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Chapter 42
Mercury in Fish: History, Sources, Pathways,
Effects, and Indicator Usage
Edward J. Zillioux
Abstract Methyl mercury is highly toxic to humans, particularly to the developing
nervous system. Virtually all mercury in muscle tissue of naturally-occurring edible
fish is in the form of methyl mercury, and fish consumption is the most common
route of human exposure to methyl mercury. The monitoring of mercury in fish thus
provides reliable indication of potential exposure of humans to mercury, and
regulatory guidelines based on threshold levels of effects due to such exposure
provides the best mechanism for effective avoidance of mercury toxicosis in
populations throughout the world. This chapter traces the development of the use
of mercury in fish as an indicator of potential harm to human health from early
recognition of the dangers associated with methyl mercury, to the first records of
major toxicity events attributable to fish consumption, through the sources of
environmental contamination by mercury today, both natural and anthropogenic,
and an overview of the mercury species, environmental conditions and pathways
leading to uptake and bioconcentration of mercury in fish.
Keywords Mercury • Methyl mercury • Fish consumption • Bioconcentration •
Indicator • Regulatory guidelines
42.1 Introduction
The concentration of mercury (Hg) in edible fish tissue is today perhaps the most
broadly-applied indicator of potential harm to human health from any xenobiotic
substance. Organic mercury, in particular monomethyl-Hg (CH
3
Hg
+
or MeHg), is
the most toxic form of mercury commonly found in the environment, and con-
sumption of contaminated fish is the most common route of human exposure to
MeHg. Virtually all Hg (>95 %) in muscle tissue of naturally-occurring (and
commonly consumed by humans) fish is in the form of MeHg (Bloom 1992).
Today, fish and products derived from fish and sea mammals are virtually the
only sources of MeHg to humans (Clarkson 1997).
E.J. Zillioux (*)
Co-founder, ISEI, Environmental Indicators Foundation, LLC,
207-209 Orange Avenue, Suite G, Fort Pierce, FL 34950, USA
e-mail: zillioux@bioindicators.org
©Springer Science+Business Media Dordrecht 2015
R.H. Armon, O. Ha
¨nninen (eds.), Environmental Indicators,
DOI 10.1007/978-94-017-9499-2_42
743
This chapter reviews the early history of organic-Hg toxicity events, the origin of
our recognition of the value of fish as the primary indicator in determining poten-
tially harmful human exposure to MeHg, the primary pathways of uptake by fish,
bioaccumulation and bioconcentration of Hg in fish, factors that exacerbate or
mitigate the uptake of Hg in fish, the toxic effects of Hg to the fish themselves, as
well as to piscivorous species both wildlife and human, and how these translate into
regulatory standards and action levels, or consumption advisories. This chapter is
an overview of MeHg poisoning with a focus on the principal vector to humans and
wildlife. It is not intended to be a comprehensive review of the literature relating to
each subject, but it is my intent within each section to provide adequate references
to assist students who wish to pursue more focused studies in greater detail.
42.2 Historical Background
42.2.1 Organic Mercury Poisoning
The earliest known deaths attributed to exposure to organic mercury, involving
dimethyl mercury, occurred at St. Bartholomew’s Hospital in Smithfield, London
in the course of research on the valency of metals and metallic compounds. Details of
the research that led to the lethal exposures were reported by Frankland and Duppa
(1863); yet, inexplicably, their publication made no mention of the poisoning and
deaths of two technicians involved in the research. The two technicians were
apparently directly exposed to dimethyl Hg for periods of 3 months and 2 weeks,
respectively. According to hospital reports, both men exhibited symptoms associated
with ataxia and died 2 weeks and 12 months, respectively, after the onset of
symptoms. Clinical details were reported in two internal hospital reports (Edwards
1865,1866), which include the statement, “That the symptoms were due to the
inhalation of [mercuric methide] is rendered almost certain.” However, circulation
of these reports was limited; Hunter et al. (1940) commented that “The story of these
deaths has been handed down verbally from one generation of chemists to another.”
Despite these early fatalities, a detailed clinical description of the toxicity of
organic mercury to humans was not published in the scientific literature until
shortly before a massive poisoning event, traced to the consumption of contami-
nated fish, occurred in Minamata, Japan. Hunter et al. (1940) reported four cases of
human poisoning by inhalation of MeHg compounds that occurred in a factory
where fungicidal dusts were manufactured. In all four subjects, only the nervous
system was involved; symptoms included generalized ataxia, dysarthria (speech
slurred, slow, and difficult to understand), astereognosis (unable to distinguish form
of objects by touch), gross constriction of visual fields, inability to perform simple
tasks, weakness of arms and legs, and unsteadiness in gait. Symptoms known
to occur in cases of metallic Hg poisoning, salivation, stomatitis and erethism
(abnormal physical sensitivity), were generally absent. All recovered with varying
744 E.J. Zillioux
degrees of disability; the most severe, a 23-year-old man (Case 4), remained totally
disabled 3 years after the onset of symptoms.
Hunter et al. (1940) also undertook four experiments with animals, which
included a pathological study. The first three experiments exposed rats to methyl
mercury nitrate through gavage feeding or vapor inhalation. The fourth experiment
exposed a female monkey (Macacus rhesus) to MeHg vapor using the same box as
previous inhalation experiments with rats, albeit at a much lower dose in proportion
to body size. Symptoms in both exposed rats and monkey mimicked the general
ataxia, involving severe neurological symptoms, as observed in human exposures.
Neurological symptoms were far more severe in the monkey than in the rats,
suggesting that primates may be more susceptible to organic mercury compounds
than rats. Animals that survived through later stages of intoxication showed degen-
eration of the cells in the granular layer of the middle lobe of the cerebellum. This is
of particular interest because similar cerebellar cortical atrophy was found when the
first human exposure to MeHg (Case 4 above) came to necropsy, following the
exposure that occurred 15 years before his death (Hunter and Russell 1954).
42.2.2 Minamata Disease
Minamata Disease (MD) was first described by McAlpine and Araki (1958) as “an
unusual neurological disorder caused by contaminated fish,” which attacked vil-
lagers living near Minamata Bay in Kyushu Island, Japan between 1953 and 1956.
During this period, 40 families were affected, “causing death in more than a third of
its victims and serious disability in most of those who survived.” In addition,
numerous animals in the immediate area died with similar neurological symptoms,
including 24 cats, 5 pigs, 1 dog and many crows. The brains of 10 of the cats
showed the granular layer of the cerebellum especially affected. Although the cause
could not then be established, the authors noted that certain metals including MeHg
had been shown to cause some of the neurological symptoms of the disease. In all
cases, the disease was directly correlated with the consumption of fish caught in
Minamata Bay. It was strongly suspected that the fish were contaminated by
pollutants contained in the effluent from a chemical factory owned by Chisso &
Company, which, in 1950, had diverted its former open sea discharge through a
newly constructed channel discharging directly into Minamata Bay. The factory
utilized a process discovered in 1881 in which mercuric sulfate was used as a
catalyst in the conversion of acetylene to acetaldehyde (Clarkson 1997). MeHg
compounds were produced as byproducts of the catalytic process, which were at
first recycled but later discharged directly into Minamata Bay because of soaring
recycling costs (Kondo 1999). The causative agent of MD was verified in 1959 as
MeHg poisoning by a Kumamoto University team (Study Group of Minamata
Disease 1968). During the life of the plant, an estimated 600 tons of Hg were
discharged (Harada 1982). In 1965, a second outbreak of MD occurred far to the
north of Minamata Bay in the Agano River area of the Niigata Prefecture. Again,
42 Mercury in Fish: History, Sources, Pathways, Effects, and Indicator Usage 745
the cause of the poisoning was the production of acetaldehyde and discharge
of waste MeHg byproducts into the Agano River and the consumption of contam-
inated fish. In all, 2,920 cases of MD in the two areas were officially recognized
before the acetaldehyde process was discontinued in Japan and elsewhere
(Kondo 1999).
42.3 Sources, Speciation and Pathways of Hg in Fish
Hg in the global aquatic environment comes primarily from atmospheric deposition
(Livett 1988; Fitzgerald et al. 1991; Iverfeldt 1991), either direct from wet and dry
deposition or indirect through Hg deposition on watersheds or floodplains, which is
subsequently transported to surface water bodies. The form in which it deposits is
primarily as Hg
++
, or HgII, which can be biotransformed to MeHg which, in turn, is
efficiently taken up by organisms at the base of the food chain. Trophic transfer and
resultant concentrations in higher trophic level organisms are influenced by food
web dynamics, including length of the food chain. A portion of Hg entering the
environment from both natural and anthropogenic sources is reemitted as gaseous
elemental Hg, or Hg
0
, which is eventually redeposited, reemitted, etc. This cyclical
history must be considered when constructing source attribution budgets.
42.3.1 Natural Sources
The earth’s crust naturally contains approximately 50 ppb Hg, varying from an
average of 40 ppb in limestone to an average of about 160 ppb in the A soil horizon.
Most natural waters contain <2 ppb Hg (Adriano 1986). Natural sources of Hg to
the atmosphere include geological, vegetative and aquatic degassing, biomass
burning, and volcanic (explosive, passive & calderas) and geothermal emissions.
Oceanic and soil degassing are probably the most important contributions to
the global atmospheric burden of Hg (Pirrone et al. 2010; Norton et al. 1990).
Considerable uncertainty exists concerning the proportion of natural sources of Hg,
as opposed to anthropogenic sources, contributing to the total atmospheric burden.
Seigneur et al. (2003) reported the contributions of natural Hg emissions, direct
anthropogenic emissions, and re-emitted anthropogenic emissions to be roughly
equal. Thus, by these estimates, one-third of the total annual Hg emissions,
estimated at 6,000–6,600 metric tons, would be attributed to natural sources.
More recent models estimate the contribution from natural sources to be on the
order of 10 % of an estimated annual total of 5,500–8,900 metric tons currently
being emitted and re-emitted to the atmosphere from all sources (UNEP 2013).
746 E.J. Zillioux
42.3.2 Anthropogenic Sources
The earliest evidence of anthropogenic releases of Hg to the atmosphere is associated
with mining. Cinnabar (HgS) has been used for the production of vermillion since
about 1500 BCE, with early mining sites in China, Spain, Greece, Egypt, Peru, and
Mexico (Rapp 2009). On the Iberian Peninsula, mat deposits built by the seagrass
Posidonia oceanica provide a paleorecord of Hg fluxes to the marine environment
going back 4,315 years (Serrano et al. 2013). The first European Hg increase
attributable to an anthropogenic source was identified in the P. oceanica record
at about 2500 BP, coinciding with the beginning of intense mining in Spain.
Lake-sediment cores collected near Huancavelica, Peru demonstrate the existence
of a major Hg mining industry at Huancavelica spanning the past 3,500 years (Cooke
et al. 2009). Artisanal and small-scale gold mining (ASGM) is the largest source of
global anthropogenic Hg emissions today (e.g., Cleary 1990) followed closely by
coal combustion. Other large sources of emissions are non-ferrous metal production,
cement production, disposal of waste from mercury-containing products, hazardous
waste sites, and sewage treatment plants (UNEP 2013). The global distribution of
estimated Hg emissions in 2010 from anthropogenic sources, ranked from highest to
lowest regional emitters, is shown in Table 42.1.
Note that estimates of regional and total anthropogenic Hg emissions change
with pollution control technologies, regulatory limits and enforcement, fuel choice,
phase-out of Hg containing products, increased usage, etc. This is illustrated by
comparing global inventories over different time periods. For example, in 1995,
approximately 11 % of the total global anthropogenic emissions originated in North
America (Pacyna et al. 2003). In the 2010 inventory given in Table 42.1, the
estimated contribution from North America had decreased to 3 %, primarily due
to advances in emission control technologies, particularly with respect to coal
combustion. On the other hand, inventory data from South America show a clear
increase from approximately 3 % of total global anthropogenic Hg emissions in
1995, to 4 % in 2000, 7 % in 2005, and 12.5 % in 2010 (Pacyna et al. 2003,2006,
2010; UNEP 2013). This increasing trend, the largest global increase in Hg
emissions over the 15-year period of record, is due almost entirely to ASGM
(Cleary 1990). Indeed, as noted by Pacyna et al. (2010), “at least 100 million people
in over 55 countries depend on ASGM – directly or indirectly – for their livelihood,
mainly in Africa, Asia and South America.”
42.3.3 Atmospheric Hg Speciation and Deposition
Mercury is emitted to the atmosphere in gaseous forms, as Hg
0
and HgII
(also known as reactive gaseous Hg, or RGM) and as particulate Hg, or Hg
p
.
The majority of Hg emissions to the atmosphere is as Hg
0
, including soil, vegeta-
tive, and oceanic degassing, volcanic and geothermal emissions, mining operations,
biomass burning and approximately half of fossil fuel emissions (Pacyna
et al. 2006). The atmospheric residence time for Hg
0
is approximately one year
42 Mercury in Fish: History, Sources, Pathways, Effects, and Indicator Usage 747
enabling distribution on a global scale. The majority of Hg
0
is eventually oxidized
to HgII, which is soluble and subject to washout. HgII is also emitted directly to the
atmosphere from various industrial processes including fossil fuel combustion
(primarily coal), municipal waste incineration, cement production, as well as
crematoria. HgII and Hg
p
, have much shorter atmospheric residence times, often
depositing on a local or, at most, a regional scale from point sources. Other species
of Hg are generally present at de minimis levels in the atmosphere and will not be
considered further here.
42.3.4 Methylation and Uptake of Hg in Fish
Deposited HgII is the primary substrate for methylation by sulphate- and
iron-reducing bacteria and/or methanogenic archaea under anoxic conditions
found in sediments, as well as in periphyton and wetland catchment areas, and is
Table 42.1 Mercury emissions from various regions, in tones per year, with the range of the
estimate, the percentage of total global anthropogenic emissions, and the primary and secondary
regional sources of emissions
a,b
Region
Emissions
(range), tones %
Primary and secondary regional
sources
East and Southeast Asia 777 (395–1,690) 39.7 1
○
Coal combustion; 2
○
ASGM
Sub-Saharan Africa 316 (168–514) 16.1 1
○
ASGM; 2
○
Coal combustion
South America 245 (128–465) 12.5 1
○
ASGM; 2
○
Non-ferrous
metals
South Asia 154 (78.2–358) 7.9 1
○
Coal combustion; 2
○
Large-
scale gold
CIS and other Eastern
European countries
115 (42.6–289) 5.9 1
○
Coal combustion
2
○
Non-ferrous metals
European Union (EU27) 87.5 (44.5–226) 4.5 1
○
Coal combustion; 2
○
Cement
production
Undefined 82.5 (70–95) 4.2 Global total from contaminated
sites
North America 60.7 (34.3–139) 3.1 1
○
Coal combustion; 2
○
Product
waste
Central America and
Caribbean
47.2 (19.7–97.4) 2.4 1
○
ASGM; 2
○
Non-ferrous
metals
Middle Eastern States 37.0 (16.1–106) 1.9 1
○
Coal combustion; 2
○
Cement
production
Australia, NZ and Oceania 22.3 (5.4–52.7) 1.1 1
○
Large-scale gold; 2
○
Non-ferrous metals
North Africa 13.6 (4.8–41.2) 0.7 1
○
Non-ferrous metals; 2
○
Product waste
Grand Total 1960 (1,010–
4,070)
100
a
UNEP (2013)
b
Estimates based on 2010 inventory
748 E.J. Zillioux
highest in sediments moderately enriched by organics and sulfate (Poulain and
Barkay 2013; Hamelin et al. 2011; Gilmour et al. 1992; Driscoll et al. 1994;
Sunderland et al. 2006 [see also reviews by Zillioux et al. 1993; Porcella 1994
and references therein]). The efficiency of MeHg production varies greatly among
species and between geobiological niches, however. Benoit et al. (2003), in an
extensive review of MeHg production and degradation, made the case that sulfate-
reducing bacteria (SRB) are the key Hg methylators in aquatic ecosystems. They
cited studies using specific metabolic inhibitors where inhibition of methanogens
increased Hg methylation, while inhibition of sulfate reduction dramatically
decreased MeHg production in saltmarsh sediment (Compeau and Bartha 1985).
In addition, Oremland et al. (1991), citing Mcbride and Edwards (1977), reported
that “Hg methylation was not detected in whole cells of methanogens or in
methanogenic sewage sludge suggesting that methanogens are not active in this
reaction.” However, Hamelin et al. (2011) presented findings that suggest “that
methanogens rather than SRB were likely the primary methylators in the periphyton
of a temperate fluvial lake.” Parks et al. (2013), although acknowledging that SRB
are the main producers of MeHg in nature, provided genetic evidence for “a
common mercury methylation pathway in all methylating bacteria and archaea.”
Kerin et al. (2006), in a paper relating mercury methylation to dissimilatory iron-
reducing bacteria (DIRB), implied that, since current models for methylation are
based on relationships between methylation and sulfate reduction, the potential
significance of methylation by iron reduction in certain environments may be
undervalued or missed entirely. Kerin concluded that “the finding that DIRB can
produce MeHg suggests that Hg methylation may be important in sediments and
soils where these organisms are dominant, e.g., iron-rich sediments with low
concentrations of sulfate.” Regardless of the methanogenic species, MeHg pro-
duced in aquatic environments is taken up rapidly by the food web, with greater
accumulation in higher trophic levels. Given that some methanogenic bacteria and
archaea are among the oldest life forms on the planet, and that a shared evolutionary
history for methanogenesis and sulfate reduction developing about 3.5 billion years
ago has been postulated (Susanti and Mukhopadhyay 2012), and that inorganic Hg
has always been present in Earth’s biosphere, it seems that fish have accumulated
MeHg throughout their evolutionary history (Clarkson 1997).
Calculations in dilute-water lakes from the ratio of total fish Hg to total Hg and
aqueous MeHg measurements indicate accumulation of MeHg in fish by a factor of
three million times, accounting for the observation that fish can contain more than
one part per million Hg in water with less than one part per trillion of total Hg
(Zillioux et al. 1993). Although accumulation of Hg in fish can occur through
uptake across both the gills and the gut, dietary uptake seems to account for more
than 90 % of total MeHg uptake with assimilation rates up to 80 % or higher. MeHg
binds to red blood cells and distributes via the circulatory system to all organs and
tissues, although much relocates to the skeletal muscle where it accumulates bound
to sulfhydryl groups in protein (Wiener et al. 2003). This process is described by a
bioaccumulation factor (BAF), i.e., the ratio of tissue chemical residue to chemical
concentration in an external environmental phase (water, sediment, or food).
42 Mercury in Fish: History, Sources, Pathways, Effects, and Indicator Usage 749
For equilibrium partitioning at steady state, the BAF may approximate the
organism-water partition coefficient (K
b
), although this varies with the degree of
uptake through the dietary route (the bioconcentration factor [BCF] is equivalent to
K
b
since it describes the ratio of tissue chemical residue directly to chemical
concentration in water with no food-web exposure). The bioaccumulation process
results in a biomagnification of Hg, or increase in tissue chemical residues at higher
trophic levels, primarily as a result of dietary accumulation (Spacie et al. 1995),
although the degree of biomagnification in a given water body varies by species and
with size and age. Figure 42.1 illustrates the range of fish species variations in
average Hg tissue (primarily axial muscle) concentrations as reported by the
U.S. Food and Drug Administration (USFDA) and the U.S. Environmental Protec-
tion Agency (USEPA).
42.3.4.1 Factors Affecting Methylation and Uptake of MeHg by Fish
Since methylation of HgII is a prerequisite for the efficient uptake of Hg in fish
in most natural water bodies, an examination of the environmental factors that
promote methylation, as well as demethylation, is important to understand the
observed differences in Hg uptake between water bodies. Methylation and demeth-
ylation should be viewed within the context of the overall cycling of mercury
species in the aquatic environment. The major compartments, fluxes, and reaction
components of mercury in a lacustrine ecosystem are illustrated in Fig. 42.2.
Many authors have considered the influence of water chemistry on the uptake of
Hg in fish. For example, Lange et al. (1993) reported that uptake of MeHg in
largemouth bass in 53 Florida lakes was shown to be positively correlated with fish
age (strongest correlation) and fish size (e.g., see Table 42.2), and negatively
correlated with alkalinity, calcium, chlorophyll ɑ, conductance, magnesium, pH,
total hardness, total nitrogen, and total phosphorus. They found that pH accounted
for 41 % of the variation in Hg concentration for standardized age three fish, while
chlorophyll ɑand alkalinity accounted for 45 % of the variation. Fish Hg concen-
trations were significantly higher in lakes with either pH <7, alkalinity <20 mg/L
as CaCO3, or chlorophyll ɑ<10 μg/L. Also, Hickey et al. (2005) studied the effects
of water chemistry on Hg in 747 fish of mixed species from 31 Ontario lakes and
11 lakes in Nova Scotia, Canada. They found that pH alone explained 77.7 % of the
variation in Hg concentration in fish, while MeHg in water and dissolved organic
carbon (DOC) accounted for only 2.7 % of the variation. They concluded that
“reducing acid rain and mitigation of pH levels will reduce Hg levels more than will
reducing Hg deposition.” Although not directly addressing this relationship, it
should be noted that, in a whole-ecosystem experiment where different isotopes
of Hg were deposited directly to a Canadian lake surface and to upland and wetland
components of its watershed and tracked over time, the Hg levels in fish responded
rapidly and directly to the changes in atmospheric deposition (spike additions)
when added directly to the lake surface (Harris et al. 2007; Engstrom 2007).
750 E.J. Zillioux
Fig. 42.1 Species variations in mercury content (ppm) (Source: John Blanchard (Sources: FDA and EPA), Sierra Magazine, Nov/Dec 2011)
42 Mercury in Fish: History, Sources, Pathways, Effects, and Indicator Usage 751
A number of studies have developed statistical or inferential models in order to
determine the biogeochemical factors that govern Hg bioaccumulation in aquatic
food webs. For example, Pollman (2012) used a variety of multivariate modeling
techniques to construct and validate empirical models relating the occurrence of Hg
in fish to chemical and other potential determinants of the variability of fish tissue
Hg concentrations. The modeling effort used data sets representative of over 7,700
lakes greater than 4 ha and 83,457 km of stream and riverine reaches in the State of
Florida, and approaches including integrating principal components analysis with
multiple linear regression and generalized linear modeling for the lake model, and
classification and regression tree analysis for the streams and rivers model. The
sequence of importance of independent variable contributions to the overall vari-
ability in Hg in largemouth bass was: for the study lakes, alkalinity >chlorophyll
Table 42.2 The effect of size (age) on the mean Hg level in 181 king mackerel sampled in 1999
from North Carolina, South Carolina, Georgia, and Florida, USA
Size category (fork length) (in.) Number of fish Average (ppm) Range (ppm)
<27 19 0.22 0.14–0.36
27–32 43 0.34 0.15–1.00
33–39 53 0.80 0.25–2.10
>39 66 1.54 0.40–3.50
Moore (2000)
Fig. 42.2 Mercury cycling in a lake and its watershed (From: Engstrom (2007) (Reprinted with
permission))
752 E.J. Zillioux
ɑ>urban runoff disturbance >atmospheric deposition >sulfate; and, for the study
streams and rivers, pH DO % saturation >conductivity >total Kjeldahl
nitrogen >sulfate >total phosphorous. Considering uncertainties in model
prediction and inferred distributions, the model results for the 90th percentile
concentrations for largemouth bass Hg, in mg/kg, were: streams – 1.295; small
lakes – 1.319; rivers – 1.136; the Everglades – 1.071; and large lakes – 0.694. The
much lower predicted fish Hg concentration in large lakes reflects both higher
alkalinities and higher productivity compared to small lakes.
Chemical and biological control of microbial methylation and demethylation of
Hg is complex and not fully understood. The degree of complexity is perhaps best
illustrated by the central role played in the biogeochemical cycling of Hg by
interactions affecting Hg methylation/demethylation among dissolved organic
matter (DOM), sulfate reduction, and sulfide inhibition (e.g., Benoit et al. 2003;
Hickey et al. 2005; Miller et al. 2007; Graham et al. 2012,2013). Sulfate and sulfide
exert conflicting influences on the extent of Hg methylation such that the highest
methylation rates are found at sites with intermediate sulfate-reduction rates and
sulfide concentration, although the point at which the highest rates occur varies with
other controlling factors. Sulfate additions increase Hg methylation rates until
sulfate concentration reaches the point where sulfide buildup is sufficient to inhibit
microbial methylation (Gilmour et al. 1998; Graham et al. 2013; Benoit et al. 2003).
Correlations between DOM and MeHg production are positive in many aquatic
sediments and wetland soils with low μM sulfide levels, and DOM concentrations
below 8 mg C/L. DOM can strongly enhance the bioavailability of HgII to SRB
under micromolar sulfide concentrations and anoxic conditions (Graham
et al. 2012,2013); however, the degree of enhancement is influenced by DOM
size, hydrophobicity, and sulfur content. The interactions of Hg with DOM in the
presence of sulfide complicate the Hg-sulfide complexation as predicted by ther-
modynamic models such that laboratory and field studies have not always been in
agreement. DOM influences numerous processes in the biogeochemical cycling of
Hg including HgII complexation and transport, MeHg complexation, transport,
precipitation and dissolution of Hg-S minerals, and MeHg production by microor-
ganisms (Graham et al. 2013 and references therein). In addition, Hg complexed
with DOM dominates the speciation of Hg under oxygenated conditions and may
influence the ultimate Hg substrate available to SRB at the primary site of methyl-
ation in aquatic sediments, just below the oxic/anoxic interface. Reported correla-
tions between DOM and MeHg concentration also can be negative (e.g., Hickey
et al. 2005; Driscoll et al. 1995), further reflecting the biogeochemical complexity
controlling these interactions.
Other factors affecting MeHg formation and uptake of Hg in fish have been
reported by many authors. Examples include: food-chain structure (Cabana
et al. 1994, Greenfield et al. 2001); salinity (Compeau and Bartha 1987; Farmer
et al. 2010); selenium (Southworth et al. 1999; Belzile et al. 2006; Peterson
et al. 2009); acid rain (Richardson and Currie 1995;Richardsonetal.1995a,b;
physical attributes of lakes (Richardson 1994); sulfate loading (Gilmour et al. 1998);
algal blooms (Pickhardt et al. 2002); temperature and season (Benoit et al. 2003).
42 Mercury in Fish: History, Sources, Pathways, Effects, and Indicator Usage 753
As mentioned above, demethylation occurs in natural aquatic systems in concert
with methylation such that MeHg uptake by fish is a function of the net microbial
production of MeHg. Bacterial demethylation through the mer operon pathway has
been well characterized (e.g., Robinson and Tuovinen 1984; Liebert et al. 1999;
Hobman et al. 2000; Barkay 2000). The mer operon contains the organomercurial
lyase gene that cleaves the carbon-Hg bond of MeHg, producing methane and HgII;
the HgII then is reduced to Hg
o
through a second step involving the Hg-reductase
enzyme (Benoit et al. 2003; Wiener et al. 2003). Oremland et al. (1991) described
an oxidative demethylation process that derives energy from single carbon sub-
strates in a wide range of environments including freshwater, estuarine, and
alkaline-hypersaline sediments and in both aerobic and anaerobic conditions.
Working in three environments that differ in the extent and type of Hg contamina-
tion and sediment biogeochemistry, Dipasquale et al. (2000) found that severely
contaminated sediments tend to have microbial populations that actively degrade
MeHg through mer-detoxification, whereas oxidative demethylation occurs in
heavily contaminated sediments as well but appears to dominate in those less
contaminated, under both aerobic and anaerobic conditions.
42.4 Effects
42.4.1 Effects of Hg on Fish
The effects of Hg in fish, as well as in other aquatic organisms, and piscivorous
wildlife have been reviewed extensively (e.g., Eisler 1987). In two early reviews
published in 1979 (Taylor 1979; Birge et al. 1979) a total of 50 discrete references that
specifically addressed the issue of Hg in fish were cited. Since the completion of these
two reviews, at least 447,000 publications have dealt with some aspect of Hg in fish
(source: Google Scholar, extrapolated from a sample size of 1,000 citations).
Although diet is the primary route of Hg uptake in fish, most laboratory studies
of Hg in fish have measured effects through gill uptake from concentrations in
water much higher than typically observed in natural water bodies, where typical
concentrations in lakes are measured in the low ng/L range (Watras et al. 1992). For
example, Zillioux et al. (1993), in a review on the effects of Hg in wetland
ecosystems, reported effects of organic Hg on fish derived from laboratory expo-
sures at concentrations in water from 0.1 μg/L (zebrafish [Brachydanio rerio]–
hatching success reduced) to 0.88 μg/L (brook trout [Salvelinus fontinalis embryo]
– enzyme disruption). Sublethal exposures of fish to MeHg can result in impaired
ability to locate, capture, and ingest prey, and to avoid predation (Kania and O’Hara
1974; Little and Finger 1990; Sandheinrich and Atchison 1990; Weis and Weis
1995; Fjeld et al. 1998; Samson et al. 2001, as cited in a comprehensive review by
Wiener et al. 2003). However, Wiener and Spry (1996) in a review on Hg in
freshwater fish concluded that reduced reproductive success was the most plausible
754 E.J. Zillioux
toxicological endpoint in wild fish populations exposed to Hg-contaminated food
webs. For example, Hammerschmidt et al. (2002) reported that exposure of fathead
minnows (Pimephales promelas) to three concentrations of dietary MeHg of 0.88,
4.11, and 8.46 μgHgg
1
dry weight prior to sexual maturity, resulted in reduced
spawning success rates of 63 %, 40 %, and 14 %, respectively, down from success
rates of 75 % for controls. Beckvar et al. (2005) linked fish tissue residues of Hg to
biological effects thresholds, primarily of growth, reproduction, development, and
behavior, using literature sources screened for data consistency. Based on an
evaluation of several approaches, the threshold-effect level (t-TEL) best
represented the underlying data. (The t-TEL is calculated as the geometric mean
of the 15th percentile concentration in the effects data set and the 50th percentile
concentration in the no-effects data set.) They concluded that a whole-body t-TEL
of 0.2 mg Hg/kg wet weight of tissue would be protective of juvenile and adult fish,
where the incidence of effects below the t-TEL is predicted to be rare.
42.4.2 Effects of Hg on Piscivorous Wildlife
Effects of Hg on piscivorous birds and mammals were reviewed by Wolfe
et al. (1998), with emphasis on the mechanisms of Hg toxicity and interpretation
of residue data. In both birds and mammals, MeHg readily penetrates the blood-
brain barrier producing brain lesions, spinal cord degeneration, and central nervous
system dysfunctions. A residue threshold for toxicity in mink is suggested at
5.0 ppm for brain and muscle tissue. From their review of the literature, Zillioux
et al. (1993) concluded that residue thresholds for significant toxic effects in wading
birds occur between 1 and 3.6 ppm wet weight (w/w) in eggs and 5 ppm w/w
in liver. However, a study by Frederick and Jayasena (2011) suggested that
dose-related increases in male-male bonding and altered sexual display behavior
in the white ibis occur at mean residue levels as low as 4.3 ppm fresh weight in
feathers (approx. equivalent to 0.37 ppm in wading bird eggs, from comparative
feather/egg effects data in Zillioux et al. 1993) and 0.73 ppm in blood. Many
investigations on ecosystem proliferation of Hg and the effects of Hg on piscivo-
rous wildlife have been conducted in the Florida Everglades, the largest freshwater
wetland in the continental United States. For example, Frederick et al. (1999)
studied the diet of great egret (Ardea albus) nestlings exposed to dietary Hg during
the breeding seasons of 1993–1996. By collecting and analyzing Hg in regurgitated
food samples from large colonies throughout the central Everglades, where fish
comprised >95 % of the nestlings’diet, Frederick et al. estimated that nestlings
would ingest 4.32 mg total Hg (Hg
T
) during an 80-day nesting period. In live tree
islands, which are the primary habitat for wading bird colonies in the Everglades,
the annual Hg deposition by bird guano was estimated at 148 μgm
2
year
1
, about
eight times the atmospheric deposition of Hg in southern Florida (Zhu et al. 2014).
Feather mercury concentrations in adults and nestlings of the great egret exceeded
30 ppm in environmental samples from the Florida Everglades in the early 1990s,
42 Mercury in Fish: History, Sources, Pathways, Effects, and Indicator Usage 755
when this area had the highest levels of Hg in fish in the entire USA (as high as
2.7 ppm in axial muscle tissue of largemouth bass, Wolfe et al. 2007).
During the same period, top predators of the fish-based food chain in the Florida
Everglades also had high tissue Hg levels. Alligators (Alligator mississippiensis)
collected on a transect through the Florida Everglades in 1999 were reported by
Rumbold et al. (2002) with Hg
T
mean concentrations (n ¼28) in liver and tail
muscle of 10.4 and 1.2 ppm w/w, respectively. A single Florida panther (Puma
concolor cori), a critically endangered species in Florida, was found dead in the
southern Everglades region with the highest Hg concentration ever reported of
110 ppm w/w in the liver; Hg toxicosis was strongly implicated in its death (Roelke
et al. 1991). Other free-ranging panthers in the same region had mean hair, liver,
and muscle concentrations of 56.4, 40.6 and 4.4 ppm Hg
T
w/w, respectively. Roelke
et al. concluded that Hg
T
in panther hair greater than 57.3 ppm fresh weight would
indicate toxicosis, and identified an “at risk” threshold value for Hg
T
in panther hair
as greater than 12.57 ppm. All of these panthers were known to be feeding on
Hg-contaminated raccoons (Procyon lotor). Raccoons are opportunistic omnivores,
but eat largely insects and crustaceans and some fish outside berry season, which
peaks in January in the Everglades region. As is the case in fish, Hg in insects is
essentially all MeHg (Mason et al. 2000). Roelke et al. (1991) reported a mean
value of 1.8 1.24 ppm Hg in raccoon muscle tissue in the central Everglades,
while in a retrospective study across all of southern Florida, Porcella et al. (2004)
found no statistical difference in raccoon Hg content over the past 50 years.
42.4.3 Effects of Hg in Humans
About 95 % of MeHg in fish ingested by humans is absorbed. In the blood, about
90 % is associated with red cells, probably bound to the sulfhydryl (SH) groups of
hemoglobin. From the bloodstream, it is taken up by all tissues, and readily crosses
the blood-brain and placental barriers. Early studies of the effects of MeHg on
humans have been described above (Section 1.2.1). More recent studies have
confirmed that the major human effects from exposure to MeHg are neurotoxicity
in adults and toxicity in fetuses of mothers exposed during pregnancy. The cortex of
the cerebrum and cerebellum are selectively involved in Hg toxicosis, with focal
necrosis on neurons, lysis and phagocytosis and replacement by supporting glial
cells. The over-all acute effect is cerebral edema, but with prolonged destruction of
gray matter and subsequent gliosis, resulting in cerebral atrophy (see reviews by
Clarkson 1997 and Goyer and Clarkson 2001, and references therein). However, the
primary human health concern today is with more subtle effects arising from
prenatal exposure, such as delayed development and cognitive changes in children.
Myers et al. (2003) studied neurodevelopmental effects in a fish-consuming popu-
lation in the Republic of Seychelles, investigating 779 mother-infant pairs. Mothers
averaged 12 fish meals per week, with fish concentrations of MeHg similar to
commercial ocean fish elsewhere. Children were followed from the prenatal period
756 E.J. Zillioux
(mean prenatal MeHg exposure was 6.9 ppm, SD 4–5 ppm) to age 9 years.
Neurocognitive, language, memory, motor, perceptual-motor, and behavioral func-
tions were assessed at 9 years. Their data did not support the hypothesis that there is
a neurodevelopmental risk from prenatal MeHg exposure resulting solely from
ocean fish consumption. However, other studies of prenatal exposure related to
fish consumption have shown effects in children, from an inverse correlation
between maternal Hg hair levels and IQ in their children (Kjellstro
¨m et al. 1989)
to cognitive developmental delays at the age of 4 years (Freire et al. 2010). A WHO
Expert Group concluded that there may be a low risk of prenatal poisoning at
maternal hair levels between 10 and 20 ppm (corresponding to blood levels of
20–40 ppb). Two independent analyses of the same data base concluded that the
lowest effect level may be anywhere from 7 to over 100 ppm in maternal hair. As a
point of comparison, a study conducted in the Florida Everglades, during the period
of highest reported concentrations of Hg in fish, measured Hg in the hair of sport
fishermen, Everglades residents, and subsistence fishermen. Of 350 participants,
119 had levels above detection limits and, of these, the mean total Hg in hair was
3.62 (SD 3.0) ppm, with a range of 2.28–15.57 ppm (Fleming et al. 1995).
42.5 Use of Fish as Indicators of Human Hg Exposure
The practice of using fish as indicators of chemical exposure is relatively new.
A permissible Hg content of 0.5 ppm in fish established in 1970 by the U.S. Food
and Drug Administration (USFDA) was the first regulatory action level for any
element in the USA (Hall et al. 1978). This temporary action level was later revised
upward to 1 ppm MeHg in fish, which “was established to limit consumers’MeHg
exposure to levels 10 times lower than the lowest levels associated with adverse
effects (paresthesia)” (USFDA 1995). This new action level was based on the
occurrence of adverse effects in adults “because the level of exposure was actually
lower than the lowest level found to affect fetuses, affording them greater protec-
tion.” Nevertheless, in January 2001 the U.S. Environmental Protection Agency
(USEPA) in apparent contradiction to the USFDA action, established a water
quality criterion of 0.3 mg MeHg/kg fish tissue screening value for fish consump-
tion (USEPA 2010). This was the USEPA’s first issuance of a water quality
criterion expressed as a fish tissue value rather than as an ambient water column
value. The more restrictive USEPA criterion is intended to be protective of recre-
ational, tribal, ethnic, and subsistence fishers who typically consume fish and
shellfish from the same local water bodies repeatedly over many years. Today,
action levels for fish consumption advisories are common throughout the world.
Table 42.3 provides the most complete compendium of these action levels available
for 53 nation states, including the 27 member states of the European Union and
12 member states of the Commonwealth of Independent States as well as general
guidelines issued by the World Health Organization/Food and Agriculture
Organization of the United Nations.
42 Mercury in Fish: History, Sources, Pathways, Effects, and Indicator Usage 757
Table 42.3 Examples of maximum allowed or recommended levels of Hg in fish in various
countries and by WHO/FAO (based on submissions to UNEP, unless otherwise noted)
Country/
organization Fish type
Maximum
allowed/
recommend
levels in
fish
a
Type of measure
Tolerable intake
levels
a
Australia Fish known to
contain high levels
of mercury, such as
swordfish, southern
bluefin tuna,
barramundi, ling,
orange roughy, rays,
shark
1.0 mg
Hg/kg
The Australian
Food Standards
Code
Tolerable Weekly
Intake: 2.8 μg
Hg/kg body weight
per week for
pregnant women.
All other species of
fish and crustaceans
and molluscs
0.5 mg
Hg/kg
Canada All fish except
shark, swordfish or
fresh or frozen tuna
(expressed as total
mercury in the
edible portion of
fish)
0.5 ppm
total Hg
Guidelines/
Tolerances of
Various Chemical
Contaminants in
Canada
Provisional
Tolerable Daily
Intake: 0.47 μg
Hg/kg body weight
per day for most of
the population and
0.2 μg Hg/kg body
weight per day for
women of
child-bearing age
and young children
Maximum
allowable limit for
those who consume
large amounts of
fish, such as
Aboriginal people
0.2 ppm
total Hg
China Freshwater fish 0.30 mg/kg Sanitation
standards for food
Croatia Fresh fish Predatory
fish (tuna, sword-
fish, molluscs,
crustaceans)
1.0 mg
Hg/kg
0.8 mg
methyl
Hg/kg
Rules on quantities
of pesticides,
toxins, mycotoxins,
metals and
histamines and
similar substances
that can be found in
the food.
All other species of
fish
0.5 mg
Hg/kg
0.4 mg
methyl
Hg/kg
Canned fish (tin
package) Predatory
fish (tuna, sword-
fish, molluscs,
crustaceans)
1.5 mg
Hg/kg
1.0 mg
methyl
Hg/kg
All other species of
fish
0.8 mg
Hg/kg
0.5 mg
methyl
Hg/kg
(continued)
758 E.J. Zillioux
Table 42.3 (continued)
Country/
organization Fish type
Maximum
allowed/
recommend
levels in
fish
a
Type of measure
Tolerable intake
levels
a
European
Community
Fishery products,
with the exception
of those listed
below.
0.5 mg
Hg/kg wet
weight
Various Commis-
sion regulations
European
Commission,
Official Journal of
the European
Communities
7 February 2002
Anglerfish, Atlantic
catfish, bass, blue
ling, bonito, eel,
emperor or orange
roughy, grenadier,
halibut, marlin,
pike, plain bonito,
Portuguese dogfish,
rays, redfish, sail
fish, scabbard fish,
shark (all species),
snake mackerel or
butterfish, sturgeon,
swordfish and tuna.
1 mg Hg/kg
wet weight
Commission
regulation
(EC) No.
221/2002
Georgia Fish (freshwater)
and fishery products
0.3 mg
Hg/kg
Georgian Food
Quality Standards
2001
Fish (Black Sea) 0.5 mg
Hg/kg
Caviar 0.2 mg
Hg/kg
India Fish 0.5 ppm
total Hg
Tolerance
Guidelines
Japan Fish 0.4 ppm
total Hg/kg
Food Sanitation
Law – Provisional
regulatory standard
for fish and
shellfish
Provisional
Tolerable Weekly
Intake: 0.17 mg
methyl Hg
(0.4 μg/kg body
weight per day)
(Nakagawa
et al. 1997).
0.3 ppm
methyl Hg
(as a
reference)
Korea,
Republic of
Fish 0.5 mg
Hg/kg
Food Act 2000
Mauritius Fish 1 ppm Hg Food Act 2000
New Zealand Fish 1.6 μg
MeHg/kg
body weight
per week
Food Standards
Australian
New Zealand
(FSANZ)
Adopted 2003
JECFA PTWI
(Karatela
et al. 2011)
Philippines Fish (except for
predatory)
0.5 mg
methyl Hg /
kg
Codex
Alimentarius
Predatory fish
(shark, tuna,
swordfish)
1mg
methyl
Hg/kg
(continued)
42 Mercury in Fish: History, Sources, Pathways, Effects, and Indicator Usage 759
Table 42.3 (continued)
Country/
organization Fish type
Maximum
allowed/
recommend
levels in
fish
a
Type of measure
Tolerable intake
levels
a
Slovak
Republic
Freshwater
non-predatory fish
and products thereof
0.1 mg total
Hg/kg
Slovak Food Code
Freshwater preda-
tory fish
0.5 mg
total Hg/kg
Marine
non-predatory fish
and products thereof
0.5 mg
total Hg/kg
Marine predatory
fish
1.0 mg
total Hg/kg
Thailand Seafood 0.5 μg Hg/g Food Containing
Contaminant
Standard
Other food 0.02 μg
Hg/g
United
Kingdom
Fish 0.3 mg
Hg/kg (wet
flesh)
European Statutory
Standard
United States Fish, shellfish and
other aquatic
animals (FDA)
1 ppm
methyl Hg
FDA action level US EPA reference
dose: 0.1 μg
methyl Hg/kg
body weight per
day
States, tribes and
territories are
responsible for
issuing fish
consumption advise
for locally-caught
fish; Trigger level
for many state
health departments:
0.5 ppm
methyl Hg
Local trigger level
WHO/FAO All fish except
predatory fish
0.5 mg
methyl
Hg/kg
FAO/WHO Codex
Alimentarius
guideline level
2003 JECFA
provisional
tolerable weekly
intake 1.6 μg
MeHg/kg body
weight per week
Predatory fish (such
as shark, swordfish,
tuna, pike and
others)
1mg
methyl
Hg/kg
From: Global Mercury Assessment, Chapter 4 (UNEP Chemicals 2002) unless otherwise noted.
Revised
a
Units as used in references. “mg/kg” equals “μg/g” and ppm (parts per million). It is assumed here
that fish limit values not mentioned as “wet weight” or “wet flesh” are most likely also based on
wet weight, as this is normally the case for analysis of fish for consumers
760 E.J. Zillioux
Compliance with regulatory guidelines, however, is often lacking. For example,
in the study of Hg in hair of exposed populations in the Florida Everglades
mentioned earlier, Fleming et al. (1995) found that, although 71 % of the 350
participants knew of the State Health Advisories concerning ingestion of
Hg-contaminated fish from the Everglades, this did not change their consumption
habits.
Conclusions
It would be difficult to find an indicator of potential harm more well-
researched than Hg in edible fish within the HgII !methanogen !MeHg
!fish !human pathway. For human consumption, the challenge is to
balance regulatory guidance for protection against exposure to MeHg at
potentially harmful levels with the well-known health benefits of fish con-
sumption. Since this review has not focused on the latter, a brief summation
of beneficial effects is warranted.
Clinical effects that support human health benefits of fish or fish oil intake
have been shown for anti-arrhythmia, anti-thrombosis and the lowering of
triglyceride, heart rate, and blood pressure. At moderate intake levels of
<750 mg per day EPA/DHA (eicosapentaenoic acid and docosahexaenoic
acid), the physiologic effects most likely to account for clinical cardiovascu-
lar benefits include modulation of myocardial sodium and calcium ion chan-
nels, and reduced left ventricular workload and improved myocardial
efficiency as a result of reduced heart rate, lower systemic vascular resistance,
and improved diastolic filling. The dose response for anti-arrhythmic effects
is initially steep, reaching a plateau at intake levels of around 750 mg/day
EPA/DHA. At increasing levels of intake up to at least 2,500 mg/day,
beneficial effects continue to accrue with respect to triglycerides, heart rate,
and blood pressure over a time course of months to years. In addition, fish or
fish oil intake may provide important beneficial effects with respect to
endothelial, autonomic, and inflammatory responses (Mozaffarian and
Rimm 2008, and references therein).
Among piscivorous wildlife, the population and ecosystem-level risks
from high environmental Hg concentrations in natural systems have proved
to be demonstrably greater than the current risk to human consumers. For
major health outcomes among adult humans, the benefits of fish consumption
generally outweigh risks; this is true even for sensitive populations of women
of child-bearing age and young children if health advisories and consumption
limits are followed. Further development of the application of Hg levels in
fish for the indication of potential threats to non-human species and to
ecological health in general is needed.
Acknowledgement Thanks are due to Dr. Curtis D. Pollman, of Aqua Lux Lucis, Inc., for his
review and helpful comments.
42 Mercury in Fish: History, Sources, Pathways, Effects, and Indicator Usage 761
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