second covers the area lying along the west coast of the Pacific, and the third runs through the western areas of the Americas, together with the Pacific Ocean adjoining them; therefore, the Pacific is surrounded by the zone naturally high in mercury (Rytuba 2003; AMAP/UNEP 2013). Ores containing cinnabar, the most widespread of natural mercury-containing minerals, are present in approximately 60 countries. Five of the richest deposits of Hg include three European sites (Almaden in Spain, Monte Amiata in Central Italy, Idrija in Slovenia) and one located in North America (including New Almaden and New Idria in California, USA) and in South America (Huancavelica in Peru). These deposits were exploited for hundreds of years but eventually were closed in the period 1982–2002 (Ferrara et al. 1999; Gnamuš and Horvat 1999; US GS 2016a). In addition to those already mentioned, areas particularly rich in mercury can be found in China and Kyrgyzstan (Scoullos et al. 2001; Hylander and Meili 2003; Gómez et al. 2007).

The concentration of Hg in environmental samples is generally low outside of these geological anomalies and areas anthropogenically contaminated by this element. Hg levels in the air in Greenland range between 0.01 and 0.06 ng m^–3 and in snow and rainwater do not normally exceed 0.2 μg L^–1. In inland surface waters, the concentration of Hg ranges from 0.2 to 1.0 μg L^–1, and it is typically lower in rivers than in lakes (Adriano 2001). Globally, the average concentration of mercury in soils assumes is about 0.16 mg kg^–1 dry weight, (range 0.06–0.20 mg kg^–1 dw), but in European agricultural soils, it is markedly less and does not exceed 0.04 mg kg^–1 dw (Adriano 2001; De Vos et al. 2006). Much higher values are listed in soils of volcanic origin, where the concentration of Hg can exceed 7 mg kg^–1 (Kabata-Pendias 2011). From the environmental and economic points of view, the mercury content in mined and processed raw materials is most essential. These are mainly ores of mercury and other metals, which are accompanied by mercury, rocks used in the cement industry and fossil fuels (Table 17.1). To obtain mercury on an industrial scale, ore with an average content of 0.6–3.2% is exploited, while there are also deposits in Almaden (Spain) which comprise 8% Hg or 80,000 ppm. In addition, some local rocks there contain small drops of native mercury (Kim et al. 2004; Gómez et al. 2007). Most mercury mines in the world have been closed, with those remaining open are located mainly in Asia.

Table 17.1 Mercury concentrations in raw materials for industry (dw, dry weight)

3.2 Mercury Production and Uses

Due to its unique properties, mercury and its compounds have been used in a variety of applications since ancient times. Over the centuries, cinnabar (vermilion) with a characteristic vivid red colour was widely used as a pigment in art, wall decorations, cosmetics and some medicines in Rome, mediaeval Europe, Egypt, India and China. Even in the twenty-first century, it is used in some ritualistic and spiritual practices. Up to now mercury was extracted in poor countries by heating cinnabar in a current of air and condensing the vapour. By 500 BC, mercury was used to make amalgams with other metals. This property of mercury to form alloys is still widely used, particularly in obtaining precious metals and the preparation of dental amalgams (“silver fillings”). Such cheap and permanent fillings have been used in dentistry since the nineteenth century. Since the last century, mercury has been used on a large scale in the chemical and electrochemical industries for electrical and electronic applications (among others in switches, batteries, fluorescent lamps and energy-saving light bulbs). It is also found in some control devices (thermometers, barometers and manometers) and some pesticides, although developed countries significantly reduced the use of mercury in various products and processes due to its high toxicity and environmental hazard (Caley 1928; Parsons and Percival 2005; Masur 2011; Teaf and Garber 2012).

The world’s richest source of cinnabar and quicksilver in Almaden (Spain) was operated for over 2000 years, with about 7 million tons of Hg extracted (Tejero et al. 2015). For comparison, from 1500 to 2000, the entire world production of Hg was less than 1 million tons, of which Almaden accounted for ~33% (Gómez et al. 2007; Hylander and Meili 2003; Tejero et al. 2015). As late as in 1971–1980, world production of Hg was very large, with an estimated production of 81,925 tons Hg, of which the former Soviet Union (including Ukraine, Russia, Kyrgyzstan and Tajikistan) accounted for 26%, Spain (Almaden) for 18% and the United States (California and Nevada) for 10.2% (Hylander and Meili 2003). Since then, the global excavation of Hg has dropped more than five times, and in the decade from 2001 to 2010, it amounted to a total of 16,310 tons (US GS 2001–2011). Table 17.2 shows the three countries with the highest production of Hg in the period 1980–2015.

Table 17.2 Three countries with the highest mercury production in selected years (metric tons)

By the end of 1970, mining mercury in European mines in Almaden, Monte Amiata, Idrija and North America, and Hg use in various sectors of the economy in those parts of the world, was significantly higher compared to developing countries in Europe and Asia. Over time, it changed significantly, and since 2005 China has been the world leader in extraction (Table 17.2). In 2015, 1600 tons Hg was excavated in China, accounting for nearly 70% of global production (US GS 2016b). It is estimated that 80% of the world’s mercury reserves have already been processed through human products (Meinert et al. 2016).

Before 1980 metallic mercury had been used in significant quantities, mainly for the extraction of gold and silver (for centuries), in dental amalgam fillings, as a catalyst in the chlor-alkali industry (where liquid Hg is the cathode, and this is one of three chlorine production technologies) and production of vinyl chloride monomer (VCM) used to synthesize polyvinyl chloride, PVC, to produce tubes, bottles, window frames and many articles. Moreover, metallic Hg is used in measuring devices, in electrical and electronic switches as well as in fluorescent lamps. Inorganic mercury compounds were used, among others, in Hg-oxide batteries, as pigments and dyes and as antiseptics in pharmaceuticals, while organic compounds of Hg (including alkyl forms) were used mainly as effective biocides in the paper industry and were added as an antifouling agent to paints and as fungicide to protect seeds and plants from fungal diseases (Hylander and Meili 2003). Due to the strong toxicity of mercury, already well documented in medical and ecotoxicological studies from 1950 to 1980, and focusing on the protection of health and care for the quality of the environment, regulations limiting economic exploitation, mining and trade of mercury have been gradually introduced in the European Union (EU) and North America. In those parts of the world, mercury mines had been shut before 2002. The consequence of the aforementioned actions was a drastic reduction in demand for mercury and a drop in its prices (Hylander and Meili 2003; Parsons and Percival 2005; Mohapatra and Mitchell 2009; UNEP 2013).

World mercury mining in 1980 was still relatively high at 6811 tons, but in 2005 it fell to 1520 tons. At that time, production and consumption of mercury shifted significantly from Europe and North America to Asia (US GS 1981, 2006). In 2005, including in Asia, Europe and North America, various sectors of the economy consumed 3188 tons Hg, of which Asia accounted for almost 67%, Europe (EU25 + CIS and other European countries) for 22.5% and North America 10.8% (AMAP/UNEP 2008). In Asia, most mercury is used in VCM and battery production (750 and 280 tons, respectively), in EU25 in mercury-cell chlor-alkali production (175 tons) and dental amalgam production (95 tons) and in North America in mercury-cell chlor-alkali production (60 tons) and production of measuring and control devices (48 tons). Several years later (in 2011), the global demand for mercury had dropped to 1930 tons, and the dominant recipient of this metal was chemical manufacturing (including 15% of the chlor-alkali industry and 21% of vinyl chloride monomer production) and artisanal and small-scale gold mining ASGM (24%) and batteries (13%), and further positions were dental amalgams 8%, measuring and control devices 7%, electrical and electronic devices 7% and fluorescent lighting 4% (UNEP 2013). According to a report by the United Nations Environmental Programme (UNEP) Global Mercury Partnership and its mercury-cell chlor-alkali production partnership area, this industry saw a very noticeable reduction in global demand for mercury. Between the base year 2005 and 2015, the consumption of mercury in the chlor-alkali industry fell by 50%, from 500 to 250 tons, resulting from the reduction in the number of plants that uses mercury in the production of chlorine and alkalis, through their closure or a shift into mercury-free technology (UNEP 2016), especially in this regard in the EU, where the use of mercury in chlor-alkali industry will have ceased in 2017 (Eurochlor 2016).

Although between 1980 and 2007 the global demand for mercury fell dramatically, and its production decreased almost six times (from 6811 to 1170 tons according to US GS in 1981, 2008), in recent years this downward trend has unfortunately changed, caused by the global economic crisis in 2008. For comparison, in 2008 and 2015, the global production of mercury was, respectively, 1320 and 2340 tons, significantly higher than in 2007 (US GS 2010, 2016b). The current increase in demand for mercury is significantly associated with an increased demand for gold, as its acquisition by the inexpensive method of amalgamation requires Hg. This method is mainly used in ASGM in developing countries (UNEP 2013).

3.2.1 Emission Sources of Mercury

Hg is released from natural (geogenic) and anthropogenic sources, including intentional (Hg acquisition from its ores, meeting the needs of certain sectors of the economy) and unintentional, that accompany various production and energy processes. Geogenic sources of mercury in nature include volcanic eruptions, weathering of rocks, natural forest fires and steppes and evaporation of the seas and oceans. Partially, these also include areas around active and abandoned Hg mines (with the deposited waste), often with significant levels of that element. Terrestrial sources and the oceans are credited with 48 and 52% of total annual emissions of mercury into the air. Researchers that from 80 to 600 tons of Hg reach from the land to the air, with the geogenic emissions mainly caused by mass burning (13%) and metal release from the desert, metalliferous and non-vegetated zones (10%), as well as some biomes such as tundra, grassland, savannah, prairie and chaparral (9%) (Pirrone et al. 2010; AMAP/UNEP 2013). In 2010 oceanic sources accounted for up to 2900 tons of Hg released into the global atmosphere, including the contribution from re-emission processes, which are emissions of previously deposited Hg originating from anthropogenic and natural sources, and primary emissions from natural reservoirs (AMAP/UNEP 2013).

Over the past few decades, the major sources of anthropogenic mercury unintentionally released into the air are the combustion of fossil fuels, mining and the processing of non-ferrous ores, cement production, natural gas cleaning, recycling and government stockpiles and incineration of sludge from biological treatment (Mohapatra and Mitchell 2009). Fossil fuels and various industrial raw materials usually contain small quantities of Hg (Table 17.1), but given the huge amounts used by man, their contribution to environmental pollution with Hg is a key position in its biogeochemical cycle. However, in 2010 it was recognized that global anthropogenic emissions of mercury to the air are mainly based on artisanal and small-scale gold mining (ASGM), before the process of burning coal for the needs of electro-energy (AMAP/UNEP 2013). It is estimated that in 2010, Hg from anthropogenic sources amounted to about 2000 tons, and another 1000 tons was released into waters, wherein the emission of water is much less recognized and evaluated in comparison to the atmospheric release. It is believed that chlor-alkali plants, paper pulp factories and mine wastes have been the major industrial sources that discharge mercury waste into water bodies (Mohapatra and Mitchell 2009; AMAP/UNEP 2013; UNEP 2013). In 2010, global atmospheric mercury emissions totalled 8900 tons, of which the current emissions from natural and anthropogenic sources account for 80–600 tons and about 2000 tons. The remaining amount of Hg (60%) in the annual amount came from re-emission, with the terrestrial and oceanic volumes estimated to be 1700–2800 and 2000–2950 tons, respectively (AMAP/UNEP 2013).

For about 200 years, we have seen a significant increase in the quantity of mercury circulating in nature. This is indicated by comparative studies of lake bottom sediments, peat deposits and core glaciers (Schuster et al. 2002; Allan et al. 2013). It is estimated that, compared to pre-industrial times, the concentration of Hg in the atmosphere and in the geochemical background has increased at least three times and probably 5–10 times in relation to the natural level (Mason et al. 2012; Horowitz et al. 2014). On a global scale, in the period 1850–2010, unintentional anthropogenic sources (from “by-product” sectors including fossil fuel combustion) issued to the atmosphere 215,000 tons of mercury. During that time, a further 540,000 tons of mercury was introduced into the environment from intentional commercial Hg uses and nonatmospheric releases from chlor-alkali plants and mining processes. From this very large pool, 20% reached the air, 30% waters, 30% soils and 20% landfill wastes. Some of this mercury remains in landfills or is associated with bottom sediments, but a significant quantity (310,000 tons) actively participates in the geochemical cycle (Horowitz et al. 2014).

Emissions of mercury into the environment have clearly differed between the Northern and Southern Hemispheres, where human economic activity releases 70% and 30% Hg, respectively (Pacyna et al. 2006; Selin et al. 2008; Pirrone et al. 2010; AMAP/UNEP 2013). This disparity in the emissions of Hg between the two hemispheres has historical, economic and demographic reasons.

Mercury released from natural and anthropogenic sources circulates in nature for a long time and is transmitted over long distances by strong atmospheric and ocean currents. Probably, it will take about a thousand years before mercury is released from stable formations in the lithosphere and circulating in the air-water-soil system, settles on the ocean floor and is permanently bound by mineral deposits in the rock formation processes (Mason et al. 2012; Horowitz et al. 2014).

Between 1980 and 2007, the mining of mercury decreased almost six times, which was driven by the results of numerous studies and regulations for the protection of health and the environment. Scientific studies provide ample evidence of the strong toxicity of Hg (especially MeHg) on humans and other warm-blooded vertebrates and document a dramatic increase in the amount of anthropogenic environment (Hylander and Meili 2003; Clarkson and Magos 2006; Horowitz et al. 2014). Out of many disasters caused by environmental Hg poisoning, the best known are the tragic events from the Japanese Minamata Bay from the 1950s, with the mass Hg poisoning of residents, their cats and wild birds, via the fish and seafood consumed. The primary source of mercury was wastewater from chemical plants discharging into the bay. The increasing awareness of risks arising from the increase in the amount of anthropogenic Hg in the environment has led to the introduction of regulations aimed at limiting the extraction, use and trade of Hg and consequently a reduction in the release of mercury into the air, water and soil from anthropogenic sources. Such pro-health and pro-environmental legislative action were taken earliest in the well-developed countries of the EU, North America and Japan, but globally more important will be the implementation of the provisions of the Minamata Convention, adopted on 10 October 2013 at a diplomatic conference held in Kumamoto, Japan. The convention entered into force on 16 August 2017 (www.mercuryconvention.org).

3.3 Biological Status of Mercury

According to current knowledge, mercury does not have any physiological function in eukaryotic and in most prokaryotic organisms. Its accumulation results in various life-threatening disorders and can lead to fatal poisoning (Clarkson 1992; Barkay and Wagner-Döbler 2005; Scheuhammer et al. 2015). Recently, Gregoire and Poulain (2016) showed a peculiar exception among prokaryotes: photosynthetic microorganisms from the group of purple non-sulphur bacteria (representing genera Rhodobacter and Rhodopseudomonas) are able to use Hg as an electron acceptor during photosynthesis.

Mercury was identified thousands of years ago and is one of the oldest toxicants known. The three forms of Hg, i.e. elemental, inorganic and organic mercury (especially CH₃Hg-R; methyl-Hg or MeHg), have different toxicological properties. Mercury can occur in compounds either in +1 or +2 oxidation state, i.e. in mercurous(I) and mercuric(II) compounds, respectively. In nature, inorganic divalent Hg(II) compounds predominate, with relatively few monovalent Hg(I) compounds. Monovalent Hg compounds are less toxic than Hg(II) compounds as they are less soluble in water (WHO 2003; Park and Zheng 2012).

The biogeochemical cycle of Hg and toxicity involve bacteria that produce MeHg. In the environment some anaerobic sulphate- and iron-reducing bacteria can methylate oxidized mercury (Hg²⁺) and to a smaller degree Hg⁰, thus generating MeHg (Hu et al. 2013; Li and Cai 2013). Biologically mediated production of MeHg predominantly occurs under anaerobic conditions in sediments of inland waters, nearshore and oceanic sea floors, as well as in peatlands, wetland soils and some rice paddy fields, for example, in China (Zhang et al. 2010; Gu et al. 2011; Windham-Myers et al. 2014; Zhao et al. 2016). MeHg is also present in most if not all aquatic organisms. Methylation of InHg to MeHg and demethylation of MeHg are the two most important processes in the cycling of MeHg, determining the levels of MeHg in aquatic and terrestrial ecosystems. Aerobic bacteria have evolved an efficient strategy of eliminating mercuric (Hg²⁺) and organic mercury compounds (including MeHg) from the environment through the reduction of Hg²⁺ to Hg⁰ (Li and Cai 2013).

Methylation and biomagnification of Hg have been well researched in aquatic ecosystems due to the consumption of Hg-contaminated fish, crayfish and molluscs, which may lead to poisoning of humans and other warm-blooded vertebrates. By contrast, studies on Hg and especially MeHg in terrestrial ecosystems are few (Clarkson 1992; Larosa and Allen-Gil 1995; Wolfe et al. 1998; Jackson et al. 2011; Douglas et al. 2012; Kalisinska et al. 2012a; Rieder et al. 2013; Scheuhammer et al. 2015). Since MeHg in aquatic ecosystems is subject to biomagnification, Hg reaches its highest levels in predatory fish, piscivorous birds and marine and semiaquatic mammals. Mercury concentrations in those biotas can be many millions of times greater than in the waters which serve as their aquatic habitat or food source (Lavoie et al. 2013; Finley et al. 2016). The greatest increase in MeHg concentration occurs in the trophic step between water and algae. It is estimated that the biomagnification factor (BMF) between water and seston often ranges from ~10⁵ to ~10⁶ with the BMF of MeHg concentrations between successive trophic levels above algae generally less than 10¹ (Wolfe et al. 2007). In terrestrial ecosystems, biomagnification of MeHg also occurs, yet this phenomenon has been much less researched (Rimmer et al. 2010; Rieder et al. 2013; Osborn et al. 2011; Jackson et al. 2015; Abeysinghe et al. 2017).

3.4 Mercury Toxicity

In the 1950s, dramatic events took place in the Japanese Bay of Minamata with many lethal mercury poisonings in humans, cats and wild birds. Over 3000 brain-damaged victims were diagnosed with “Minamata disease”, and veterinary medicine introduced the term “dancing cats” to describe the neurological symptoms observed in cats. Both “Minamata disease” and “dancing cats” were the result of Hg poisoning accompanied by other contaminants spilled into the gulf from a nearby chemical factory. In the gulf’s sediments, bacteria transformed inorganic mercury into MeHg, whose levels progressively increased in organisms from successive trophic levels. Large amounts of MeHg in fish, crustaceans and mussels were consumed by humans and animals inhabiting those areas, resulting in diseases and fatal poisonings (D’Itri 1991; Aronson 2005; Hachiya 2006; Ekino et al. 2007; Grandjean et al. 2010). Also in the 1950s, MeHg toxicity in the developing brain was first recognized in cases of congenital Minamata disease among newborns and children. At the same time, it was noted that the mothers had no symptoms of Hg toxicity or were minimal (Clarkson and Magos 2006; Ekino et al. 2007).

A few later studies from the 1960s to 1970s were conducted by Swedish naturalists on birds and rodents feeding on grains and on predators feeding on these granivores. They showed that Hg poisoning can also occur in terrestrial environments, not just aquatic environments. Inorganic and organic Hg compounds (including MeHg) were then common components of pesticides (fungicides) serving as seed dressing. Large quantities of Hg from the fungicides were detected in granivores and even larger levels in predatory birds and mammals preying on the passerines and rodents (Borg et al. 1969; Johnles and Westermark 1969). From 1960 to 1990, Hg-containing fungicides had been banned in Northern Hemisphere countries with highly developed agriculture (UNEP 2002). After all those years, it is very difficult to determine how much of the Hg pesticides has been introduced into the environment since the usage (launched in the first quarter of the twentieth century) lasted dozens of years. In the United States, Sweden and Japan, it is estimated that 800, 600 and 1600 tons of Hg fungicides were sprayed each year in rural areas of those countries (with Japan being more than 20 times smaller in area than in the United States) (Smart 1968; Kiesling and Lloyd 1971). Currently, agricultural soils are also being contaminated with anthropogenic Hg due to fertilization with sewage sludge, but this process is much less intense. It is estimated that in the EU, the Hg concentrations in sewage sludge recycled to agriculture vary among its member states from 0.2 to 4.6 mg kg^–1 dw (Milieu Ltd. WRc and RPA 2010). In the 2000s the amount of mercury introduced into agricultural soils in the 27 EU countries probably exceeded 4 tons per year (AMAP/UNEP 2013). Total Hg from atmospheric deposition (derived from natural and anthropogenic sources) of agricultural origin and released from soil rocks contributes to pollution of the terrestrial environment. Mercury is washed away from these areas and is transported to various waters bodies where it is methylated and (partly as Hg⁰) is released into the atmosphere and transported over considerable distances. In addition, soils in river valleys are exposed to various forms of Hg during periodic inundations. However, in aquatic environments, as compared to land, Hg is to a much greater degree integrated into food chains, and aquatic food can be a significant threat to the health of humans and wildlife. Generally free-living terrestrial animals are chronically exposed to low concentrations of Hg contained in food, water and ambient air. Mercury toxicity has been studied at the levels of molecules, cells, tissues, organisms, species and ecosystems (Borg et al. 1969; Wren 1984; Scheuhammer et al. 1998a, b; Aschner 2000; Schurz et al. 2000; Silva-Pereira et al. 2005; Wolfe et al. 2007).

The toxicity of mercury has been attributed to its high affinity to protein-containing sulfhydryl (thiol) groups (–SH). These groups are especially abundant in proteins containing cysteine and methionine, which are sulphur amino acids. Proteins rich in cysteine include glutathione peroxidase (GSH-Px), metallothioneins (MTs) and keratins. GSH-Px belongs to the family of very important antioxidant enzymes, which also contain selenium (Se) (Clarkson and Magos 2006). MTs and keratin structures (including hair and feathers) may contain up to 30% and 26% of cysteine, respectively (Clarkson and Magos 2006; Agarwal and Behari 2007; Greenwold and Sawyer 2013). The MTs are low-molecular-weight proteins and are present in various cells (especially in the liver and kidneys) and serum of vertebrates, but they were also discovered in invertebrates. MTs have a few main hypothesized functions: homeostasis of essential metals such as zinc (Zn) and copper (Cu), detoxification of non-essential Hg and cadmium (Cd), protection against oxidative damage and free radical scavenging (Isani and Carpenè 2014).

All mercury species are accumulated by eukaryotic organisms. Vertebrates can uptake toxic mercury from the environment through the lungs, gills, skin and from the digestive tract. In wildlife the alimentary tract plays the most important route. From avian and mammalian gastrointestinal tracts, MeHg is most effectively absorbed at a rate over 90%. InHg is absorbed from the diet, at most at a rate of a few to a dozen percent, and Hg⁰ at <0.01% (Serafin 1984; Clarkson and Magos 2006; Park and Zheng 2012; Ye et al. 2016). Inhaled Hg⁰ vapour in the lungs of mammals is absorbed at up to 85%, as demonstrated by experimental research on mammals and epidemiological studies of humans occupationally exposed to mercury vapour (Pendergrass et al. 1997; Falnoga et al. 2000; Bose-O’Reilly et al. 2010; Bernhoft 2012).

Mercury toxicity studies have taken into account many factors, including the physico-chemical properties of this element. Mercury is classified as a chalcophile element (alongside Se, Cd and Pb), with a typically higher affinity to sulphur (S) and a lower affinity to oxygen (O) than iron (Fe). In living organisms, Hg is highly competitive in relation to essential metals, mainly Zn and Cu, which are displaced from the S binding sites in cysteine to be replaced by Hg⁺² and/or MeHg⁺. Sulphur amino acids (cysteine, Cys, and methionine, Met) are constituents of enzyme, transport and structural proteins, which after binding to Hg change their properties and structure (Grosicki and Kowalski 2002; Fraga 2005; García-Barrera et al. 2012; Dobrakowski et al. 2013). In the case of Cys, over the course of evolution, S has been replaced by Se to form the 21st amino acid, selenocysteine (SeCys). It is a natural component of selenoproteins in all animal kingdoms including vertebrates (Lu and Holmgren 2009). From this group of proteins, the most important are enzymes such as GSH-Px, thioredoxin reductase and iodothyronine deiodinase. These proteins participate in the antioxidant protection of cells and the metabolism of thyroid hormones and of immunological processes. Selenoproteins may contain from 1 to 15 SeCys per protein subunit (Ralston et al. 2008; Mehdi et al. 2013). MeHg⁺ ions possess electrophilic properties, and they interact with and oxidize nucleophilic groups of various biomolecules, especially those containing sulfhydryl groups. Besides proteins (i.e. antioxidant enzymes, neurotransmitter receptors, transporters), sulphydryl groups contain nonprotein thiols such as cysteine and glutathione, GSH (Farina et al. 2013). GSH is an important antioxidant in animals, preventing damage to cellular components caused by reactive oxygen species and other factors including Hg⁺² and MeHg⁺ (Schurz et al. 2000; Pompella et al. 2003; Clarkson and Magos 2006; Wolfe et al. 2007).

As the binding affinity of Hg for Se is up to a million times higher than for S, Hg (especially Hg²⁺ and MeHg⁺) inexorably sequesters Se, directly impairing selenoenzyme activity and synthesis. At the same time, Se compounds are able to decrease the toxicity of Hg, which has been established in all investigated species of mammals, birds and fish (Dietz et al. 2000; Belzile et al. 2009; Ralston and Raymond 2010).

3.4.1 Mercury Cytotoxicity, Genotoxicity, Cancerogenicity and Teratogenicity

The cytotoxicity and genotoxicity of the various forms of Hg are evaluated mainly in vitro assays on human and non-human cell lines (De Flora et al. 1994; Silva-Pereira et al. 2005; Robinson et al. 2010; Polunas et al. 2011; Fernandes Azevedo et al. 2012; Roy et al. 2013; Wang et al. 2013, 2016). The results of in vivo Hg genotoxicity tests (based mostly on leucocytes) that assessed the damage of nuclear genetic material (comet assay, micronucleus test, chromosome aberration tests) do not always confirm differences between the material obtained from warm-blooded vertebrates exposed to Hg and from control/comparison groups (Hansteen et al. 1993; Rozgaj et al. 2005; Kenow et al. 2008; Crespo-López et al. 2009). Various ions of Hg exhibit a high ability to bind –SH groups of protein and nonprotein compounds, and on this ground a number of hypotheses have been formulated about molecular mechanisms of Hg genotoxicity. In this respect, the most commonly mentioned are four mechanisms: oxidative stress connected with generation of free radicals, effect on microtubules, influence on DNA repair mechanisms and direct interaction with DNA molecules (De Flora et al. 1994; Myers and Davidson 1998; Burbacher et al. 1990; Crespo-López et al. 2009).

In vitro cytotoxicity studies have shown that in various human and animal cell lines, both InHg and MeHg induce numerous adverse changes. These changes mainly rely on altering mitochondrial function and raising oxidative stress by generating free radicals or by interacting with sulfhydryl groups (Polunas et al. 2011; Farina et al. 2013; Agrawal et al. 2015; Wang et al. 2013, 2016). Additionally, in human embryonic neural progenitor cells, MeHg induces oxidative damage to mitochondrial DNA (Wang et al. 2016).

The carcinogenetic potential of InHg is still being debated. In the 1990s the International Agency for Research on Cancer (IARC 1993) classified MeHg compounds as possibly carcinogenic to humans (Group 2B), but metallic mercury and InHg compounds were not classifiable as carcinogenic to humans (Group 3). The classifications of MeHg, Hg⁰ and InHg are still in use by the IARC (2017). Drasch et al. (2004) reviewed papers concerning the influence of Hg on laboratory rats and mice and revealed that male rats receiving extremely high oral doses of HgCl₂ or MeHgCl had an increased number of renal tubule tumours. These compounds caused severe nephropathy in the rodents. It is likely that dietary MeHgCl may act in mice as a tumour promoter. However, the connection between Hg exposure and carcinogenesis remains controversial (Drasch et al. 2004; Crespo-López et al. 2009).

Methylmercury is known as an embryotoxic and teratogenic agent. The teratogenicity of MeHg is well documented in fish, birds, mammals and humans. This compound especially affects normal development of the central nervous system (Domingo 1994; Samson and Shenker 2000; Schurz et al. 2000; Heinz et al. 2011; Gandhi et al. 2013). In addition to the already mentioned adverse effects of Hg on wildlife, its effects on endocrine and immune systems are also important (Zhu et al. 2000; Kenow et al. 2007; Tan et al. 2009). Tan et al. (2009) listed five main endocrine-related mechanisms of Hg across these systems: (a) accumulation in the endocrine system, (b) specific cytotoxicity in endocrine tissues, (c) changes in hormone concentrations, (d) interactions with sex hormones and (e) upregulation or downregulation of enzymes within the steroidogenesis pathway. However, disorder and impairment of endocrine and immune systems by Hg and the net effects on the demography of wild animals are poorly understood (Kenow et al. 2007; Frederick and Jayasena 2011).

3.4.2 Mercury Neurotoxicity and Lethal Levels of Total Mercury in Soft Tissues

All three Hg species may occur in the brain, including elemental Hg. A certain part of inhaled Hg⁰ is deposited in the brain as demonstrated in humans and laboratory animals (Warfvinge et al. 1992; Tjälve and Henriksson 1999; Bose-O’Reilly et al. 2010; Park and Zheng 2012). Also Hg⁰ can be transported through the olfactory pathway to the olfactory bulbs and later into other brain areas (Galić et al. 1999; Tjälve and Henriksson 1999; Park and Zheng 2012). As Hg⁰ is lipid soluble and highly diffusible, it can cross the blood-brain barrier and other cellular and intracellular membranes (Park and Zheng 2012). In humans inhalation of Hg⁰ vapour can cause acute and chronic intoxication. Typical symptoms connecting with this include airway symptoms and many neurological problems (tremor, ataxia, coordination disturbances, abnormal reflexes, polyneuropathy with sensation difficulties, loss of memory, neurocognitive disorders) as well as kidney problems such as proteinuria (Bose-O’Reilly et al. 2010). In laboratory animals, the neurological symptoms following exposure to Hg⁰ are poorly understood, but in rats a significant increase in Hg concentrations in different parts of the brain (primarily in the neocortex, in the basal nuclei and in the cerebellar granule cells) and in the kidneys were shown in comparison to the control group (Warfvinge et al. 1992; Galić et al. 1999). Unlike elemental and organic mercury species, the oxidized Hg forms (Hg²⁺) are not able to effectively cross the blood-brain barrier, but such process could not be excluded (Park and Zheng 2012). Organic Hg compounds, especially MeHg, can easy cross the blood-brain barrier (however, less efficiently as Hg⁰) and are accumulated in vertebrate brains. The probable active transport of MeHg (via neutral amino acid transporters) into the brain is preceded by the formation of MeHg-cysteine complexes (ADSTR 1999; Clarkson and Magos 2006). MeHg does not uniformly affect the nervous system, and Hg concentration in the brain varies between the compartments (Eto et al. 1999, 2010; Farina et al. 2013).

Speciation analyses of brain Hg in vertebrates show that a much larger proportion of THg is present in the form of MeHg (typically >80%) and a small fraction as InHg. Depending on the degree and duration of exposure to MeHg, the percentage of brain THg may change over time and varies greatly between individuals of the same species and between various species. In extreme cases in some people exposed to MeHg in childhood and for more than 20 years, up to 80% of brain THg may be InHg (Farina et al. 2013). Most wildlife is exposed to long-term exposure to small amounts of MeHg contained in the diet, with the exception of long-living piscivorous species. MeHg, which has penetrated into the brain, is gradually demethylated and transformed into InHg. The demethylation of MeHg has been found in the brain of humans and several wild species of birds and mammals from inland environments (Eto et al. 1999; Gnamuš and Horvat 1999; Scheuhammer et al. 2008; Strom 2008; Eagles-Smith et al. 2009; Haines et al. 2010; Basu 2012; Kalisinska et al. 2014a; Jo et al. 2015). Presumably, the remaining part of brain InHg can occur in insoluble and biologically inert compounds with selenium such as tiemannite (HgSe) (Whanger 2001; Nakazawa et al. 2011). In long-lived animals and humans, the half-life for MeHg in the brain is determined in days or months, but for InHg it is many years (Vahter et al. 1994; ADSTR 1999; Rice et al. 2014). Until recently it had been assumed that MeHg that gets into the brain did not leave, similar to InHg produced by MeHg demethylation or oxidation of Hg⁰. However, works by Lohren et al. (2015, 2016), investigating MeHg and InHg transfer across the blood-brain barrier in a primary in vitro porcine model, may lead to the revision of this view. In the latter paper, Lohren et al. (2016), the researchers documented the transport of MeHg via the barrier in both directions, with diffusion as the transfer mechanism. Additionally for HgCl₂, their data delivered evidence that the blood-brain barrier transfers InHg out of the brain.

Lethal brain levels of Hg have not yet been established for most mammals and birds. In literature, in the brains of piscivorous mammals experimentally intoxicated with MeHg, river otter Lontra canadensis and American mink Neovison vison (previously Mustela vison), Hg levels were 23.7 and 19.9 mg kg^–1 ww and 11.9 mg kg^–1 ww, respectively (Aulerich et al. 1974; Wobeser et al. 1976; O’Connor and Nielsen 1981). In field studies in North America, brain Hg in single dead or dying river otter and mink were ~30 and 13.4 mg kg^–1 ww, respectively (Wren 1985; Sleeman et al. 2010; Wobeser and Swift 1976). A lower range was shown by THg concentrations (8.1–18.6 mg kg^–1 ww) in experimentally and non-intentionally intoxicated domestic cats from Japan and Northwestern Ontario Reserve (Canada), which revealed neurological symptoms typical for Minamata disease (Takeuchi et al. 1977). Shore et al. (2011) defined >10 mg THg kg^–1 ww as a lower indicative value in mammalian brains, which may be associated with adverse effects on survival and resulting in death. Krey et al. (2015) analysed a large number of reports on mammalian brain Hg concentrations and proposed a THg threshold concentrations for toxic endpoints: clinical symptoms >6.75 mg kg^–1 ww (29 mg kg^–1 dw), neuropathological signs >4 mg kg^–1 ww (17.2 mg kg^–1 dw), neurochemical changes >0.4 mg kg^–1 ww (1.72 mg kg^–1 dw) and neurobehavioral changes >0.1 mg kg^–1 ww (0.43 mg kg^–1 dw).

In adult passerines (starling Sturnus vulgaris, grackle Quiscalus quiscula, red-winged blackbird Agelaius phoeniceus, brown-headed cowbird Molothrus ater, zebra finch Poephila guttata and piscivorous great egret Ardea albus), which were experimentally intoxicated with MeHg, the concentration of brain THg was in the range of 20–45 mg kg^–1 ww (Finley et al. 1979; Scheuhammer 1988; Spalding et al. 2000). The highest THg residues in brains among wild birds found dead in fields were within the range of 13–14 mg kg^–1 ww: in tawny owl Strix aluco from Norway feeding on small rodents, piscivorous white-tailed eagle Haliaeetus albicilla from Sweden and common loon Gavia immer from Canada (Jensen et al. 1972; Holt et al. 1979; Scheuhammer et al. 2008). The values found in wild birds were clearly smaller than in experimental studies, but free-living animals are more exposed to various adverse environmental factors, including food shortages, than captive ones (Van der Molen et al. 1982; Wiener et al. 2003). A combination of the environmental factors can cause premature death before brain THg levels in birds reach ≥20 mg kg^–1 ww, established as lethal in laboratory conditions. In addition, dead animals are quickly eaten by scavengers, which is why they are rarely obtained for analysis. It has been experimentally shown that chicks are more sensitive to the toxic effects of Hg than adult birds. Data presented by Heinz and Locke (1976) indicate that lethal brain THg levels can be as low as 3–7 ppm in mallard ducklings. Shore et al. (2011) suggested indicative values of THg concentrations for avian brains of non-marine species, which may be associated with bird deaths at >15 mg kg^–1 ww and >3 mg kg^–1 ww in adults and developing youngs, respectively, and correspond well to data from Jensen et al. (1972), Holt et al. (1979), Scheuhammer et al. (2008) and Heinz and Locke (1976). Neurological symptoms (e.g. trembling) have been observed in 1 hatch-year osprey with 1.2 mg kg^–1 ww THg in the brain (or 6.2 mg kg^–1 dw) (Hopkins et al. 2007). However, THg threshold concentrations for toxic endpoints analogous to those proposed for mammals have not been developed, i.e. ones that would include subclinical, neuropathological, neurochemical and neurobehavioral changes, although some attempts have been made in this regard (Scheuhammer et al. 2008; Rutkiewicz et al. 2011; Rutkiewicz 2012).

Mercury is not only neurotoxic but also nephrotoxic for elemental and inorganic mercury species. The kidney is a major repository of InHg in warm-blooded vertebrates. Within the kidney divalent Hg accumulates primarily in the cortex and outer stripe of the outer medulla (Aschner and Aschner 1990; Wolfe et al. 1998; Bridges and Zalups 2010). It should be underlined that birds differ from mammals in having a renal portal system. In birds the venous blood from the terminal portion of the digestive tract flows to the kidney rather than the liver, as in mammals. This may make the avian kidney more vulnerable than the mammalian (Wolfe et al. 1998). Indicative values of THg concentrations in mammalian kidney associated with death, as proposed by Shore et al. (2011), are lower than in avian species, >25–30 mg kg^–1 ww compared to >40 mg kg^–1 ww. Also THg indicative value estimated for the adult mammalian brain are lower than in the analogous avian organ. However, in the case of the liver, the indicative value is higher in mammals than birds: >25–30 THg kg^–1 versus >20 mg THg kg^–1 (Shore et al. 2011).

Lethal concentrations of THg in the soft tissues of mammals and birds are most commonly determined in the liver and kidney, followed by the brain. Muscles are rarely taken into consideration, although they constitute a large part of the body weight of the animals, and the collection of muscle samples is easy when compared to the brain (Finley et al. 1979; O’Connor and Nielsen 1981; Wren et al. 1987; Farrar et al. 1994; Thompson 1996; Shore et al. 2011; WVDL 2015). In addition, the efficient functioning and coordination of skeletal muscles play a key role, especially in predators, because they co-determine the effectiveness of hunting. Based on relatively scarce data concerning THg concentrations in tissue pairs: liver (L)–muscle (M) and muscle–brain (B) in adult individuals of wild animals and birds, and those experimentally intoxicated with organomercury, correlation coefficients (r) and the values of two indices M_THg/L_THg and B_THg/M_THg were calculated, and potentially lethal muscle THg concentrations were estimated. In both animal groups, an increasing hepatic THg concentration was initially accompanied by a marked increase in muscle levels (Fig. 17.1, panels a and b). After exceeding ~10 mg kg^–1 ww in the muscle, the increase in THg slowed down and stabilized at 25–35 and 25–40 mg kg^–1 ww in mammals and birds, respectively, while the hepatic THg significantly exceeded 100 mg kg^–1 ww over time in some cases. Among inland mammalian and avian species, the highest hepatic THg levels were detected in river otter (96 mg kg^–1 ww) and common loon (200 mg kg^–1 and 370 mg kg^–1 ww) (Wren 1985; Stone and Okoniewski 2001; Scheuhammer et al. 2008). In the livers of marine mammals and birds, levels exceeding 1000 and 200 mg THg kg^–1 ww, respectively, were found in some cases (Kim et al. 1996; Storelli et al. 1999; Pompe-Gotal et al. 2009).

Fig. 17.1

Relationship between total mercury (THg) concentrations (mg kg^–1 = ppm ww, wet weight) in liver and muscle, muscle and brain in terrestrial mammals and birds. Panel (A) Used data of intoxicated mammals such as river otter Lontra canadensis (O’Connor and Nielsen 1981), American mink Neovison vison (Aulerich et al. 1974; Wobeser et al. 1976), cat Felis catus (Charbonneau et al. 1974), ferret Mustela putorius furo (Hanko et al. 1970) and wild animals such as river otter (Sheffy and St Amant 1982; Wren 1985; Langlois and Langis 1995; Fortin et al. 2001; Strom 2008; Sellers 2010; Sleeman et al. 2010; Dornbos et al. 2013), Eurasian otter Lutra lutra (Hernández et al. 1985; Hyvärinen et al. 2003; Lodenius et al. 2014), American mink Neovison vison (Sheffy and St Amant 1982; Langlois and Langis 1995; Fortin et al. 2001; Wobeser and Swift 1976). Panel (B) Used data of intoxicated birds such as cowbird Molothrus ater (Finley et al. 1979), redwing Agelaius phoeniceus (Finley et al. 1979), starling Sturnus vulgaris (Finley et al. 1979), grackle Quiscalus quiscula (Finley et al. 1979), American kestrel Falco sparverius (Bennett et al. 2009), mallard Anas platyrhynchos (Hough and Zabik 1972) and wild birds from nature such as osprey Pandion haliaetus (Holt et al. 1979; Norheim and Frøslie 1978; Evers et al. 2005; Hopkins et al. 2007; Kalisinska et al. 2014a), bald eagle Haliaeetus leucocephalus (Evers et al. 2005; Evans 1993), white-tailed eagle H. albicilla (Norheim and Frøslie 1978; Kalisinska et al. 2014a; Henriksson et al. 1966; Falandysz et al. 2000), common loon Gavia immer (Evers et al. 2005; Scheuhammer et al. 1998b), common merganser Mergus merganser (Langlois and Langis 1995; Scheuhammer et al. 1998b; Kalisinska et al. 2010). Panel (C) Used data of intoxicated mammals: river otter (O’Connor and Nielsen 1981), American mink (Aulerich et al. 1974; Wobeser et al. 1976), cat (Charbonneau et al. 1974), ferret (Hanko et al. 1970) and wild mammals from nature such as river otter (Sheffy and St Amant 1982; Wren 1985; Fortin et al. 2001; Strom 2008; Sleeman et al. 2010; Dornbos et al. 2013), American mink (Sheffy and St Amant 1982; Fortin et al. 2001; Wobeser and Swift 1976), Eurasian otter (Kalisinska et al. 2016, 2017); intoxicated birds, cowbird (Finley et al. 1979), redwing (Finley et al. 1979), starling (Finley et al. 1979), grackle (Finley et al. 1979), mallard (Hough and Zabik 1972); and wild birds from nature, osprey (Holt et al. 1979; Hopkins et al. 2007; Kalisinska et al. 2014a), bald eagle (Evans 1993), white-tailed eagle (Holt et al. 1979; Kalisinska et al. 2014a; Henriksson et al. 1966; Jensen et al. 1972), common merganser (Kalisinska et al. 2010)

In the multispecies groups of mammals and birds, the correlation coefficient between the concentration of THg in the liver and muscle exceeded 0.95, and the values of r were, respectively, 0.928 and 0.964 (Fig. 17.1, panels a and b). Using the appropriate equations from panels A and B, we calculated THg concentrations for avian and mammalian muscle when the concentration of hepatic THg reached the lower limit values of the estimated lethal range (25 and 20 mg kg^–1 ww, respectively) (Shore et al. 2011). At these hepatic THg concentrations in mammalian and avian muscle, potentially lethal values were 9.8 and 7.3 mg kg^–1 ww. Other researchers had also found a significant correlation (r ranging from 0.60 to 0.98) between muscle and hepatic THg concentrations in inland mammals (Lord et al. 2002; Millan et al. 2008; Strom 2008; Kalisinska et al. 2009; Lodenius et al. 2014) and birds (Hopkins et al. 2007; Eagles-Smith et al. 2008), although not always (Halbrook et al. 1994; Kalisinska et al. 2010; Lanocha et al. 2014). These ambiguous results may be related to the large variations of hepatic THg concentration and M_THg/L_THg index in endothermic animals exposed to Hg. The mean value of the index is statistically higher in mammals than birds (0.42 versus 0.31, t = 2.34; p < 0.03). Wolfe et al. (2007) emphasized a poor correlation between liver THg concentration and its effects. Unlike the liver, the muscle THg concentration is more representative of brain THg concentration and correlates better with its effect. Moreover, MeHg is a dominant species of Hg in the brain and muscle tissues. These suggestions are supported by our analysis of data on THg concentration in the muscle and brain of mammals and birds combined into one group (Fig. 17.1, panel c). The correlation coefficient for this relationship exceeded 0.97, and values of the index B_THg/M_THg for mammals and birds were close, at 0.73 and 0.82, respectively. In another study, Shore et al. (2011) suggested that the lethal concentrations of THg in the brain of mammals and birds are >10 and >15 mg kg^–1 ww, respectively. Taking into consideration the equation from panel c, it may be assumed that the lethal THg level in muscle is about 13 mg kg^–1 ww for mammals and 18 mg kg^–1 ww for birds. On the basis of equations from Fig. 17.1, it can be assumed that the lethal concentration of THg in the muscles of mammals and birds is in the range 10–13 mg kg^–1 ww and 7–18 mg kg^–1 ww, respectively. Heinz (1996), based on literature data, estimated that muscle Hg concentrations associated with harmful Hg exposure in adult birds ranged from 15 to 30 mg kg^–1 ww. In the context of our analysis of avian muscle, it seems that the lower value suggested by Heinz (1996) is more likely.

3.4.3 Inorganic and Organic Mercury Distribution in Bodies of Mammals and Birds

The three forms of Hg (elemental, inorganic and organic) that penetrate the organisms of vertebrates differ with respect to their toxicokinetics regarding absorption, distribution and accumulation. In laboratory studies, the influence of MeHg (in MeHgCl form) and mercury compounds of Hg(II) (especially HgCl₂) are most frequently investigated. Mercurous mercury Hg(I), for example, in the form of mercurous chloride (Hg₂Cl₂), is little absorbed in the body. This compound readily dissociates in body fluids where, from Hg₂Cl₂, double atom cations of Hg₂ ²⁺ are realized and from this is formed one atom of divalent Hg²⁺ and another of elemental mercury (Hg⁰). Elemental mercury from this unimportant source and the vapour of this metal from inhaled air are oxidized into the mercuric form (Hg²⁺) in erythrocytes and tissues. Both inorganic and organic Hg species are excreted primarily in faeces. Absorption of MeHg from the digestive tract in warm-blooded vertebrates is very high (about 90%), with a great amount of it excreted in faeces (about 85–90%) and 5% with urine. Scientists have estimated that only up to 15% of absorbed MeHg is incorporated in various tissues and organs. Fur or hair in mammals as well as feather in birds are also an important route of excretion, especially MeHg (Farris et al. 1993; Clarkson and Magos 2006; Wolfe et al. 2007).

Mammalian pelt and avian plumage sometimes incorporates even >80% of THg in the body. MeHg is permanently built into hair and feathers during their growth. It is a dominant species of Hg in these tissues and becomes biologically inactive there, as confirmed in studies on experimental animals and wildlife from inland ecosystems. After long exposure to MeHg in laboratory experiments and chronic exposure of wildlife, MeHg and/or THg concentrations in these keratin skin structures usually reach the highest values in comparison to liver, kidney, brain and muscle THg (Thomas et al. 1988; Farris et al. 1993; Wood et al. 1996; DesGranges et al. 1998; Mierle et al. 2000; Hyvärinen et al. 2003; Bennett et al. 2009; Lieske et al. 2011; Nam et al. 2012; Wang et al. 2014). However, THg and/or MeHg are rarely assayed in all of the mentioned tissues in the same individuals. Eventually, MeHg is removed from mammalian and avian bodies during moulting, and therefore hair and feathers are also an important additional route of Hg excretion (Honda et al. 1986; Farris et al. 1993; Clarkson and Magos 2006; Wolfe et al. 2007; Wang et al. 2014; Evans et al. 2016). After Hg in fur and feathers, the second largest Hg pool can be found in skeletal muscles, with up to 50% of the remaining MeHg in the body (Farris et al. 1993; DesGranges et al. 1998; Saeki et al. 2000; Nam et al. 2005) from the large proportion of skeletal muscles in the body mass of vertebrates and their vasculature. For example, in the body of predatory mammals and birds, these muscles represent on average 50–55% and 30–40% of body mass, respectively (Honda et al. 1986; Biewener 2011; Muchlinski et al. 2012), and in the case of fish, it is up to 70% of their body weight (Kisia 1996). In the muscles of warm-blooded vertebrates from inland ecosystems, Hg occurs mainly in the form of MeHg (70–95% of THg), and the concentration is usually low (<0.50 mg kg^–1 ww), with the exception of the muscles of fish species near the top of a food web and piscivorous wildlife (Wren et al. 1980; Mason et al. 1986; Rothschild and Duffy 2005; Kinghorn et al. 2007; Strom 2008; Ruelas-Inzunza et al. 2009; Chumchal et al. 2011; Burger et al. 2013; Hall et al. 2014; Kalisinska et al. 2014a, b, 2017; Wentz et al. 2014). Observed transient storage of large amounts of MeHg in the muscle may protect other tissues against MeHg toxicity.

Because of the large proportion of muscles in body weight and easily digestible MeHg contained in them, they play an important role in the transfer of this substance from freshwater invertebrates and fish to semiaquatic piscivores and benthophages and from carrion of these animals to terrestrial scavengers (Sheffy and St Amant 1982; Halbrook et al. 1994; Langlois and Langis 1995; Fortin et al. 2001; Evers et al. 2005; Chumchal et al. 2011; Kalisinska et al. 2009, 2016). However, mercury, especially MeHg, is rarely assayed in the muscles of warm-blooded vertebrates. Among the tissues of terrestrial vertebrates, Hg achieves the highest concentration in the liver and kidneys, so THg is most frequently analysed in these organs, although in total they account for no more than 4–6% of the body weight of mammals and birds (Fischer and Bartlett 1957; Holliday et al. 1967; Hughes 1970; Kruska and Schreiber 1999; Lanszki et al. 2008; Balk et al. 2009; Kalisinska et al. 2010). In the kidney and livers of many fish-eating mammalian and avian species, the percentage of MeHg decreases as THg concentration increases in the organs (Norheim and Frøslie 1978; Wiener et al. 2003; Gamberg et al. 2005a). The liver and kidney have been suggested as one of the major sites of MeHg demethylation in mammals and birds. Above the threshold value of 10 mg THg kg^–1 dw (~3 mg THg kg^–1 ww), hepatic %MeHg declines rapidly from a high value (~90%) (Eagles-Smith et al. 2009). However, interspecies differences are observed in this respect, and hepatic intensification of MeHg demethylation in birds can occur already in the range of 5–7 mg THg kg^–1 dw, because then %MeHg in THg falls below 70% (Scheuhammer et al. 1998b; Dietz et al. 2013; Kalisinska et al. 2014c). Some researchers (Gamberg et al. 2005a; Martin et al. 2011) suggest that in piscivorous mammals (such as mink), the demethylation process of hepatic MeHg is activated well below the 30 mg THg kg^–1 dw threshold (10 mg kg^–1 ww) suggested by Wiener et al. (2003). Energy costs of MeHg demethylation in avian and mammalian livers are probably high but to date have not been estimated (Eagles-Smith et al. 2009; Dietz et al. 2013; Kalisinska et al. 2014c). Methylmercury demethylation is observed in tissues other than the liver and kidney but at a lower intensity and efficiency. This process is well documented in the brain of a number of mammals and birds, including terrestrial species. However, species of endothermic animals differ in the proportion of brain MeHg to THg (Vahter et al. 1994; Farina et al. 2003; Scheuhammer et al. 2008, 2015). It is generally assumed that demethylation of MeHg in fish and other vertebrate muscles does not occur or is negligible, with the percentage of MeHg in THg usually exceeding 80–90% (O’Connor and Nielsen 1981; Houserova et al. 2006; Strom 2008; George et al. 2011; Kalisinska et al. 2014b; Harley et al. 2015; Scheuhammer et al. 2015). However, in a few papers concerning the muscle of fish, birds and mammals, we can find data indicating that %MeHg can be ≤70%, especially in cases where THg < 0.5 mg kg^–1 ww. Pal et al. (2012) and Park et al. (2010) found in 8 out of 13 (8/13) and 5/13 investigated Asian freshwater fish species (generally with muscle THg 0.05–0.45 mg kg^–1 ww) mean values of %MeHg were in the range 50–69%. Sometimes in predatory freshwater fish (such as Elops machnata and Pelates quadrilineatus from Taiwan), whose muscles contained >1 mg THg kg^–1 ww, MeHg did not exceed 70% of THg (Huang et al. 2008). In three aquatic birds species from Mexico THg muscle levels varied from 0.32 to 0.45 mg kg^–1 ww, and the %MeHg was in the range of 26–61% (Ruelas-Inzunza et al. 2009). In two populations of white-tailed eagle from Europe, the share of MeHg reached 45 and 58% when the mean THg in the eagle muscle was just 4.8 and 0.46 mg kg^–1 ww, respectively (Norheim and Frøslie 1978; Kalisinska et al. 2014a). In the muscle of the piscivorous river otter mean, 72% MeHg of THg was sporadically revealed (THg = 0.89 mg kg^–1 ww, Wren et al. 1980), but in marine cetaceans a value <70% was very often noticed. In 11 out of 16 studied species, the means were in the range 36–67%, and THg concentrations varied from 1.0 to 39.5 mg kg^–1 ww (Endo et al. 2005). The data quoted above may indicate that MeHg demethylation in vertebrate muscles does occur, although this process requires further investigation and collection of more data. On the basis of comparative studies of two populations of blue shark (Prionace glauca) from the Azores and the Canary Islands, Branco et al. (2004) speculated that the diet of migrating animals may differ significantly in MeHg content due to differences in exposure to Hg at different locations. Periodic stays in areas where prey contains less MeHg promote gradual demethylation and elimination of MeHg already accumulated in the muscles of sharks, and at the same time the supply of new portions of MeHg with food to their organisms is then lower. Branco et al. (2004) found in the shark’s muscle from the Canaries %MeHg much lower than in sharks from the Azores 55–70% and >80%, respectively, although muscle THg concentrations were similar.

3.5 Mercury in Elements of Inland Food Chains

Food is the main source of Hg for humans and wildlife, but its absorption from digestive tract is strictly dependent on the chemical form and amount of Hg in various diets. Mercury concentrations increased from autotrophic organisms to herbivores < detritivores < omnivores < carnivores (Rimmer et al. 2010). For terrestrial herbivorous and omnivorous animals, plant, fungi and invertebrates are the most important components of their food. Soil invertebrates, insects, spiders and other arthropods or small- and medium-sized birds and mammals are eaten by different predators depending on their body size and food preferences. Some carnivore mammals and aquatic birds of inland habitats are highly specialized in catching fish. Below are presented some aspects of Hg transfer between different environmental components, including soil and plants, plants and warm-blooded herbivores as well as preys and predators.

3.5.1 Mercury in Plants and Mushrooms

The amount of MeHg in soils is low relative to THg, and the dominant form in soils is InHg (Burton et al. 2006). Bioavailability of soil InHg for plants is very low. A significant part of the InHg taken from the soil is retained in the roots, which are a barrier to mercury uptake. There is a positive correlation between the concentration of InHg in the soil and roots, but it does not occur between soil Hg and its content in shoots and leaves, which are about ten times lower than in the roots, and probably the transport of Hg from the roots to the stems either does not occur or is a very slow process (Wang and Greger 2004; Tomiyasu et al. 2005). The main soil factors affecting the collection of this toxic metal by plants include organic matter, oxygen and carbon, redox potential, Hg species and their concentrations and the presence of other metals in the soil solution (Tomiyasu et al. 2005; Patra and Sharma 2000; Azevedo and Rodriguez 2012). In plants, the dominant form is InHg, which is >97% THg (Mailman and Bodaly 2005; Dombaiová 2005). In unpolluted areas, THg concentration in leaves is negligible and is characterized by considerable variability, ranging from several to several dozen micrograms per kilogram of dry matter (μg kg^–1 dw). In addition to the species diversity of plants, it is related to seasonal variation. In young leaves, compared to older ones, at the end of the growing season, the concentration of mercury is an order of magnitude smaller. The mercury detected in the leaves basically comes from the surrounding air, most likely Hg⁰, entering through the stomata, and probably leaf uptake of Hg is irreversible (Bushey et al. 2008; Laacouri et al. 2013). In areas where Hg was mined (e.g. Almaden zone in Spain), Hg concentration in soils is many thousand mg kg^–1 dw, and in some herbal plant species, it reaches 7–23 mg THg kg^–1 dw (or 7000–23,000 μg kg^–1 dw), thousands of times greater than in plants in uncontaminated areas (Moreno-Jimenez et al. 2006; Laacouri et al. 2013). In contrast to InHg, which is absorbed by the root system and kept there, in wetlands MeHg enters more efficiently both to water plants and through the roots to the aerial parts of plants (Patra and Sharma 2000; Windham-Myers et al. 2014). This organic species of mercury in plants from paddy fields may reach levels up to 63 μg kg^–1 dw in rice grain and pose a significant health risk to people, as has been demonstrated for rice grown on soils with a high concentration of Hg in Asia (Qiu et al. 2012; Rothenberg et al. 2014). Probably, due to the consumption of rice grain from such areas, not only humans but also grain-feeding animals (especially granivorous birds) are at risk of MeHg intoxication.

Of the nonanimal inland organisms, fungi are considered the greatest accumulator of Hg from the soil (Falandysz and Borovička 2013). Usually, higher Hg concentrations are detected in these than in their substrates, and fungi accumulate especially high levels in the areas of geochemical anomalies such as the mercuriferous Eurasian belt (including Almaden in Spain, Monte Amiata in Italy and Chinese Yunnan Province). In the mushrooms found there, the average concentration of THg varies from 1 to 100 mg kg^–1 dw (Bargali and Baldi 1984; Esbri et al. 2011; Falandysz et al. 2015). The concentration of MeHg in mushrooms is generally low and ranges between 0.01 and 3.7 mg kg^–1 dw, with the proportion of MeHg in THg not exceeding 5% (Bargali and Baldi 1984; Rieder et al. 2011).

3.5.2 Mercury in Earthworms

For some terrestrial invertebrates and vertebrates, the source of mercury may be soil contaminated with Hg. It is the essential food of earthworms or is an admixture for the intaken plant and animal foods of soil invertebrates, birds and mammals (Hargreaves et al. 2011; Rieder et al. 2013). In soils, over 90% of the invertebrate biomass may consists of earthworms. That is why, they are a significant object in ecotoxicological studies on Hg (Zhang et al. 2009; Teršič and Gosar 2012; Rieder et al. 2011; Abeysinghe et al. 2017). Concentrations of Hg in earthworm bodies depend on animal species and various soil conditions such as Hg forms and their amount, content of organic matter, pH and oxygen availability (Zhang et al. 2009; Rieder et al. 2013; Abeysinghe et al. 2017). Additionally, Rieder et al. (2011) demonstrated that earthworms inhabiting topsoils (endogenic) contained the highest concentrations THg and MeHg, followed by deep-burrowing earthworms (anecic) and litter-inhabiting organisms (epigeic). Methylated organic Hg species bioaccumulate more readily, and much higher bioconcentration factors (BCFs) from soil to earthworms are reported for MeHg than for THg (BCFs are calculated as THg or MeHg concentrations in the organisms divided by the corresponding concentrations in the soils). For example, in earthworms from Swiss forest, soils non-contaminated with Hg (mean THg level at 0.18 mg kg^–1 dw) mean THg and MeHg in all investigated earthworm groups were 1.04 and 0.09 mg kg^–1 dw, respectively. The share of MeHg in THg did not exceed 9%. BCF for THg and MeHg differed significantly: 7.2 vs. 83.1 (Rieder et al. 2011).

Analogical data has also been presented for earthworms living in rice paddy soils (Abeysinghe et al. 2017) sampled at various distances from abandoned mercury mines in Guizhou (China) and at control sites without a history of Hg mining. The highest mean THg concentrations were detected in the soil near the mines (80–125 mg kg^–1 dw). However, even at sites distant from a mine (7–8 km) and in control samples THg levels were quite high (~20 and ~0.6 mg kg^–1 dw, respectively). On the other hand, in those samples the concentration of MeHg was negligible and did not exceed 0.001 mg kg^–1 dw, with the proportion of MeHg in THg estimated at ≤0.01%. In earthworm bodies, mean concentrations of THg and MeHg decreased with the increasing distance from the mine. In the animal samples at sites distant 7–8 km from the mine and at control areas, the average THg concentrations were approximately 10 and 0.60 mg kg^–1 dw and for MeHg 0.10 and 0.05 mg kg^–1 dw, respectively. Share of MeHg in THg in the two group of earthworms was ~8 and ~2%, similar to levels reported by other researchers (Zhang et al. 2009; Rieder et al. 2011). In the study, BCFs for THg and especially for MeHg increased with distance from the mine. In earthworms sampled at 7–8 km from mines and at reference sites mean values of BCFs for THg were in the range 0.5–1.0 and almost three orders of magnitude lower than BCFs calculated for MeHg. Mean values of BCF for MeHg at control sites and 7–8 km from mines were about 900 and 300, respectively. Abeysinghe et al. (2017) suggested that specific soil conditions in rice paddies may make the earthworms important biomagnifiers of MeHg. Such large differences observed between the BCF for THg and BCF for MeHg in the case of earthworms and soils (even with negligible Hg contamination) are influenced by very high absorption of lipophilic MeHg from their intestine compared to InHg. This is probably due to the methylation of InHg occurring in their digestive tract due to the activity of the microbiota. At least two arguments for this are given by Rieder et al. (2013) on the basis of their experimental studies. Firstly, earthworms contained about six times higher concentrations of MeHg if they grew in soils treated with InHg than in soils without Hg. Secondly, the concentrations of MeHg in earthworm casts and in the soils were similar and did not change over time.

3.5.3 Mercury in Spiders and Insects

Studies of MeHg contamination of food webs have historically focused on aquatic organisms including those inhabiting inland reservoirs. However, recent reports have shown that terrestrial organisms such as songbirds, bats and reptiles can exhibit elevated Hg burden by feeding on MeHg-contaminated spiders and insects (Cristol et al. 2008; Jackson et al. 2011; Drewett et al. 2013; Yates et al. 2014; Gann et al. 2015). Studies in this field are mainly conducted in floodplains, riparian and wetland ecosystems of North America, which have documented historical influence of Hg pollutants. It has been shown that in such areas (especially not too distant from Hg point sources) MeHg in terrestrial predatory spiders from the Lycosidae family reach high concentrations, in the range 0.60–1.29 mg kg^–1 dw, which may be comparable or greater than in fish from neighbouring waters (Cristol et al. 2008; Speir et al. 2014; Gann et al. 2015; Standish 2016). In areas with negligible contamination or unpolluted with Hg, average concentration of MeHg in Lycosidae varies from 0.06 to 0.15 mg kg^–1 dw (Bartrons et al. 2015; Gann et al. 2015; Tavshunsky et al. 2017). Depending on trophic position (which can be derived from δ¹⁵N), other arthropods in the areas with the historically proven exposure to Hg may exhibit MeHg concentrations from 0.02 mg kg^–1 dw in herbivorous leafhoppers to 1.18 mg kg^–1 dw in detritofagous isopods (Cristol et al. 2008; Standish 2016).

Long-lived cicadas are another example of increased concentrations of MeHg in arthropods. The larvae of these insects live in the ground (2–17 years) and feed on root juice. In the Hg-contaminated soils, the effective absorption of MeHg occurs through the roots from where it can be taken up by cicadas. In Huludao City (NE China), with a chlor-alkali plant and two zinc smelters (industrial sources of Hg), its soils contained on average 4.08 mg THg kg^–1 dw and 0.009 mg MeHg kg^–1 dw. Cicadas Cryptotympana atrata from such soils accumulated in their bodies on average 0.124 mg MeHg kg^–1 dw, in a range from 0.021 to 0.319 mg MeHg kg^–1 dw (Zheng et al. 2010). Thus, these insects, although not associated with aquatic food chains, may constitute an important local source of MeHg intoxication for predatory arthropods, insectivorous birds, bats and other animals. The number of studies on bioaccumulation and biomagnification of MeHg in terrestrial food webs is gradually increasing, which should result in a better understanding and explanation of these processes. Importantly, this requires close cooperation between specialists in various fields, including zoology, ecology, toxicology of animals, plants and soils.

3.5.4 Transfer of Mercury from Inland Aquatic Ecosystems to Terrestrial Vertebrates

Compared to the Hg transfer between the links of the aforementioned food chains, much more data has been gathered on predatory warm-blooded vertebrates (including semiaquatic mammals and aquatic birds) that inhabit inland areas and feed on aquatic food, especially fish. Studies on the relationships between these organisms, taking into account Hg forms and their concentration levels, have been conducted at least since the mid-twentieth century. Their initiation was closely related to the dramatic events in the Gulf of Minamata and documented the neurotoxic and disruptive effects of Hg on reproductive processes in humans and other homeothermic animals. Fish (and in less degree shellfish) are considered most significant source of MeHg exposure for humans and wildlife. Therefore, many countries have set standards to protect humans from Hg in food. For example, in the EU the limit for Hg in freshwater fish for humans is 0.5 mg kg^–1 ww or 500 μg kg^–1 ww (1000 μg kg^–1 ww for pike Esox lucius and eel Anguilla anguilla) (Commission Regulation, EC 2006), and in the United States 300 μg MeHg kg^–1 ww is recommended (US EPA 2001, 2010). According to the EU Water Framework Directive, Environmental Quality Standards (EQS) for some chemicals in biota have been set, with mercury being defined as a priority hazardous substance (Directive 2008/105/EC). EQS are intended to protect top predators against secondary poisoning and refer to THg; for freshwater fish, the EQS for Hg (EQS/Hg) is at 20 μg kg^–1 ww. Apart from the EU, only Canada has a standard designed of Hg (MeHg) to protect fish-eating animals at 33 μg kg^–1 ww (Canadian Environmental Quality Guidelines 2000). The Canadian standard concerning Hg in freshwater fish is 65% higher than the European EQS/Hg. In North America the value of 100 μg kg^–1 ww in fish is of concern for the protection of piscivorous mammals, including mink and otters (Scudder et al. 2009). However, robust data on the dietary Hg exposure thresholds that result in deleterious effects, including disturbances in reproduction, exist only for very few bird species. Typical range of Hg effect thresholds are approximately from 200 to over 1400 μg kg^–1 ww in natural and/or experimental diets (Fuchsman et al. 2017). In North America, the piscivorous common loon has been intensively studied in field and laboratory settings (Evers et al. 2003; Kenow et al. 2008; Scheuhammer et al. 2008; Kenow et al. 2011). The dietary screening benchmark of 180 μg Hg kg^–1 ww in whole body prey fish was established for this species, characterized as moderately sensitive to Hg intoxication (Heinz et al. 2009; Depew et al. 2012).

The concentration of Hg in fish depends on the degree of environmental pollution with this metal, the intensity of Hg methylation, the size of fish (closely correlated with their age) and their trophic level (Depew et al. 2013a; Eagles-Smith et al. 2014). Because of the higher cost of MeHg analysis (2–3 times greater than that for Hg analysis), THg in various animal tissues is assayed in the most investigations, including monitoring studies. It is generally accepted that the Hg in fish muscle occurs in the form of MeHg, which accounts for ~90% of THg (US EPA 2010). Concentration of Hg in freshwater fish in various parts of the world varies considerably. The United States and Canada have very large databases on Hg concentration in many species of fish. These data (after appropriate selection, standardization and statistical processing) allow an estimate of Hg concentrations in prey (Hg_PREY) of piscivorous fish and wildlife and evaluate their potential. In North America ecological monitoring of Hg depends crucially on top piscivorous fish such as walleye Sander vitreus and northern pike Esox lucius and among fish from lower trophic levels—yellow perch Perca flavescens and largemouth bass Micropterus salmoides. Among piscivorous wildlife, Hg monitoring uses common loon, bald eagle (to a smaller extent), mink and river otter (Evers and Clair 2005; Evers et al. 2007; Depew et al. 2013a). The United States Geological Survey (USGS) developed the National Descriptive Model for Mercury in Fish (NDMMF, http://emmma.usgs.gov; Wente 2004), which was later adopted in Canadian reports (Depew et al. 2013b). For example, in standardized fish (collected in 1998–2005) coming from streams across the United States, fish Hg concentrations at 27% of sampled sites exceeded the US EPA human health criterion (300 μg kg^–1 ww). However, THg concentrations in fish from >66% of the sites exceeded the value of 100 μg kg^–1 ww that is of concern for the protection of piscivorous mammals. The highest mean Hg concentrations (between 1800 and 1950 μg kg^–1 ww) were noticed in fish from blackwater coastal-plain streams draining forests or wetlands in eastern and south-eastern part of the United States as well as from streams draining gold- or Hg-mined basins in the Western United States (Scudder et al. 2009). Clearly lower concentrations of Hg were found in fish living in 21 national parks in the Western United States, with average value at ~78 μg kg^–1 ww (Eagles-Smith et al. 2014). According to Depew et al. (2013b) Hg concentration in Canadian fish (gathered in years 1967–2010) averaged 370 μg kg^–1 ww (from below detection to 10,430 μg kg^–1 ww). In fish from years 1990–2010 estimated Hg_PREY ranged from 10 to 960 μg kg^–1 ww with a mean of 90 μg kg^–1 ww, decreasing westwards. This is consistent with spatio-temporal tendency in the United States of a decrease in Hg_PREY from east to west (Evers et al. 2007). This situation is closely related to the strong industrialization of the south-eastern regions of Canada and the Northeastern United States, where winds carry air masses anthropogenically contaminated with Hg. Mercury contamination is gradually deposited westwards, but the influence of mercury from bedrocks cannot be ruled out either (Page and Murphy 2005; Evers et al. 2007; Wentz et al. 2014).

The European Union as a whole lacks a common database on Hg concentration in fish that could be comparable to the North American one. In Scandinavian countries Hg concentrations in freshwater fish have been reported regularly since the late 1960s and early 1970s. The most data was collected for top predator northern pike followed by Eurasian perch Perca fluviatilis, which represents a lower trophic level (Munthe et al. 2007; Danielsson et al. 2011; Akerblom et al. 2014). Munthe et al. (2007) took into account all lacustrine data for Sweden, Norway and Finland from 1965 to 2004. In a standardized size of pike and perch, they found mean Hg concentrations of 730 and 400 μg kg^–1 ww, respectively, in the three Scandinavian countries. Importantly, mean value of Hg in “standard fish” (1 kg pike or 0.3 kg perch or 3.2 kg brown trout Salmo trutta or 1.4 kg Arctic char Salvelinus alpinus) was estimated to be as high as 690 μg kg^–1 ww (Munthe et al. 2007). The authors of that report stated that the data from Scandinavia show some similarity with data from a large survey in NE North America, when considering the mean values for various fish species. In the recent past in both regions, similar levels of atmospheric Hg pollution were noticed, and the geographic characteristics of bedrock and soils exhibit many analogies. For Eurasian perch and North American yellow perch, the mean concentrations were comparable: 400 versus 440 μg kg^–1 ww. However, an important difference was observed between these two regions, with Hg concentration in the pike from Scandinavia higher than in NE North America: 730 vs. 640 μg kg^–1 ww (Munthe et al. 2007). Miller et al. (2013) analysed data concerning Eurasian perch from Sweden and Finland covering the period 1974–2005. Swedish data from a later period (post-1996) show that in the fish from 22 and 72% lakes Hg concentrations were as high as >500 μg kg^–1 ww and between 200 and 500 μg kg^–1 ww, respectively. By contrast, after 1996 more lakes in Finland showed Hg concentrations in fish greater than 500 μg kg^–1 ww (31%), while fewer lakes had fish Hg concentrations below 500 μg kg^–1 ww (68%). Despite considerable reductions in Hg use and production as well as lower Hg atmospheric deposition in these countries, Miller et al. (2013) indicated that Hg concentrations in the fish exceeded the EQS/Hg (and EQS/Hg for the Nordic region was 200–250 μg kg^–1 ww). Moreover, in both Finland and Sweden, the perch from over 90% lakes exhibited Hg concentration exceeding 100 μg kg^–1 ww, which in North America is a level of concern for the protection of piscivorous mammals. One of the probable reasons for the persistence of elevated Hg concentrations in fish may be significantly lower selenium concentration in the Scandinavian environment (similar to Poland and eastern Germany). The deficiency of this element in the diet of vertebrates is accompanied by an increased accumulation of Hg, and in the case of fish from Scandinavia and Poland, this was indicated by Julshamn et al. (1986), Lindqvist et al. (1991), Hultberg (2002) and Kalisinska et al. (2017). In addition to the aforementioned species, bream Abramis brama is used to assess the quality of the aquatic environments in Europe, a common benthofagous species sampled in the German Environmental Specimen Bank (Wellmitz 2010). German biomonitoring research from the years 1994–2009 showed that on average Hg concentrations changed from ~100 to 350 μg kg^–1 ww and exceeded the EQS/Hg in all analysed years and all 17 sites from which breams came from: rivers Rhine, Danube, Saar, Elbe and its tributaries Saale and Mulde (Wellmitz 2010). Between 2007 and 2013, Hg levels were analysed in breams from five riverine places in France, Netherlands, Sweden and United Kingdom and one German lake as reference site (Nguetseng et al. 2015). Means of Hg concentration ranged from 18 to 246 μg kg^–1 ww. However, the EQS/Hg was exceeded in all years and at all riverine sites including the reference site except for the year 2012. The available data show that in Europe, the areas with not exceeded EQS/Hg in various fish species (even in non-piscivorous breams) are not very often reported; exceptions include some freshwater aquifers in Poland and Croatia (Zrncic et al. 2013; Szkoda et al. 2014).

In Asia, several year-long and systematic biomonitoring of Hg in freshwater fish has only been conducted in South Korea. In other countries occasional research has usually concerned individual species and reservoirs (Jin et al. 2006; Kim et al. 2012; Pal et al. 2012; Zhu et al. 2012). In 2006–2008 in Korea, analysis covered 55 species of wild freshwater fish, among which seven species predominated. The most numerous of them were two piscivores (largemouth bass Micropterus salmoides and Far Eastern catfish Silurus asotus) and five omnivores (steed barbell Hemibarbus labeo, Korean bullhead Pseudobagrus fulvidraco, pale chub Zacco platypus, crucian carp Carassius auratus, carp Cyprinus carpio). Each freshwater fish species was assigned to an appropriate trophic level (piscivore, carnivore, omnivore, planktivore). The piscivores had the highest median Hg concentration (148 μg kg^–1 ww) than carnivores and omnivores (83 μg kg^–1 ww and 68 μg kg^–1 ww, respectively). The median in planktivores was the lowest, at 30 μg kg^–1 ww. In most piscivorous species (including largemouth bass) from 12 sites Hg level exceeded 500 μg kg^–1 ww, which is recommended by the Korea Food and Drug Administration and the World Health Organization to protect human health (Kim et al. 2012).

The fish bioaccumulation factor (BAF), which expresses the ratio of THg (or MeHg) concentration in fish to the concentration in ambient water, depends on many factors including trophic position and fish size (US EPA 2000; Yu et al. 2011). BAF is mainly presented in a logarithmic form (log10), and in freshwater prey fish and larger predatory fish, it is usually in the range from 5.9 to 6.6 (Yu et al. 2011; Scudder Eikenberry et al. 2015; Wu 2017). In ecotoxicological studies, analysis of MeHg biomagnification is very important, including indicators of changes in concentration between different trophic levels (TMF, trophic magnification factor). Extensive analysis of Lavoie et al. (2013) shows that in freshwater food webs MeHg levels increase by a factor of 8.1 per trophic level. In addition, they stated that TMF is higher in lentic than lotic waters (7.6 vs. 9.8), and values of this factor increase from tropical via temperate to polar climatic zones (TMF, 3.9, 7.5 and 12, respectively). Finally, biomagnification factors (BMFs) are also estimated within food web, and the factor concerning MeHg (or THg) is expressed as the ratio of concentration in animal bodies to the concentration in their food (in ppm or ppb) (Rolfhus et al. 2011). BAF and BMF are very seldom presented in piscivorous inland birds and mammals, which participate in transport of Hg from aquatic to terrestrial environments. BAF and BMF calculated for piscivorous birds take into account Hg concentrations in water, fish and avian blood, feathers or eggs but rarely in soft tissues including muscle (Henny et al. 2009; Yu et al. 2011; Falkowska et al. 2013). In the transmission of Hg (especially MeHg) from freshwater fish to piscivorous wildlife of inland ecosystems muscle tissue plays important role for at least two reasons. Firstly, among soft digestible tissues, skeletal muscles represent the largest percentage of body weight, and secondly Hg present in them is almost all in MeHg form, which is easily absorbed. Therefore, it seems reasonable to analyse BMF using Hg concentrations in fish and wildlife muscle (not fish muscle and indigestible fur or feathers). For example (based on muscle tissue), in two pairs American mink—fish and Eurasian otter—fish from Western Poland BMFs were 27.3 and 10.9, respectively (Kalisinska et al. 2017). In addition, they found that these mammals quite often die on the roads and are later eaten by scavengers, thus contributing to the further transmission of Hg in the local terrestrial food web.

3.6 Mercury Concentrations in Soft Tissues in Various Groups of Inland Wildlife

The literature available in English includes many publications on the concentration of THg and much fewer investigate MeHg in soft and hard tissues of wild animals, especially in North America and Europe. In spite of this, there are basically no studies which estimate the average concentrations of THg representing the main ecotrophic groups. In order to characterize and compare the concentration of THg in wild terrestrial mammals and birds of the Northern Hemisphere, 140 studies from the years 1973–2017 were selected, including data on at least one of four soft tissues: liver, kidney, muscle and brain. The average concentrations of THg (mainly arithmetic means) from these reports concerned a minimum of three specimens of an individual species. If several groups of animals of the same species were included in the study (due to sex, age, temporal or territorial division), mean THg concentrations selected for the analysis referred to the largest number of individuals, preferring adults. Since the concentrations in soft tissues were given in mg kg^–1 in dry or wet weight, we made appropriate calculations, and the final results are presented in mg kg^–1 ww. Mammalian livers, kidneys, muscles and brains contain on average 70%, 75%, 75% and 80% of water, respectively, as calculated on the basis of several works (Weiner 1973, Blus and Henny 1990, Reinoso et al. 1997; Gamberg et al. 2005a, b; Sleeman et al. 2010; Kalisinska et al. 2012a, b). In the case of birds, it was assumed that their liver, kidney, muscle and brain contain 70%, 75%, 70% and 80% water, respectively (mean values were calculated based on the work of Cosson et al. 1988; Cosson 1989; Binkowski et al. 2013; Kalisinska et al. 2010, 2014a).

Data on mammals and birds were grouped according to their ecotrophic category. Among mammals, three groups were identified: Herb-M (predominantly herbivorous), Carn-M (terrestrial carnivores) and SemCarn-M (semiaquatic carnivores). Four groups were distinguished among the birds: TerrOmn-B (terrestrial omnivores and herbivores), TerrPred-B (diurnal and nocturnal predators), W-B (non-piscivorous waterfowl) and Pisc-B (piscivores). Groups, names of species and data sources are given in Table 17.3.

The analysis excluded cases indicating very high THg concentrations in the liver and kidneys, which were recorded in warm-blooded vertebrates living in areas heavily contaminated with mercury. For carnivorous mammals and those who prefer a different diet, excessive concentrations of mercury in liver and/or kidney were assumed to be above 16.5 and 12.5 mg kg^–1 ww, respectively, i.e. two-thirds and one-half of the value associated with mortality of mammals (lower value of the range: <25–30 mg kg^–1 ww), which was reported by Shore et al. (2011). The data of piscivorous birds and other ecotrophic groups that indicated very high Hg exposure were not included in this analysis. The threshold levels in the livers and kidneys in Pisc-B were over 2/3 of the levels, shown by Shore et al. (2011) to be associated with the death of non-marine birds (20 mg kg^–1 ww and >40 mg kg^–1 ww, respectively). Therefore, cases were excluded from statistical calculations when hepatic and nephric THg concentrations were higher than 13.2 and 26.4 mg kg^–1 ww. In relation to other ecotrophic bird groups, the exclusion limit was 1/2 of the levels indicated for avian liver and kidney by Shore et al. (2011), i.e. 10 and 20 mg kg^–1 ww.

Figure 17.2 shows the mean concentrations of THg in soft tissues of the various ecotrophic groups of birds and mammals inhabiting the inland areas in the Northern Hemisphere. Many species included in Table 17.3 occur in both Eurasia and North America. Some of them are native species on both continents (such as common loon, mallard, osprey, Eurasian elk/moose, reindeer/caribou), but some of them have been introduced, for example, fallow deer from Europe to North America and American mink and raccoon from North America to Europe (Genovesi et al. 2012; Bradley et al. 2014). Belonging to the same species and/or genus, occurrence on both continents and the large biological similarity (e.g. bald eagle and white-tailed eagle) are a justification for using their THg data in joint analyses.

Fig. 17.2

Medians of total mercury concentrations (ppm = mg kg^–1; ww, wet weight) in soft tissues of inland mammals (Herb–M, predominantly herbivore; Carn–M, terrestrial carnivore; SemCarn–M, semiaquatic carnivore) and birds (TerrOmn–B, terrestrial omnivore and herbivore; TerrPred–B, diurnal and nocturnal predators; W–B, non-piscivore waterfowl; Pisc–B, piscivore birds) from the Northern Hemisphere (for more details see Table 17.3)

Table 17.3 Analysed ecotrophic groups of mammals and birds, species names and source of data

3.6.1 Mercury Concentrations in Mammalian Soft Tissues

The low concentration of THg in the aboveground parts of plants (with the predominant share of InHg poorly absorbed in the gastrointestinal tracts of mammals and birds) results in a negligible exposure of most herbivorous animals to this toxic metal. In the three groups of mammals we distinguished above, THg concentrations were the smallest in Herb-M, but they can be arranged in the following ascending order: muscle < brain < liver < kidney (0.015, 0.026, 0.056 and 0.173 mg kg^–1 ww). According to Wisconsin Veterinary Diagnostic Laboratory (WVDL 2015), normal THg concentration in cervid kidney and liver does not exceed 0.1 mg kg^–1 ww, which only in the case of liver is consistent with the level established for the multispecies group of Herb-M. Two reports on special cases were excluded from this group, which indicated that increased concentrations of THg may be found even among herbivores. In the early 1990s among herbivorous ungulate mammals, there were exceptions such as roe deer from zones of a mercury mine in Idrija (Slovenia), which was active in the 1990s, and caribou from Canadian Arctic, Northwest Territories and Nunavut (Gnamuš and Horvat 1999; Gamberg et al. 2005b). In the roe deer, liver and kidney Hg levels were 0.64 and 15.56 mg kg^–1 ww. Hepatic and nephric tissues of the caribou contained 2.04 and 12.80 mg Hg kg^–1 ww, respectively. In both cases, the main reason for such a high concentration of Hg was the specific diet of these animals, containing large amounts of Hg. In the aboveground parts of plants from a smelter area of Idrija, average Hg was ~50 mg kg^–1 dw. Leaves of plants in those areas intensively absorbed Hg⁰ released during the roasting of ores containing this metal (Gnamuš and Horvat 1999). Caribou in the far north, on the other hand, mainly feeds on mosses and lichens, perennial plants which lack root systems and absorb contaminants (including Hg), along with their nutrients, from atmospheric deposition. In addition, high Hg levels in detoxification organs were related to the caribou weight loss in spring, resulting in lower absolute organ weights (Gamberg et al. 2005b).

The diet of carnivores (Carn-M) is very diverse. Their prey consists mostly of rodents, lagomorphs and birds, with the admixture of carrion, reptiles, frogs, insects, fruits and other parts of plants. In these predators, the average THg concentration in the liver and kidneys was similar and amounted to 0.105 mg kg^–1 ww and 0.140 mg kg^–1 ww, respectively. Farrar et al. (1994) argue that in the liver and kidneys of the dog, the concentration of THg usually does not exceed 0.1 mg kg^–1 ww, and in the WVDL list for canids from 2015 (including domestic dog), the normal Hg concentration in tissues is <0.1 mg kg^–1 ww and <0.200 mg kg^–1 ww, respectively. In both cases, these values do not differ from those calculated by us for the multispecies Carn-M group. In their muscles and brain, the THg concentration was an order of magnitude lower than in the liver and kidneys, and they did not exceed 0.018 mg kg^–1 ww and 0.030 mg kg^–1 ww, respectively (Fig. 17.2).

SemCarn-M group represents four species of the superfamily Musteloidea, including piscivorous mustelids. Among them, the diet of otters is 90% fish, American mink 60%, and raccoon 30% (Table 17.3; Kalisinska et al. 2017). This mammalian group is characterized by the largest body of data (especially with regard to the liver and kidney). Median THg concentration in the liver, kidney, muscle and brain of SemCarn-M were, respectively, 1.70, 1.09, 0.51 and 0.34 mg kg^–1 ww. Comparisons of median hepatic and nephric THg concentrations between Eurasian otter (liver n = 12, 2.57 mg kg^–1 ww; kidney n = 7, 1.30 mg kg^–1 ww) and river otter from North America (liver n = 25, 1.78 mg kg^–1 ww; kidney n = 11, 1.42 mg kg^–1 ww) showed no significant differences.

According to WVDL (2015), normal THg concentrations in the liver and kidneys of mustelids are <0.20–0.70 mg kg^–1 ww and <1.0 mg kg^–1 ww, respectively, much lower than our results. In North American studies from the 1980s, when Hg intoxication of otters and minks was much more frequent, background hepatic THg in those piscivorous species was indicated as <4–5 mg kg^–1 ww and ~2 mg kg^–1 ww, respectively (Wren 1986; O’Connor and Nielsen 1981; Carmichael and Baker 1989). In the light of the quoted papers from 1980s and our statistical analysis (taking into account European and North American reports from 1981 to 2017), it can be assumed that currently the values of hepatic background level for otters and American mink are <3.0 mg kg^–1 ww and <1.5 mg kg^–1 ww, respectively. According to our analysis, THg levels in the kidney, muscle and brain of the piscivorous mammals are <1.5, 1.0–1.3 and 0.3–0.6 mg kg^–1 ww, respectively.

Comparisons of hepatic THg concentration showed statistically confirmed differences between all three ecotrophic groups, and their values can be arranged in a decreasing series of SemCarn-M > Carn-M > Herb-M (1.700 > 0.105 > 0.015 mg kg^–1 ww). In relation to SemCarn-M, the concentrations of THg in the kidneys, muscles and brain of Carn-M were about an order of magnitude lower, and in the Herb-M groups, it was two orders of magnitude lower. No significant differences were found in kidney and brain THg between Carn-M and Herb-M (Fig. 17.2). In muscle, the concentrations of THg in Carn-M and Herb-M were more than 28 and 100 times lower than in SemCarn-M. In comparison to Carn-M, Herb-M had a 3.6 times lower level of muscle THg. In the muscles, similar to the liver, the concentrations could be arranged in a descending order (0.508 > 0.018 > 0.005 mg kg^–1 ww), and all intergroup differences were statistically significant.

3.6.2 Mercury Concentrations in Avian Soft Tissues

In birds, the lowest levels of THg in the liver, kidneys, muscles and brain occurred in the TerrOmn-B group, and their medians ranged from 0.024 to 0.067 mg kg^–1 ww. In some reports, especially in the case of muscles and the brain, the concentrations were very low, below the limit of detection, but increased levels (≥0.10 mg kg^–1 ww) were found in tissues of granivorous birds from Scandinavia in the 1970s, when large amounts of organic Hg fungicides were used in agriculture in that part of Europe (Holt et al. 1979).

Although in the liver and kidneys of waterfowl (W-B group) we found an order of magnitude higher THg concentration than in the TerrOmn-B group, the differences between these groups were not statistically significant. The largest number of differences were recorded between piscivorous birds (Pisc-B) and other analysed groups of birds. Pisc-B had the highest concentration of THg in the liver, kidneys, muscles and brain (3.21, 2.69, 0.78, 0.72 mg kg^–1 ww, respectively) and significantly differed in this regard from TerrOmn-B and W-B. Pisc-B and TerrPred-B showed no statistically confirmed difference in muscle THg (0.78 vs. 0.44 mg kg^–1 ww) and brain THg (0.72 vs. 0.35 mg kg^–1 ww).

In contrast to mammals, the WVDL list (2015) does not include the normal THg level for or different systematic groups of birds. Normal concentration is proposed of avian liver in the range of 0.01–0.10 mg kg^–1 ww and for the kidney at <0.02–0.30 mg kg^–1 ww. Puls (1988) suggested that normal concentrations of THg in the liver, kidney, muscle and brain of poultry were 0.01–0.10, 0.05–0.30, 0.008–0.100 and 0.10 mg kg^–1 ww, respectively. These levels coincide with those we calculated for TerrOmn–B, i.e. typical terrestrial birds, including galliformes. Other researchers, investigating various wild water birds, argue that hepatic and renal THg residues represent background concentrations when they are 0.3–3.0 mg kg^–1 ww (Ohlendorf 1993; Badzinski et al. 2009). This range includes median hepatic THg concentrations calculated by us for two groups of bird: W-B and TerrPred-B (Fig. 17.2). In the group of piscivorous birds (Pisc–B), hepatic THg exceeds 3.0 mg kg^–1 ww (3.21), but for the kidneys it is lower (2.69 mg kg^–1 ww).

Mammals and birds are characterized by different sensitivity to Hg, and depending on the type of intaken food, they accumulate different amounts of this toxic element. Significantly, the lowest adverse effect level (LOAEL) has not been established for most wild endothermic animals. In the common loon, in the case of the liver, kidney, breast muscle and brain, LOAEL values do not exceed 4.0, 2.3, 1.2 and 0.80 mg kg^–1 ww, respectively (Zhang et al. 2013). The quoted values coincide with the THg levels proposed by us for piscivorous birds, with the exception of muscle THg, which we estimated to be ~0.80 mg kg^–1 ww.

3.7 Mercury Concentrations in Hair and Feathers of Inland Wildlife

Hair (fur) and feathers are often used in ecotoxicological studies because they can be taken from living individuals. Here, the dominant form of Hg is MeHg, which reaches hair/fur/feathers in the period of their growth and reflects only that period. As in the case of soft tissues, the concentration of Hg in fur/feathers is closely related to the ecotrophic association of wildlife and Hg contamination of habitats. Sheffy and St Amant (1982) based on various furbearers from Wisconsin (USA) considered that Hg 1–5 mg kg^–1 ww (ppm dw) in hair to be normal background levels. In herbivorous mammals (such as American beaver, muskrat, white-tailed deer and lagomorphs), the average concentration of hair Hg does not exceed 0.3 ppm dw, and in many individuals it is below the limit of detection (Cumbie and Jenkins 1975; Sheffy and St Amant 1982; Stevens et al. 1997; Lourenco et al. 2011). In omnivorous mammals, such as common opossum Didelphis marsupialis, average hair Hg, depending on the environmental Hg pollution, ranged from 1.3 to 44 ppm dw (Cumbie and Jenkins 1975). Until recently, piscivorous mammals were thought to have the highest hair Hg concentrations among terrestrial mammals. The average hair Hg concentrations in these mammals from North America in the twentieth and twenty-first centuries usually exceeded 5, and sometimes 15 ppm dw (Sheffy and St Amant 1982; Stevens et al. 1997; Wolfe & Norman 1998; Mierle et al. 2000; Yates et al. 2005; Strom 2008). The maximum values in river otter and mink from the United States (Maine) reached 234 and 68.5 ppm dw, respectively (Yates et al. 2005). However, several years ago even greater concentrations were detected in the hair of insectivorous bats from Virginia (the South River, USA): little brown bat Myotis lucifugus and big brown bat Eptesicus fuscus, at 274 and 65.4 mg kg^–1 dw, respectively (Wada et al. 2010; Nam et al. 2012). Probably, the concentrations greater than 30 ppm dw in fur are associated with the clinical neurological effects, or they may be lethal (Wobeser and Swift 1976; Evers 2005; Basu et al. 2007), but there is little data on wild mammals in this respect.

In monitoring programs, feathers have low priority status for several reasons. Feather Hg concentration is characterized by high variability even in the same individual (depending, among others, on the type and location of feathers). Moreover, it relatively weakly correlates with the Hg concentration in soft tissues. Usually, the times of moulting and replacement of certain types of feathers are not known for most species, and it is even more complicated for migratory birds. In addition, the period of feather growth is accompanied by the redistribution of Hg in internal organs and its increased transport to feathers, both in chickens and older individuals (Honda et al. 1986; Eagles-Smith et al. 2008; Ackerman et al. 2011, 2016; Odsjo et al. 2012). In general, bird feathers have average Hg concentrations in the range of 0.1–5 ppm dw (Lodenius and Solonen 2013), but in some European herbivores, such as wood pigeon Columba palumbus and red-legged partridge Alectoris rufa, it may be <0.1 ppm (Hahn et al. 1993; Lourenco et al. 2011). Natural background levels of Hg in feathers of non-piscivorous raptorial birds are in the range 1–5 ppm dw (Scheuhammer 1991). Among adult piscivorous birds, it is estimated that this level for bald eagle in North America is much higher and in some regions ~20 ppm dw (DeSorbo et al. 2008). Among piscivorous birds the maximum concentration of Hg in feathers sometimes exceeds 190 ppm, for example, in osprey from Canada (DesGranges et al. 1998) and white-tailed eagle from Germany (Niecke et al. 1998). Mercury levels in feathers that are associated with adverse effects in birds are 5 ppm fresh weight or 7.5 mg kg^–1 dw. Concentrations of 15 ppm are required for adverse effects of mercury in some predatory birds (Burger and Gochfeld 2009). In raptorial birds concentrations >20 ppm may be connected with toxic effects, but in bald eagle it is probably >60 ppm (Scheuhammer 1991; DeSorbo et al. 2008).

Despite the large number of works with Hg concentration in mammalian fur and bird feathers, huge species and ecological diversity of wildlife make interpretation of results difficult, especially since the correlation between Hg concentration in these tissues and concentration in soft tissues in general are usually very weak or non-existent. Therefore, information on Hg obtained from fur and feather samples is not sufficient to clearly assess Hg intoxication of wildlife and their habitats.

4 Conclusions

Long-term studies of the abiotic environment, human toxicology and the ecotoxicology of Hg hold major gaps in knowledge on the behaviour of Hg in nature (including MeHg biomagnification) and subsequent long-term ignoring of the evidence of the negative effects of this metal on humans and other vertebrates. Maintaining the functioning of the various economic sectors based on Hg and coal-based energy has led to a dramatic increase in the environmental problems associated with Hg. Currently, the most important way to reduce anthropogenic Hg emissions and to reduce the health risks to humans and ecosystems globally is to act in international agreements. The first formal and very important preventive action was the signing of the Minamata Convention on Mercury in October 2013 (Kessler 2013; Larson 2014). However, its ratification, implementation and raising of awareness of entire societies and individuals will determine not only the health condition of this and future generations and the different environments and also the survival of many sensitive species, especially those directly or indirectly dependent on aquatic food chains. This requires, among other things, control of the presence of Hg in abiotic and biotic environments, including biomonitoring.