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Arsenic in Marine Mammals, Seabirds,
and Sea Turtles
Takashi Kunito, Reiji Kubota, Junko Fujihara, Tetsuro Agusa,
and Shinsuke Tanabe
T. Kunito, R. Kubota, J. Fujihara, T. Agusa, S. Tanabe(*)
Center for Marine Environmental Studies (CMES), Ehime University, Bunkyo-cho 2-5,
Matsuyama 790-8577, Japan (shinsuke@agr.ehime-u.ac.jp)
T. Kunito
Department of Environmental Sciences, Faculty of Science, Shinshu University, 3-1-1 Asahi,
Matsumoto 390-8621, Japan.
R. Kubota
Division of Environmental Chemistry, National Institute of Health Sciences, Kamiyoga 1-18-1,
Setagaya-ku, Tokyo 158-8501, Japan.
J. Fujihara
Department of Legal Medicine, Shimane University School of Medicine, 89-1 Enya, Izumo,
Shimane 693-8501, Japan.
1 Introduction ......................................................................................................................... 32
2 Arsenic Species and Cycling in the Marine Ecosystem ..................................................... 34
2.1 Arsenic Species .......................................................................................................... 34
2.2 Microbial Degradation of Arsenobetaine ................................................................... 40
3 Distribution of Arsenic Species in the Tissues of Marine Mammals, Seabirds,
and Sea Turtles .................................................................................................................... 41
3.1 Arsenic in Marine Mammals ...................................................................................... 41
3.2 Arsenic in Seabirds ..................................................................................................... 44
3.3 Arsenic in Sea Turtles ................................................................................................ 46
3.4 Toxicological Significance of Arsenic in Marine Mammals, Seabirds,
and Sea Turtles ........................................................................................................... 48
3.5 Newly Identified Arsenicals ....................................................................................... 49
4 Maternal Transfer of Arsenic Species ................................................................................. 50
4.1 Maternal Transfer of Arsenic in Marine Mammals ................................................... 50
4.2 Maternal Transfer of Arsenic in Seabirds .................................................................. 51
5 Arsenobetaine: Accumulation Mechanism and Origin ....................................................... 51
5.1 Origin and Synthetic Pathway for Arsenobetaine ...................................................... 51
5.2 Accumulation Mechanism of Arsenobetaine in Marine Animals .............................. 53
5.3 Arsenobetaine in Freshwater and Terrestrial Environments ...................................... 55
D.M. Whitacre (ed.), Reviews of Environmental Contamination and Toxicology. 31
© Springer 2008
32 T. Kunito et al.
6 Lipid-Soluble Arsenic ......................................................................................................... 56
6.1 Lipid-Soluble Arsenicals in Marine Organisms at Low Trophic Levels ................... 56
6.2 Lipid-Soluble Arsenicals in Marine Animals ............................................................ 58
6.3 Promising Analytical Methods for Lipid-Soluble Arsenicals ................................... 58
7 Future Areas of Study ......................................................................................................... 59
8 Summary ............................................................................................................................. 60
References................................................................................................................................. 61
1 Introduction
Arsenic, a chalcophilic element, is widespread in the environment. Although
arsenic may possibly be an essential element for life (Cox 1995) and some micro-
organisms are known to use arsenic for energy generation (Oremland and Stolz
2003), no firm data are available on its essentiality for biological systems
(Francesconi 2005). In contrast to its possible essentiality in life, many studies
have focused on its high toxicity, which has been well known from various cases of
poisoning throughout the ages (Nriagu 2002). The toxicity is especially high for
inorganic arsenic; trivalent inorganic arsenic [arsenite; As(III)] is known to bind
readily to sulfhydryl groups of enzymes leading to enzyme inhibition, whereas
pentavalent inorganic arsenic [arsenate; As(V)], which is structurally similar to
phosphate, may disrupt metabolic reactions that require phosphorylation (Cox
1995). Symptoms of acute intoxication in humans by inorganic arsenic include
severe gastrointestinal disorders, hepatic and renal failure, and cardiovascular dis-
turbances, whereas chronic exposure causes skin pigmentation, hyperkeratosis,
and cancers in the lung, bladder, liver, and kidney as well as skin (Gorby 1994;
WHO 2001). At present, arsenic contamination of groundwater is a worldwide
problem (Mandal and Suzuki 2002), particularly in the Bengal Delta where
chronic ingestion of arsenic in groundwater poses a significant health risk to about
36 million people (Nordstrom 2002). Thus, the development and use of techniques
to remove arsenic from polluted groundwater is an urgent necessity (Chowdhury
2004). In contrast to the hazards of arsenic, it is useful in medicine. For example,
arsenic trioxide (As
2
O
3
) has recently attracted considerable attention as a thera-
peutic agent for treatment of acute promyelocytic leukemia and other cancers,
although the precise mechanisms by which it produces results are not fully under-
stood (Zhu et al. 2002).
Arsenic is used in agriculture, livestock, medicine, electronics, industry, and
metallurgy (Azcue and Nriagu 1994). Worldwide anthropogenic emission of
arsenic was estimated to be ∼5,000 t/yr in the mid-1990s, of which more than half
was accounted for by nonferrous metal production (Pacyna and Pacyna 2001).
Emission from natural sources, estimated to be ∼12,000 t/yr (Pacyna and Pacyna
2001), is more than twice that from anthropogenic sources. The major natural
source is volcanoes (Nriagu 1989). Therefore, both anthropogenic and natural
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 33
sources should be factored into evaluations when assessing the environmental risk
of arsenic.
Generally, no significant difference is observed between arsenic concentra-
tions in seawater and freshwater: arsenic concentration is about 1.5 µg L
−1
in sea water,
0.1–2.0 µg L
−1
in river water with absence of significant nearby emission sources
(e.g., mining activity and geothermal sources) and < 1 µg L
−1
in lake water (Plant
et al. 2005). However, it is widely known that marine organisms contain arsenic
at much higher concentrations than do terrestrial organisms (Lunde 1977), and
some marine species show arsenic levels exceeding 2,000 µg g
−1
dry wt on a
whole-body basis (Gibbs et al. 1983). Hence, many studies have been conducted
on arsenic levels and its speciation in marine organisms at low trophic levels
(e.g., algae and shellfish). In contrast, few studies are available for marine mam-
mals and seabirds occupying higher trophic levels, or sea turtles. It is known that
marine mammals and seabirds accumulate organochlorine compounds (e.g.,
polychlorinated biphenyls) at high levels by biomagnification through the
marine food chain (O’Shea 1999; Braune et al. 2005; Tanabe and Subramanian
2006). In particular, marine mammals have a unique tissue, blubber, which serves
as the main repository for organochlorine compounds. Furthermore, it has been
reported that marine mammals (Thompson 1990; Law 1996; O’Shea 1999), sea-
birds (Thompson 1990), and sea turtles (Anan et al. 2001) accumulate certain
metals, such as cadmium, mercury, and copper, at high concentrations in their
tissues. For example, some marine mammals show hepatic mercury levels of
13,000 µg g
−1
dry wt and a renal cadmium level of 800 µg g
−1
dry wt (O’Shea
1999). Such high accumulation of metals seems to depend not only on biomag-
nification through the food chain but also on various biological factors, such as
species, feeding habits, and lifespan (Thompson 1990). Indeed, it has been sug-
gested that organic mercury is virtually the only metal that can be biomagnified
through the food chain (Langston and Spence 1995). Although many studies
have been conducted on the accumulation of metals such as cadmium, copper,
mercury, and zinc in tissues of marine mammals, seabirds, and sea turtles, there
have been few efforts to study the presence of arsenic species in these animals.
To illustrate, although more than 18,000 papers have been published on the
accumulation of organochlorine compounds and metals in marine mammals
since the 1960s (O’Shea and Tanabe 2003), no report was published on arsenic
species in marine mammals (Eisler 1994; Law 1996) until the study of Goessler
et al. (1998); this is probably because sensitive speciation techniques for arsenic
were unavailable for many years. Because marine mammals, seabirds, and sea
turtles display unique features in metal accumulation, it may be useful to char-
acterize arsenic accumulation in these animals. Furthermore, studying arsenic
species and their presence in high-trophic-level marine animals is crucial for
understanding arsenic cycling in the marine ecosystem. In this review, we focus
attention on the pattern of accumulation of arsenic species in marine mammals,
seabirds, and sea turtles and also summarize the state of current knowledge
related to this topic (e.g., newly identified arsenicals in other marine organisms).
34 T. Kunito et al.
Fig. 1 Water-soluble arsenicals found in the marine ecosystem: As(V), arsenate; As(III), arsenite;
MA(V), methylarsonic acid; MA(III), methylarsonous acid; DMA(V), dimethylarsinic acid;
DMA(III), dimethylarsinous acid; DMAA, dimethylarsinoyl acetate; DMAE, dimethylarsinoyl
ethanol; DMAP, dimethylarsinoyl propionate; TMAO, trimethylarsine oxide; AC, arsenocholine;
AB, arsenobetaine; TMAP, trimethylarsoniopropionate; TETRA, tetramethylarsonium ion
2 Arsenic Species and Cycling in the Marine Ecosystem
2.1 Arsenic Species
Arsenic is present in various chemical forms (Fig. 1), and its toxicity depends on
the particular chemical form. Therefore, an understanding of arsenic speciation is
essential to understanding its environmental behavior and ecotoxicological effects.
Recently, the new scientific field, “metallomics,” which focuses on identification of
metallomes (metalloproteins, metalloenzymes, and other metal-containing biomolecules)
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 35
and the elucidation of their functions in biological systems, has received increasing
attention (Haraguchi 2004). Arsenic speciation is also of interest from the point of
view of this new field (Suzuki 2005). High performance liquid chromatography-inductively
coupled plasma-mass spectrometry (HPLC-ICP-MS) is the technique most
employed to determine arsenic speciation. Arsenic speciation is conducted using
ICP-MS as a detector after separation of arsenic species by HPLC. The HPLC-ICP-
MS technique is the most sensitive method for detecting arsenic species and often
allows a detection limit lower than 0.5 µg L
−1
(e.g., Hirata et al. 2006). The choice
of the HPLC column (e.g., anion-exchange, cation-exchange, or reversed-phase)
affects the separation of arsenic species, because different arsenic species behave
differently on each column. In part, this variability may occur because the dissocia-
tion constant, functional groups, and molecular size are largely different among
arsenic species. Therefore, HPLC columns should be selected based on character-
istics of the expected analytes and coexisting arsenicals. It should be noted that the
HPLC-ICP-MS technique requires the appropriate authentic reference material for
successful identification of arsenic species because no structural information is
inherently provided by this method. Identification of arsenic species depends on a
comparison of the retention times of unknowns versus such authentic standards
using selected HPLC column(s).
Seawater is considered as the starting point for arsenic cycling in the marine
ecosystem. In seawater, most arsenic exists in the inorganic form (Fig. 2), with pen-
tavalent arsenic, HAsO
4
2−
, predominating in oxygenated surface water (Cullen
and Reimer 1989). Arsenic shows a nutrient-like vertical profile in the water col-
umn (i.e., depletion at the euphotic zone), suggesting biological uptake of arsenic
Fig. 2 Arsenicals found in seawater and organisms at each trophic level
36 T. Kunito et al.
by marine phytoplankton, in spite of its high toxicity (Cullen and Reimer 1989;
Shibata and Morita 2000). In addition to inorganic arsenic, pentavalent mono-
methylated and dimethylated arsenicals, methylarsonic acid [MA(V)] and
dimethylarsinic acid [DMA(V)], respectively, are also present. Santosa et al.
(1996) reported that the ratio of MA(V) + DMA(V) to total arsenic increased
with water temperature and also was influenced by nutrient levels in Pacific
Ocean surface waters, which suggests that the abundance of organoarsenicals
reflect the biological activity [i.e., uptake of As(V) and subsequent methylation
to MA(V) and DMA(V)] of phytoplankton in the surface water. Furthermore, tri-
valent methylarsonous acid [MA(III)] and dimethylarsinous acid [DMA(III)]
were also detected in seawater (Hasegawa 1996). It should be noted that the
methylation pathway of inorganic arsenic is not yet firmly established. It has
been generally accepted that inorganic arsenic is methylated oxidatively (Fig. 3a),
but recently a reductive methylation pathway has also been proposed (Hayakawa
et al. 2005; Naranmandura et al. 2006). In the latter pathway, MA(V) and
DMA(V) are shown as the end products of transformation (Fig. 3b), which is
consistent with the abundant presence of these pentavalent arsenicals in animals
(Aposhian and Aposhian 2006).
In marine organisms, arsenic is known to exist mainly as organic forms,
although elucidation of the actual structures involved only took place over many
years. In 1977, arsenobetaine (AB; see Fig. 1) was first identified in the western
rock lobster (Panulirus cygnus) (Edmonds et al. 1977). AB was first synthesized in
the 1930s for pharmacological studies, but its presence in biota and the environ-
ment was not reported until 1977 (Edmonds et al. 1993). In 1981, arsenosugars
(Fig. 1) were also identified in brown kelp, Ecklonia radiata (Edmonds and
Francesconi 1981). Subsequent studies on arsenic species in various low-trophic-
level marine organisms revealed that marine algae, which rest at the base of the
marine food chain, accumulate arsenic (mainly as arsenosugars) at levels of
1,000–50,000 times that of seawater. Low-trophic-level marine animals contain
arsenic mainly as AB (see Fig. 2) at levels comparable to those in marine algae
(Francesconi and Edmonds 1993). Arsenosugars found in marine algae and in some
marine animals comprise the largest group (more than 20) of naturally occurring
arsenicals (Francesconi 2005). Interestingly, the composition of arsenosugars in
marine algae is related to their phylogeny: red and green algae contain arsenosug-
ars of rather simple structure, whereas in brown algae the structure of arsenosugars
is more complicated (Morita and Shibata 1990). Although AB was not detected in
marine algae until recently, a study by Nischwitz and Pergantis (2005a) revealed
the presence of AB in these organisms.
Major arsenicals found in marine ecosystems are shown in Figs. 1 and 2.
Arsenobetaine, arsenocholine (AC), trimethylarsine oxide (TMAO), and tetram-
ethylarsonium ion (TETRA) are the arseno-analogues of the nitrogen-containing
compounds, glycine betaine, choline, trimethylamine oxide, and tetramethylammonium
ion, respectively (Shibata et al. 1992). Thus, in uptake and retention, marine animals
do not discriminate these arsenicals from their natural nitrogen analogues (Shibata
et al. 1992).
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 37
Marine animals generally contain arsenic mainly as AB (Figs. 2, 4), with two
exceptions: the first, a marine teleost fish, the silver drummer (Kyphosus syd-
neyanus), which digests its macroalgal diet by fermentation (Edmonds et al.
1997), and second, the dugong (Dugong dugon), which feeds on seagrass
(Kubota et al. 2002a, 2003b). The proportion of AB in marine animals varies
depending on their feeding habit and trophic position, with animals of higher
trophic levels containing higher proportions of AB (Francesconi and Kuehnelt
2002). For example, AB comprises the major arsenic species in pelagic carnivo-
rous marine fish, whereas various arsenicals are contained in detritivorous and
herbivorous marine fish, with the corresponding proportion of AB being rela-
tively low (Kirby and Maher 2002). In general, the arsenic composition in
marine animals reflects the distribution found in their prey, because marine
animals take up arsenicals mainly through their diets (Phillips 1990). However,
the trophic transfer coefficient differs among species of arsenicals. Although
inorganic arsenic predominates in seawater, dimethylated arsenosugars and
Fig. 3 Hypothesized oxidative (a) and reductive (b) methylation pathways of inorganic arsenic
38 T. Kunito et al.
Fig. 4 Relationship between total arsenic and arsenobetaine concentrations in marine animals.
Total arsenic in marine mammals, seabirds, and sea turtles includes nonextractable arsenic in liver.
Data on crustaceans, mollusks, and fishes are from references cited in Francesconi and Edmonds
(1993); hepatic arsenic concentrations of marine mammals are from Kubota et al. (2003a),
Fujihara et al. (2003), and Goessler et al. (1998); those of seabirds are from Kubota et al. (2003a)
and Fujihara et al. (2003); and those of sea turtles are from Kubota et al. (2003a) and Fujihara
et al. (2003)
trimethylated arsenicals (i.e., AB) are contained as major arsenicals in marine
algae and marine animals, respectively (Fig. 2), with proportions of more methyl-
ated forms increasing with trophic level. Hence, the proportion of AB and
degree of methylation of predominant arsenicals are related to the trophic posi-
tion of the organism. However, the concentrations of total arsenic and AB vary
greatly among species (Francesconi and Edmonds 1993) and are not related to
the trophic position of the organism (Table 1).
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 39
The widely used HPLC-ICP-MS technique is effective in identification and
quantification of arsenicals for which corresponding authentic standards are available,
but it is not applicable in the absence of such standards. Recently, electrospray
tandem mass spectrometry (ES-MS/MS) has been increasingly used for identification
of arsenicals without using standard compounds, and has, therefore, contributed
significantly to the understanding of arsenic metabolism in marine organisms and
arsenic cycling in marine ecosystems (Edmonds and Francesconi 2003). For example,
McSheehy et al. (2002) identified 15 organoarsenicals in the kidney of giant clams
(Tridacna derasa), which have symbiotic unicellular algae in their tissues. These
identifications were achieved using ES-MS/MS after successive chromatographic
fractionation of the arsenicals by size-exclusion chromatography, anion-exchange
chromatography, and cation-exchange chromatography (or anion-exchange chro-
matography with a high resolution column). More recently, Nischwitz and Pergantis
(2006) established a HPLC-ES-MS/MS method that is capable of analyzing for 50
arsenic species, including various thio-arsenicals. Furthermore, in a recent study
using ES-MS/MS, dimethylarsinoyl acetate (DMAA), dimethylarsinoyl propionate
(DMAP) and dimethylarsinoyl ethanol (DMAE), which are postulated intermediates
in AB biosynthesis, were found in various marine animals and marine algae (Sloth
et al. 2005a). Also, thio-arsenosugars containing As=S have been recently identi-
fied in mussels and marine algae using this method (Fricke et al. 2004; Schmeisser
et al. 2004; Nischwitz et al. 2006) and HPLC-ICP-MS (Meier et al. 2005). It is
noteworthy that these newly identified arsenicals have also been quantified for
certified reference materials (CRMs) from marine animals (tuna fish, BCR-627;
dogfish muscle, DORM-2; mussel tissue, CRM278R and oyster, 1566b) (Nischwitz
and Pergantis 2005b)
Table 1 Arsenic concentrations in marine organisms (µg g
−1
dry wt)
Organism
Range
of means
(Phillips 1990)
Geometric
means (Neff
1997)
Means
(Kubota et al.
2001)
Range of
means (Saeki
et al. 2000)
Range
of means
(Kubota et al.
2003b)
Marine algae 1.5–84 43.7
Crustaceans 8.4–179 14.9
Bivalve mollusks 5.0–1025 10.4
Gastropod mol-
lusks
9.0–407 52.0
Cephalopod mol-
lusks
6.4–99 16.1
Fish < 0.2–216 5.6
Marine mammals
Pinnipeds (liver) 1.85
Cetaceans (liver) 1.88
Sea turtles (liver) 1.76–15.3
Seabirds (liver) 2.25–12.2
40 T. Kunito et al.
Fig. 5 Hypothesized degradation pathways of arsenobetaine
2.2 Microbial Degradation of Arsenobetaine
Because AB is by far the dominant arsenical in most marine animals, degradation
of AB to inorganic arsenic after its release into the environment from the decom-
posing dead animals is essential for completion of arsenic cycling in marine eco-
systems. There are two possible pathways for AB degradation (Fig. 5): the
conversion from AB to TMAO or from AB to DMAA. The TMAO or DMAA is
further degraded to inorganic arsenic through DMA(V) in both pathways. The AB-
degrading bacteria are ubiquitous in the marine environment. It was shown that
microbial communities from marine sediments, marine algae, mollusk intestine, or
suspended particles were able to convert AB to TMAO, DMA(V), and even to
inorganic arsenic (Hanaoka et al. 1992). Microbial communities on suspended par-
ticles collected at a depth of 3500 m were also able to degrade AB (Hanaoka et al.
1997). It is likely that aerobic microorganisms are primarily involved in degrada-
tion of AB, because AB is more rapidly degraded under aerobic than anaerobic
conditions (Hanaoka et al. 1992). Khokiattiwong et al. (2001) suggested that AB
is rapidly degraded when present at its environmentally relevant low level. In contrast,
degradation took several weeks in an incubation experiment conducted by Hanaoka
et al. (1992) in which a relatively high level of AB was employed. More than 95%
of AB was converted to DMA(V) within 24 hr by microorganisms in seawater to
which low levels of AB (100 and 750 µg As L
−1
) were added and in which shore
crabs (Carcinus maenas) were maintained (Khokiattiwong et al. 2001). A more
detailed investigation revealed that AB was first converted to DMAA, reaching a
maximum concentration after 3 hr incubation, and was then totally converted to
DMA(V) after 48 hr. Thus, the authors expect that AB is not usually detected in
seawater because of such rapid degradation. In these experiments, TMAO was not
detected, suggesting that AB was degraded to DMA(V) primarily via DMAA.
In addition to studies on AB degradation by microbial communities, isolation
and characterization of each AB-degrading bacterium have also been reported. Two
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 41
bacterial strains of the Vibrio-Aeromonas group isolated from coastal sediment by
the culture enrichment method converted AB to DMA(V) under aerobic but not
anaerobic conditions (Hanaoka et al. 1992). A microbial community isolated from
the blue mussel (Mytilus edulis) converted AB to TMAO, MA(V), and DMA(V).
Four AB-degrading bacterial strains (one of Paenibacillus, two of Pseudomonas,
and one of Aeromonas) were isolated from this community to characterize the
degradation pathway of AB (Jenkins et al. 2003). Degradation of AB to DMA(V)
by each strain occurred after 21 d incubation. TMAO was detected during incubation
with the microbial community, whereas DMAA but not TMAO was observed
during incubation in the pure culture. One isolate further degraded DMAA to
As(V) after 28 d incubation. Jenkins et al. (2003) assumed that the conversion of AB
to DMAA would be a fortuitous reaction, whereas conversion of DMAA to DMA(V)
would provide carbon or energy to the bacteria. Also, the degradation of AB was
shown to be mediated intracellularly by Paenibacillus sp. (Jenkins et al. 2003).
Devesa et al. (2005) observed that microbial communities from the hepatopancreas,
tail, and remaining parts of the red swamp crayfish (Procambarus clarkii) degraded
AB to TMAO, DMA(V), MA(V), and an unidentified arsenical. Interestingly, in the
incubation experiments using either AC, TETRA, TMAO, DMA(V), or MA(V),
only AC was converted to AB, but the other arsenicals were not transformed by
these microbial communities. Five AB-degrading strains isolated from these microbial
communities were all identified as Pseudomonas putida and were shown to degrade
AB to DMA(V) and MA(V) in the incubation experiment (Devesa et al. 2005).
Generally, microbial communities could degrade AB to inorganic arsenic, whereas
most of the AB-degrading bacteria could not degrade AB completely by themselves
(Hanaoka et al. 1992). Thus, degradation of AB to inorganic arsenic requires the
cooperation of various microorganisms. The microbial community structure in AB
degradation is important, because metabolites formed by the degradation are different
among the sources of microbial communities (Hanaoka and Kaise 1999).
3 Distribution of Arsenic Species in the Tissues of Marine
Mammals, Seabirds, and Sea Turtles
3.1 Arsenic in Marine Mammals
There are few studies that have examined the types of arsenic species in marine
mammals, seabirds and sea turtles. Therefore, we have undertaken a detailed
characterization of arsenic accumulation in such large marine animals.
Influences of feeding habits, age (or body size), and gender on the hepatic
arsenic level in marine mammals were examined by analyzing in-house measure-
ments of 16 species of marine mammals (n = 226), as well as data from the literature
(Kubota et al. 2001). The highest level of 7.68 µg g
−1
on a dry weight (dry wt) basis
was observed in liver of the harp seal (Pagophilus groenlandicus); levels were
42 T. Kunito et al.
Fig. 6 Influence of feeding habits on arsenic concentration in liver of marine mammals (Kubota
et al. 2001)
lower than those for animals at lower trophic levels (see Table 1). Hepatic levels were
comparable between pinnipeds and cetaceans (Table 1). Influences of gender and
age (or body size) on the arsenic level were not found (Kubota et al. 2001). The
relatively low arsenic level in marine mammals is probably because arsenic is
mainly present as AB, which has a short biological half-life in marine mammal tis-
sues. However, hepatic levels in marine mammals vary by species and depend on
feeding habits (Fig. 6); species feeding on cephalopods and crustaceans tend to
contain higher arsenic concentrations than those feeding on fish, which is consistent
with the pattern observed in prey organisms (Table 1). Generally, concentrations
of other trace elements also vary with feeding habits. For example, marine mam-
mals feeding on cephalopods show higher concentrations of cadmium and radioac-
tive cesium, whereas animals feeding on fish exhibit higher mercury levels
(Watanabe et al. 2002; Yoshitome et al. 2003).
Goessler et al. (1998) were among the first to report arsenic species in marine
mammals. These authors found AB to be the predominant arsenical in all the liver
samples of the ringed seal (Pusa hispida; n = 10), bearded seal (Erignathus barba-
tus; n = 1), pilot whale (Globicephala melas; n = 2), and beluga (Delphinapterus
leucas; n = 1), accounting for 68%–98% of extractable arsenic. Arsenocholine and
DMA(V) were also found in almost all samples, whereas MA(V) was detected only
in 5 specimens. Because TETRA was detected in pinnipeds but not in cetaceans,
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 43
these authors hypothesized that the presence or absence of TETRA reflects differences
in metabolism between these two groups of marine mammals (Goessler et al. 1998).
However, the limited number of samples used in their study (only 3 for cetaceans)
made it difficult to draw a definitive conclusion about the origin of TETRA.
Arsenic speciation analyses were performed in liver of Dall’s porpoise (Phocoenoides
dalli), short-finned pilot whale (Globicephala macrorhynchusfive), harp seal,
ringed seal, and dugong, also by Kubota et al. (2002a, 2003a). Arsenobetaine was
the dominant arsenical in all samples tested except in the dugong. Lower AB (42%)
and higher DMA(V) percentages (38%) were found in the Dall’s porpoise than in
other species (Fig. 7). Although TETRA was not detected in cetaceans by Goessler
et al. (1998), this arsenical was found in the Dall’s porpoise and also in pinnipeds,
the harp seal, and the ringed seal (Kubota et al. 2003a), indicating that TETRA is
present in both cetaceans and pinnipeds. Interestingly, AB was present in only a
trace amount in the dugong (n = 1), with MA(V) being the major arsenical followed
by DMA(V) (Kubota et al. 2003a). To confirm the generality of this unique com-
position of arsenicals in the dugong, arsenic speciation was conducted in four
additional liver samples of the dugong (Kubota et al. 2003b); the animals did not
contain AB at a detectable level and had appreciable amounts of MA(V) and
smaller quantities of DMA(V) (Kubota et al. 2003b), which was in agreement with
the result of Kubota et al. (2003a). These results can be attributed to the seagrass
diet of the dugong. Although only limited data are available, it is believed that, in
contrast to marine algae, seagrass might not contain arsenosugars (Shibata and
Morita 2000). Because seagrass is phylogenetically related to terrestrial higher
plants rather than to marine algae, it may contain principally MA(V) and DMA(V)
as do terrestrial plants (Kuehnelt and Goessler 2003).
Fig. 7 Arsenic species in liver of marine mammals, seabirds, and sea turtles (Kubota et al. 2003a;
Fujihara et al. 2003)
44 T. Kunito et al.
Inorganic arsenic was not detected in the liver samples of marine mammals we
examined. Sloth et al. (2005b) proposed a new method for determining inorganic
arsenic in animal samples that probably successfully extracts As(III) bound to thiol
groups of proteins. In their procedure, all inorganic arsenic was determined as As(V)
by HPLC-ICP-MS after microwave-assisted alkaline digestion of the sample
[oxidizing As(III) to As(V) in alkaline media]. Inorganic arsenic was not detected
in the minke whale (Balaenoptera acutorostrata), harp seal, and hooded seal
(Cystophora cristata), even though the aforementioned procedure was used (the tissue
was not mentioned in the paper, but it was probably muscle; Sloth et al. 2005b).
Almost all studies on arsenic have focused on its speciation in the liver (the main
metabolic organ) of marine mammals, but little is known about the distribution of
arsenic species in other tissues. Ebisuda et al. (2002) analyzed arsenic species in
liver, kidney, muscle, and gonad and total arsenic in blubber and hair of the ringed
seal (n = 18), and found that arsenic levels were highest in blubber, followed by
liver and kidney, and lowest in muscle, gonad, and hair on a wet weight basis.
Assuming that the respective tissue weight ratio of liver:kidney:muscle:blubber:
hair is 10:1:100:200:5, about 90% of the arsenic burden of the five tissues is esti-
mated to be present in ringed seal blubber. It is reported that the forms of arsenic
differ between dogfish muscle and liver (Wahlen et al. 2004). In these samples, AB
accounted for 96% of arsenic in muscle, and AB and DMA(V), respectively,
accounted for 79% and 16% of arsenic in liver, suggesting differences in arsenic
metabolism between the two tissues. However, there was no such difference in the
ringed seal, where AB accounted for more than 70% of extractable arsenic in all
the liver, kidney, muscle, and gonad (Ebisuda et al. 2002). It should be noted that
lipid-soluble arsenicals prevail only in the blubber, accounting for about 90% of
total arsenic (Ebisuda et al. 2002). Interestingly, AC was detected in all samples of
liver, kidney, and gonad, but not in all muscle samples (Ebisuda et al. 2002). The
predominant arsenical in the ringed seal was AB, followed by DMA(V) in stomach
contents, but levels of both decreased after passing through the gastrointestinal
tract whereas residual arsenic (nonextractable arsenic) increased. Arsenocholine
and TMAO were also detected in stomach contents, and AB, DMA(V), and AC
were present in tissues of the ringed seal, suggesting that AB, DMA(V), and AC
were derived from the diet. In contrast, TMAO was detected in stomach contents
but not in the tissues or contents of the intestine, whereas MA(V) was present in
tissues but not in stomach or intestinal contents. These differences in arsenical
forms between tissues of the ringed seal and contents of its stomach might be the
result of metabolism by the ringed seal itself and/or metabolism by the intestinal
bacteria it harbors.
3.2 Arsenic in Seabirds
Arsenic levels are higher in seabirds than terrestrial birds (Fig. 8), although there are few
studies that define arsenic species in birds. Arsenic speciation in birds was first reported
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 45
by Kubota et al. (2002b, 2003a). These authors reported average arsenic levels up to
12.2 µg g
−1
dry wt, and up to 26.7 µg g
−1
dry wt in liver of the black-footed albatross
(Phoebastria nigripes, n = 5), but only 2.25 µg g
−1
dry wt in liver of the black-tailed gull
(Larus crassirostris, n = 5) (see Fig. 7). The levels found in the black-tailed gull are
comparable to those found in marine mammals (see Table 1). We also analyzed 24
additional liver samples of the black-footed albatross and found concentrations up to
42 µg g
−1
dry wt (Fujihara et al. 2004), which is comparable to those in marine animals
at lower trophic levels (Table 1). Arsenobetaine predominated in liver of both the black-
footed albatross and black-tailed gull, with DMA(V), AC, and TETRA also being
detected (Fig. 7; Kubota et al. 2003a). Interestingly, a low level of AB (0.19 µg g
−1
dry
wt) was observed in one liver sample from a jungle crow (Corvus macrorhynchos)
specimen, although arsenic was not detected in four other specimens (Kubota et al.
2003a). Low levels of AB were also reported in some other terrestrial birds (Koch et al.
2005). The jungle crow might have obtained AB by eating leftover marine fish and
shellfish from garbage. Alternatively, because some terrestrial organisms have recently
been reported to contain AB at trace amounts (Kuehnelt and Goessler 2003), AB might
have originated in terrestrial organisms consumed by the jungle crow.
Tissue distribution of arsenicals in avian species has, so far, been reported only
for the black-tailed gull (Kubota et al. 2002b) and black-footed albatross (Fujihara
et al. 2004). Among the 13 tissues analyzed from the black-tailed gull, the concentration
of arsenic was highest in liver (mean, 1.59 µg g
−1
dry wt), followed by kidney
Fig. 8 Comparison of hepatic arsenic concentrations among birds from marine, coastal, and ter-
restrial environments (Kubota et al., unpublished results). Filled, shaded, and open bars corre-
spond to birds from marine, coastal, and terrestrial environments, respectively
46 T. Kunito et al.
(mean, 1.17 µg g
−1
dry wt), and was lowest in feathers (mean, 0.18 µg g
−1
dry wt),
with AB predominating in all the tissues (75%–97% of extractable arsenic; arsenic
speciation was not conducted for feathers) (Kubota et al. 2002b). In contrast, the
arsenic level was highest in lung (mean, 16 µg g
−1
dry wt) and muscle (mean, 15 µg
g
−1
dry wt), and lowest in bone (mean, < 0.1 µg g
−1
dry wt) and feathers (mean,
0.70 µg g
−1
dry wt), among the 17 tissues analyzed from the black-footed albatross
(Fujihara et al. 2004). Similar to the black-footed albatross, marine animals at low
trophic levels tend to accumulate arsenic in muscle. It is noteworthy that As(V) was
detected in the muscle and testis of the black-footed albatross but that inorganic
arsenic was not found in marine mammals.
The trophic transfer coefficient (TTC), defined as the ratio of the concentration
in a consumer’s body to the concentration in diet (stomach content) (Suedel et al.
1994), was found to be 1.0 for the black-footed albatross using arsenic levels of
17 tissues and diet (stomach content) (Fujihara et al. 2004). Because the TTC value
is usually below unity for trace elements, other than those that are highly accumulative
(e.g., mercury) (Anan et al. 2001), the black-footed albatross is believed to be very
efficient in absorbing arsenic although arsenic does not biomagnify, thus leading to the
levels reported (Fig. 8).
3.3 Arsenic in Sea Turtles
Few data are available on accumulation of arsenic or its species in sea turtles. In
studies conducted in our laboratories (Saeki et al. 2000; Kubota et al. 2003a;
Fujihara et al. 2003; Agusa et al. 2007), liver samples from the hawksbill turtle
(Eretmochelys imbricata, n = 19) showed the highest arsenic levels (mean, 20.9 µg
g
−1
dry wt), followed by the loggerhead turtle (Caretta caretta, n = 9) (mean, 9.0 µg
g
−1
dry wt), and the lowest in the green turtle (Chelonia mydas, n = 34) (mean,
2.9 µg g
−1
dry wt). Although the dugong, which feeds on seagrass, exhibited rela-
tively high arsenic concentrations in the liver (see Fig. 6), the carnivorous species
(i.e., hawksbill and loggerhead turtles) tended to show higher arsenic levels than
did herbivorous sea turtles (i.e., green turtle). This pattern in sea turtles is similar
to that observed in other low-trophic-level marine animals such as mollusks
(Cullen and Reimer 1989).
Edmonds et al. (1994) described the first characterization of arsenic species in
sea turtles. Arsenobetaine, As(III), and AC accounted for 50%, 35%, and 15% of
water-extractable arsenic, respectively, in liver of the leatherback turtle
(Dermochelys coriacea), whereas methanol extracts of AB and AC were 80% and
20%, respectively. The relatively high percentage of AC and As(III) is character-
istic of the leatherback turtle, but this does not occur in marine mammals and sea-
birds. In studies conducted in our laboratories, AB predominated in liver of the
green turtle, loggerhead turtle, and hawksbill turtle (see Fig. 7). Interestingly,
the loggerhead turtle showed a relatively high percentage of AC (30% of extractable
arsenic species) (Fig. 7), which was similar to that of the leatherback turtle
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 47
(Edmonds et al. 1994). It is known that most AC is converted to AB and also to a
small amount of lipid-soluble arsenicals when administered to marine animals
(Edmonds and Francesconi 2003). The high percentage of AC in these two turtle
species may originate in their diets, because they feed primarily on jellyfish
(Bjorndal 1997), some species of which are known to contain high AC levels
(Hanaoka et al. 2001a). Alternatively, it is assumed that these sea turtles have a
low capacity to convert AC to AB. It is noteworthy that green turtles feeding on
marine algae and seagrass (Bjorndal 1997) have high proportions of AB in their
livers (Fig. 7). Similar to the green turtle, the luderick (Girella tricuspidata), a
herbivorous fish species, contained 67% AB and 15% DMA(V) of total arsenic
extractable from the liver (Kirby and Maher 2002). Arsenosugars predominate in
marine algae, which comprise the primary diet of the green turtle (Francesconi
and Kuehnelt 2004). In our studies, although arsenosugars have not been meas-
ured, no significant HPLC peak other than AB, DMA(V), and AC has been
detected in the HPLC-ICP-MS analysis for the green turtle (Kubota et al. 2002a,
2003a). It is known that arsenosugars absorbed through the diet are converted
primarily to DMA(V) and are then excreted in the urine by humans (Le et al.
2004) and sheep (Martin et al. 2005). In the green turtle, however, the percentage
of DMA(V) was low (Fig. 7), despite possible uptake by this species of large
amounts of arsenosugars from marine algae. There are three possible explanations
for these discrepancies: first, AB may be synthesized from arsenosugars by the
green turtle or by intestinal bacteria they harbor; second, AB absorbed from diet ani-
mals (jellyfish and zooplankton) may be efficiently retained [adult green turtles
feed on small amount of jellyfish, and zooplankton is the chief diet of juvenile
turtles (Bjorndal 1997)], whereas DMA(V) converted from arsenosugars may be
rapidly excreted; and, third, the green turtles might efficiently retain in the body
any AB gleaned from marine algae, although only a small amount of AB may be
present in marine algae (Nischwitz and Pergantis 2005a).
High concentrations of arsenic were observed in the liver (up to 32.8 µg/g dry
wt) and muscle (205 µg/g) of hawksbill turtles (Saeki et al. 2000). These turtles may
have a peculiar mechanism for arsenic accumulation, because their main food
source, sponges, have rather low arsenic levels, when compared to other low-
trophic-level marine organisms (Saeki et al. 2000). Fujihara et al. (2004) summa-
rized the distribution of arsenic in tissues of various marine animals and concluded
that species with high arsenic levels (e.g., hawksbill turtle and the black-footed
albatross) tend to accumulate AB in the muscle. Such accumulation is also characteris-
tic of some fish (Shiomi et al. 1996; Amlund et al. 2006a,b). Agusa et al. (2007),
in reviewing the literature for arsenic levels in various marine animals, found the
ratio of arsenic concentration in muscle versus liver to be high in sea turtles (5.87).
Generally, inorganic arsenic is retained in mammalian tissues whereas organoars-
enicals are rapidly excreted in urine (Shiomi 1994). In contrast, AB and AC tend
to accumulate in fish tissues (especially muscle) while inorganic arsenic, DMA(V),
and TMAO are readily excreted (Shiomi et al. 1996; Amlund et al. 2006b).
As(III) was detected at low levels in two of five green turtle liver samples and one
of five loggerhead turtle liver samples (Kubota et al. 2003a). According to Agusa
48 T. Kunito et al.
et al. (2007), As(III) was detected in all examined tissues of green and hawksbill
turtles. Remarkably, high levels of As(III) were found in spleen of the hawksbill
turtle (2.83 µg g
−1
dry wt; Agusa et al. 2008). As(III) comprised 35% of water- extractable
arsenic in the liver of the leatherback turtle (Edmonds et al. 1994). Storelli and
Marcotrigiano (2000) analyzed organic and inorganic arsenic levels in the loggerhead
turtle and found that inorganic arsenic comprised 3% and 11% of total arsenic in the
muscle and liver, respectively. Although inorganic arsenic was not detected in liver of
the hawksbill turtle by Fujihara et al. (2003), sea turtles generally have higher levels of
inorganic arsenic than do marine mammals and seabirds.
A strong positive correlation was observed between AB and total arsenic concen-
trations in the liver of seabirds, sea turtles, and marine mammals (see Fig. 4). However,
some differences exist in AB accumulation among species. Arsenobetaine was not
detected in the dugong (not included in Fig. 4). Kubota et al. (2003a) reported that
the proportion of AB increased with total arsenic concentration in marine mammals,
seabirds, and sea turtles. Loggerhead turtles have a high arsenic level (mean, 11.2 µg
g
−1
dry wt; Fig. 7) and would, therefore, be expected to have high proportion of AB;
however, the value was relatively low (mean, 54.0%). The low proportion of AB was
attributed to high levels of AC in the loggerhead turtle (Fig. 7).
3.4 Toxicological Significance of Arsenic in Marine Mammals,
Seabirds, and Sea Turtles
In general, inorganic arsenic is more toxic than organic arsenic (Shiomi 1994).
Because AB is the dominant form in most marine mammals, seabirds, and sea
turtles, risk to these marine animals may be rather low despite retention of high
concentrations in their tissues. However, the more toxic inorganic form was
detected in some specimens of sea turtles and seabirds. Inorganic arsenic acts as a
carcinogen by forming certain reactive oxygen species (Kitchin 2001; Kitchin and
Ahmad 2003; Hei and Filipic 2004). Oxidative damage to DNA is indeed reported
in humans exposed to inorganic arsenic through contaminated groundwater (Feng
et al. 2001; Basu et al. 2005; Kubota et al. 2006). However, oxidative stress induced
by arsenic has not received much attention in marine organisms. Furthermore,
arsenic has recently been accused of being a potent endocrine disruptor (Darbre
2006). Stoica et al. (2000) showed that As(III) activated the estrogen receptor-α
(ER-α) through formation of a high-affinity complex with the hormone-binding
domain of the receptor in human breast cancer cells. Bodwell et al. (2004, 2006)
revealed that at very low levels As(III) stimulated transcription [mediated by
glucocorticoid receptors (GR), progesterone receptors, and mineralocorticoid
receptors of humans and rats), whereas at slightly higher but not cytotoxic concen-
trations, inhibition of transcription was observed. Waalkes et al. (2004) reported that
exposure of inorganic arsenic can cause overexpression of ER-α through its pro-
moter region hypomethylation in mice and humans. According to Stanton et al.
(2006) and Shaw et al. (2007), inorganic arsenic may act as an endocrine disruptor
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 49
in killifish; inorganic arsenic inhibits the ability of killifish to adapt to increased
salinity by altering GR-mediated posttranscriptional steps that regulate cystic fibrosis
transmembrane regulator (CFTR) protein abundance. Furthermore, some organoars-
enicals such as DMA(V) and DMA(III), as well as inorganic arsenic, show carci-
nogenic action, probably by inducing oxidative stress (Kitchin 2001; Kitchin and
Ahmad 2003; Hei and Filipic 2004). DMA(V) may be a carcinogen and tumor
promoter in some experimental animals (Yamanaka et al. 2004). Presumably,
dimethylarsenic peroxide may act as a tumor promoter and the dimethylarsenic
radical and dimethylarsenic peroxy radical act as tumor-initiating factors, all of
which seem to be metabolites of DMA(V) (Yamanaka et al. 2004). In HeLa S3
cells, As(III) induced oxidative DNA damage at 0.075 µg ml
−1
, MA(III) and
DMA(III) at 7.5 µg ml
−1
, and MA(V) and DMA(V) at 750 µg ml
−1
(Schwerdtle et
al. 2003). Because inorganic arsenic exerts adverse effects at low levels, its risk
should be assessed in marine animals in which As(III) and As(V) are found.
However, MA(V) and DMA(V) have not been found at levels that could adversely
affect marine mammals, seabirds, and sea turtles (Fig. 7), but effects of their
chronic exposure remain uncertain. Highly toxic MA(III) and DMA(III) have not
been detected in marine organisms.
Arsenobetaine is known to be scarcely metabolized in animals, but small amounts
of TMAO, TETRA, MA(V), DMA(V), As(V), and As(III) were detected in the urine
of rats administered orally with AB (Yoshida et al. 2001). Excreted forms in the rat
may result from degradation of AB by intestinal bacteria. Hence, effects of degrada-
tion products of AB, especially toxic inorganic arsenic, should be evaluated in marine
animals known to absorb large amounts of AB from the organisms they consume.
Mammal species vary considerably in their capacity to methylate arsenic (Aposhian
1997; Vahter 1999); inorganic arsenic is methylated to MA(V) and DMA(V) in most
mammalian species, but some species, such as the marmoset monkey and the chim-
panzee, have low or no methylation capacity. In humans, genetic polymorphisms are
known to affect arsenic biotransformation (Aposhian and Aposhian 2006), but no such
information is available for marine organisms. For the future, information is needed
on the metabolism of various forms of arsenic, not only in marine mammals, seabirds,
and sea turtles but also in other marine animals and algae.
3.5 Newly Identified Arsenicals
Recently, various new arsenicals have been identified in tissues of marine mammals
and other animals. Geiszinger et al. (2002) detected trimethylarsoniopropionate
(TMAP) in muscle, liver, kidney, and lung of the sperm whale (Physeter catodon).
The concentrations of TMAP were considerably lower than those of AB, but higher
than those of DMA(V) and AC, and accounted for 3%–5% of total arsenicals found.
Sloth et al. (2005a) detected DMAA, DMAP, and DMAE in the liver of hooded seal
and DMAE in the kidney of harp seal. Mancini et al. (2006) identified a novel
polyarsenic compound (arsenicin A; C
3
H
6
As
4
O
3
) in the marine sponge Echinochalina
50 T. Kunito et al.
bargibanti from the coast of New Caledonia. Recently, various thio-arsenicals were
also identified in the urine of sheep feeding on marine algae and humans exposed
to arsenic from groundwater. In addition, 2-dimethylarsinothioyl acetic acid
[(CH
3
)
2
As(=S)CH
2
COOH] was detected in urine of wild sheep feeding on brown
kelp (Laminaria hyperborea, L. digitata, etc.), which was the first identification of
a thio-arsenical in mammals (Hansen et al. 2004a). Thio-dimethylarsinate
[(CH
3
)
2
As(=S)OH; thio-DMA(V)] was also identified in urine of sheep (Hansen et al.
2003, 2004b). It is reported that humans convert arsenosugars to thio-arsenicals such
as thio-DMAE and thio-DMAA (Raml et al. 2005). Furthermore, thio-DMA and
thio-methylarsonate [CH
3
As(=S)(OH)
2
; thio-MA(V)] were identified in urine of
humans exposed to inorganic arsenic in groundwater in Bangladesh (Raml et al.
2007). The sulfur of these thio-arsenicals is thought to be derived from H
2
S, produced
by sulfate-reducing bacteria in the gastrointestinal tract (Conklin et al. 2006) and
released from cysteine degradation within cells (Hansen et al. 2004c). It is assumed
that the oxo (As-O) and thio (As-S) forms have been readily interconverted (Raml et
al. 2005). Surprisingly, Raab et al. (2007) identified a complex between thio-
DMA(V) and glutathione in shoots of cabbage (Brassica oleracea) exposed to
DMA(V), even though this is not a trivalent arsenic compound, suggesting that
pentavalent arsinothioyl species may interact with proteins. We have not examined
whether TMAP, DMAA, DMAP, and DMAE are present in other marine mammals,
seabirds, and sea turtles because the corresponding standard compounds necessary
for HPLC-ICP-MS analysis are unavailable. However, an unidentified arsenical,
with behavior on HPLC similar to that of TMAP, was found in extracts from vari-
ous marine mammals, seabirds, and sea turtles (Ebisuda et al. 2002; Kubota et al.
2003a, 2005). Also, thio-arsenicals, a new group of arsenic species, could be
present in marine mammals, seabirds, and sea turtles, although they were not iden-
tified in the studies we conducted. We used ion-exchange columns for separation
of arsenic species, but this method is not suitable for analysis of thio-arsenicals (Raml
et al. 2006). Instead, a reverse-phase HPLC method would enable analysis of thio-
arsenicals (Raml et al. 2006); thus, prospects for analyzing thio-arsenicals in
marine mammals, seabirds, and sea turtles look promising.
4 Maternal Transfer of Arsenic Species
4.1 Maternal Transfer of Arsenic in Marine Mammals
Very few studies have been performed on the maternal transfer of arsenic species
in marine mammals, seabirds, and sea turtles. Previously, it was reported that
inorganic arsenic would pass through mammalian placentas but that organic
arsenic would not (Morton and Dunnette 1994). However, women exposed to
inorganic arsenic from drinking water contained mainly DMA(V) in cord blood,
demonstrating placental transfer of some organoarsenicals in humans (Concha et al.
1998). Meador et al. (1993) detected arsenic in brain, liver, and kidney of fetal pilot
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 51
whales, confirming placental transfer of arsenic in marine mammals. However, the
arsenic species was not determined in either the mothers or fetuses. Kubota et al.
(2005) studied maternal transfer of arsenic to the fetus of the Dall’s porpoise. Analytes
included liver, kidney, muscle, and blubber of both mother and fetus. Arsenobetaine,
DMA(V), AC, and MA(V) were found in liver, kidney, and muscle of the fetus (arsenic
speciation was not conducted in the blubber), and the arsenical composition was
similar to that of the mother, suggesting that these arsenicals can transfer from
mother to fetus. However, the arsenic level in the fetus was less than one-half of
that in the mother. In the fetal Dall’s porpoise, 25.2% and 59.0% of total arsenic
burden was distributed in blubber and muscle, respectively, whereas 59.6% and
33.5% of the burden was distributed in blubber and muscle of the mother, respec-
tively. This difference reflects the low placental transferability of lipid-soluble
arsenicals from the mother’s blubber (Ebisuda et al. 2002, 2003). It is reported that
hydrophobic chemicals are much less transferable from mother to fetus in these
marine mammals (Tanabe et al. 1982).
4.2 Maternal Transfer of Arsenic in Seabirds
Limited information is available on the maternal transfer of arsenicals to bird eggs.
To our knowledge, only one study on arsenic species in bird eggs has been reported;
DMA(V) and As(III) but not AB were detected in eggs of the spoonbill (Platalea
leucorodia), although arsenic speciation was not conducted for the mother bird
(Gómes-Ariza et al. 2000). Kubota et al. (2002b) studied the maternal transfer of
arsenicals to eggs of the black-tailed gull. Arsenic composition in the eggs was
similar to that in tissues of the mother bird, with AB being predominant, followed
by DMA(V). However, the arsenic level in eggs was low compared to that in the
mother bird. The eggs weighed 32% of the body weight of the mother black-tailed
gull, but the percentage of arsenic in eggs was only 11% of that existing in the mother.
For the black-tailed gull (Agusa et al. 2005), the transfer rate of arsenic from
mother to eggs was comparable to that of vanadium, chromium, and antimony,
which are generally less transferable in birds. Arsenic was detected in eggs of sea
turtles (Lam et al. 2006), and, thus, is confirmed to transfer to eggs. However, as
far as we know, no studies exist on the nature of arsenic species in sea turtle eggs.
5 Arsenobetaine: Accumulation Mechanism and Origin
5.1 Origin and Synthetic Pathway of Arsenobetaine
The origin and synthetic pathways followed by AB are controversial. Principally,
four pathways for synthesis of AB have been proposed (Fig. 9). In the first two, AB
is transformed from dimethylated arsenosugars through DMAA or AC (Fig. 9a,b).
52 T. Kunito et al.
Fig. 9 Hypothesized synthetic pathways of arsenobetaine
In the third, AB is converted from trimethylated arsenosugars (Fig. 9c). Finally, it
is postulated that AB is synthesized from DMA(III) and 2-oxo acids, glyoxylate
(Fig. 9d) and pyruvate, a similar pathway as exists for amino acid biosynthesis.
DMAE and DMAA, at low levels, were observed in several marine animals and
marine algae (Sloth et al. 2005a). The existence of these two forms supports the
concept that a pathway exists from dimethylated arsenosugars (Fig. 9a,b) to AB.
Despite the natural occurrence of trimethylated arsenosugars in marine organisms,
their very low concentrations do not account for the presence of AB at a high
concentration in marine animals (Edmonds and Francesconi 2003). However, relatively
high concentrations of a trimethylated arsenosugar (2’3’-dihydroxypropyl 5-deoxy-
5-trimethyl arsonioriboside) were detected in abalone (Haliotis rubra) from New
South Wales, Australia. This arsenosugar accounted for 28% (5 µg g
−1
dry wt) of all
arsenicals in intestinal tissue and 0.9% (0.4 µg g
−1
dry wt) of the total in muscle
(Kirby et al. 2005). Hence, the trimethylated arsenosugar might contribute to
synthesis of AB in this marine animal. The pathway that produces AB in deep-sea
organisms and some terrestrial ones that are not dependent on marine algae is
unclear (Edmonds and Francesconi 2003). A pathway (Fig. 9d) starting from DMA(III),
recently proposed by Edmonds (2000), could explain the presence of AB in these
animals and also some other arsenicals found in marine organisms (Edmonds and
Francesconi 2003). For example, DMAA could be synthesized in this pathway,
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 53
although DMAE, an arsenical occasionally found in marine organisms, is not part
of this pathway. Furthermore, DMAP and TMAP, which are found in various
marine animals and algae, could be synthesized from oxaloacetate instead of gly-
oxylate (Fig. 9d). The major pathway leading to production of AB may vary with
location, organism, and ecosystem. To be sure, a comprehensive understanding of
this topic requires further research studies.
Arsenobetaine has been detected in marine animals but not in marine algae
(Francesconi and Edmonds 1993), so it is presumed that synthesis of AB is ascribed
to metabolic capacities of the animals and their intestinal bacteria. Involvement of
bacteria in AB synthesis has been reported by several researchers. Ritchie et al.
(2004) observed that AB was synthesized from DMAA and the methyl donor
S-adenosylmethionine by lysed-cell extracts of Pseudomonas fluorescens A (NCIMB
13944) isolated from the blue mussel. Interestingly, this bacterium was known to
degrade AB to DMAA (Jenkins et al. 2003), so the reaction in both directions might
be catalyzed by a methyltransferase (Ritchie et al. 2004). The results of Ritchie et
al. (2004) also suggest a direct involvement by bacteria in synthesis of AB within
marine animals. In 2005, Nischwitz and Pergantis (2005a) identified and quantified
AB in marine algae, by HPLC-ES-MS/MS, for the first time. In this study, commercially
available brown algal powder, and fresh green, brown, and red algae, which were
carefully washed to remove epifauna and contaminants from the surface, were
employed; visible epifauna with body size > 0.1 mm were fully removed, especially
from transparent green algae. Furthermore, extraction was performed by a mild pro-
cedure using only deionized water, and methanol and sonication were avoided to pre-
vent arsenic transformation. Analysis under these conditions revealed that AB
accounted for 7.5% of extracted arsenic in green algae and 0.25%–1.3% in other algal
samples. These authors also indicate that the chromatographic peak of AB present in
trace amounts cannot be separated chromatographically from the larger peaks of the
major arsenicals (i.e., arsenosugars) in marine algae, and therefore the presence of
AB could not be confirmed in marine algae by HPLC-ICP-MS analysis (Nischwitz
and Pergantis 2005a). It should be noted that the HPLC-ES-MS/MS enables analy-
sis of arsenicals co-eluting from the HPLC column, in contrast to HPLC-ICP-MS.
These results cast doubt on the general assumption that AB is not present in marine
algae and the belief that this arsenical is either synthesized or accumulated only in
marine animals. Therefore, further studies are needed to confirm these findings.
5.2 Accumulation Mechanism of Arsenobetaine
in Marine Animals
It has been pointed out that high levels of AB in marine animals may be related to
the salinity of seawater. Organisms are known to utilize various osmolytes, low
molecular weight osmotically active solutes, to adapt to osmotic stress. Glycine
betaine [GB, (CH
3
)
3
N
+
CH
2
COO
−
] is the nitrogen analogue of AB (Shibata et al.
1992) and behaves as an osmolyte in marine animals at lower trophic levels (Yancey
54 T. Kunito et al.
et al. 1982), in mammals (Burg et al. 1997), and in birds (Lien et al. 1993). It is
suggested that once AB is synthesized in the marine food chain, it may be taken up
into cells through the same route that absorbs GB and thereby behave as GB does
in the cells (Shibata et al. 1992; Shibata and Morita 2000). Indeed, it has been
reported that neither GB nor AB is bound to macromolecules (e.g., proteins) (Vahter
et al. 1983).
Arsenobetaine was not detected in the bivalve, Corbicula japonica, which lives
in a low-salinity estuary (Shibata and Morita 1992). This result suggests the possibility
that AB was not accumulated in the bivalve because osmolytes are not necessary in
a low-salinity environment. It was also shown that the blue mussel accumulated AB
efficiently from seawater, but the accumulation decreased in the presence of GB in
seawater (Gailer et al. 1995). Clowes and Francesconi (2004) reported that AB lev-
els increased in the blue mussel when the animals were maintained at high salinity.
When the blue mussel that had been maintained at high salinity was transferred to low-
salinity seawater, AB levels decreased in the gill but not in other tissues (Clowes and
Francesconi 2004). These results suggest that AB behaves as an osmolyte in the blue
mussel and that the gill responds sooner to osmotic changes than did other tissues.
Also, for herring (Clupea harengus), cod (Gadus morhua), and flounder (Platichthys
flesus), total arsenic concentrations in muscle were correlated with salinity at loca-
tions where the fish were collected, which may be because arsenic levels (probably
AB) in fish or their diet animals reflected the ambient salinity (Larsen and Francesconi
2003). Although controversial, Amlund and Berntssen (2004), after studying the
retention capacity of AB in seawater- and freshwater-adapted Atlantic salmon
(Salmo salar), found no significant difference between such groups, despite the
AB level in muscle of seawater-adapted wild salmon being more than 10 fold that
of freshwater-adapted wild salmon. Thus, the high AB level of seawater-adapted
wild salmon might be caused by the AB level in diet rather than an adaptation to
salinity. On the other hand, the bacterium Escherichia coli (Pichereau et al. 1997) and
Madin Darby canine kidney (MDCK) cells (Randall et al. 1996) absorbed AB and
GB in response to osmotic stress, although the uptake rate of AB was lower than that
of GB (Randall et al. 1996). Furthermore, it was shown that AB was efficiently
absorbed through two GB transporters, ProP and ProU, in Escherichia coli (Randall
et al. 1995). An alternative explanation for high concentrations of AB in marine ani-
mals is that AB might be largely distributed in cellular organelles. Vahter et al.
(1983) reported that urinary excretion of AB was slower in rabbits, which have more
AB in cellular organelles, than in mice or rats. The relationship between AB accu-
mulation and its subcellular distribution has not yet been examined in wildlife.
To gain insight into the mechanisms of the high AB accumulation, we determined
subcellular distribution of arsenic and the relationship between AB and GB concen-
trations in livers of the northern fur seal (Callorhinus ursinus), ringed seal, black-
footed albatross, black-tailed gull, hawksbill turtle, and green turtle (Fujihara et al.
2003). Results indicated that arsenic levels were not related to the subcellular dis-
tribution in these marine animals. However, a significant negative correlation was
observed between AB and GB concentrations for all animals examined (Fig. 10a);
a strong negative correlation was observed, especially for the black-footed albatross
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 55
(Fig. 10b). These results were contrary to our expectation; we had assumed that GB
might increase with increasing AB levels in the animals, but, instead, a negative
correlation was observed (Fig. 10). It is assumed that AB and GB are both taken up
and efficiently retained when osmolyte content (e.g., GB) is insufficient in these
animals or in their food supply. If true, this condition would lead to a negative corre-
lation between AB and GB levels in the marine animals. It is likely that the contribu-
tion of AB to osmoregulation was low because the AB level was remarkably lower
than that of GB (Fig. 10).
5.3 Arsenobetaine in Freshwater and Terrestrial Environments
Arsenic concentrations in seals from freshwater environments, the Baikal seal
(Pusa sibirica) (Kubota et al. 2001) from Russia, and the harbor seal (Phoca vitu-
lina) from northern Quebec in Canada (Langlois and Langis 1995), were lower
Fig. 10 Relationship between arsenobetaine and glycine betaine concentrations in liver of marine
mammals, seabirds, and sea turtles (Fujihara et al. 2003)
56 T. Kunito et al.
than those in seals from marine environments (see Fig. 6), suggesting that these
freshwater species are exposed to low AB concentrations in their food supply
because of the low salinity of freshwater habitats. It is frequently reported that AB
is not a dominant arsenical in freshwater animals, although arsenosugars dominate
in both freshwater and marine algae (Francesconi and Kuehnelt 2002). Nondetectable
or very low concentrations of AB was observed in freshwater animals from the
River Danube in Hungary (Schaeffer et al. 2006). Jankong et al. (2007) reported
relatively low accumulation of AB in tissues, especially liver, of four species of
freshwater fish collected from highly arsenic-contaminated ponds (550 and 990 µg
L
−1
). Soeroes et al. (2005) described the absence of AB in the common carp
(Cyprinus carpio) from lakes in Hungary. Low concentrations of AB in freshwater
fish may reflect the low salinity of their ambient environment. However, some
freshwater fish species contain predominantly AB (Shibata and Morita 2000;
Francesconi and Kuehnelt 2002). Šlejkovec et al. (2004) suggest that composition of
arsenic species is different among freshwater fish species; AB predominates espe-
cially in species of salmonids even though they inhabit freshwater environments
all through their life. Increasing evidence suggests that AB is present also in vari-
ous terrestrial organisms (Kuehnelt and Goessler 2003) such as mushrooms
(Kuehnelt et al. 1997a), earthworms (Geiszinger et al. 1998), and ants (Kuehnelt et
al. 1997b), although the levels are remarkably low. Future elucidation of the origin,
behavior, and function of AB, not only in the marine ecosystem, but also in fresh-
water and terrestrial ecosystems, is necessary.
6 Lipid-Soluble Arsenic
6.1 Lipid-Soluble Arsenicals in Marine Organisms
at Low Trophic Levels
It is well known that various marine organisms contain lipid-soluble arsenic. The
ascidian (Halocynthia roretzi), the turban shell (Turbo cornutus), the short-necked
clam (Tapes japonica) (Shinagawa et al. 1983), the red sea urchin (Pseudocentrotus
depressus), the abalone (Haliotis diversicolor supertexta), the three-line grunt
(Parapristipoma trilineatum), and the Japanese surfperch (Neoditrema ransonneti)
(Kaise et al. 1988) all showed relatively high concentrations of lipid-soluble
arsenic, although the levels were lower than those of water-soluble arsenic.
A phosphatidylarsenosugar was identified for the first time in a brown alga
(Undaria pinnatifida) as a lipid-soluble arsenical in 1988 (Fig. 11; Morita and
Shibata 1988). Phosphatidylarsenocholine (Fig. 11), a phosphatidylcholine ana-
logue, was also identified in the digestive gland of the Western rock lobster (Edmonds
et al. 1992). Because direct introduction of organic solvents, used for extraction of
lipid-soluble arsenicals, into ICP-MS is generally problematic, lipid-soluble arsenicals
in marine organisms are poorly studied. Until now, only water-soluble arsenicals
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 57
released from the lipid-soluble arsenicals have been studied after alkaline or acid
digestion. Unfortunately, these analyses cannot provide direct information on the
structure of lipid-soluble arsenicals.
In the star spotted shark (Mustelus manazo), almost all lipid-soluble arsenic was
detected in the polar lipid fraction. Lipid-soluble arsenic levels in this fraction were
particularly high in liver, gallbladder, and kidney (Hanaoka et al. 1999). Lipid-soluble
arsenic in the shark was fractionated into alkali (0.027 M NaOH)-stable and alkali-
labile portions, with the former predominating in liver and gallbladder, and the latter
in kidney (Hanaoka et al. 1999). Thus, the results suggest that the nature of lipid-
soluble arsenicals varies among tissues of the shark. Altogether, four arsenicals
were detected in the alkali-labile fraction of 12 shark tissues, with one of these
arsenicals predominating in kidney, spleen, and brain (Hanaoka et al. 2001b). After
digestion with 6 M HCl of three arsenicals detected in the alkali-labile fraction, AC
or DMA(V) was released from two of three of these arsenicals; HCl failed to digest
the third arsenical. Hydrolysis of the alkali-stable fraction with saturated Ba(OH)
2
gave primarily DMA(V) for liver, and DMA(V) and AC for muscle, skin, stomach,
and intestine (Hanaoka et al. 2001b). Therefore, at least six lipid-soluble arsenicals
exist in and contain precursors of DMA and AC in the star spotted shark. In contrast,
water-soluble arsenicals were not released by alkaline hydrolysis (1 M NaOH), but
DMA(V) was detected after acid digestion (conc. HNO
3
) in fish oil (Kohlmeyer et al.
2005). Lipid-soluble arsenicals were only detected in the polar lipid fraction of
Fig. 11 Lipid-soluble arsenicals identified in marine organisms
58 T. Kunito et al.
fish oil, where neutral lipids constituted more than 90% of the total (Kohlmeyer
et al. 2005); this is consistent with the distribution of lipid-soluble arsenicals in the
star spotted shark (Hanaoka et al. 1999).
6.2 Lipid-Soluble Arsenicals in Marine Mammals
As far as we know, no information was available on lipid-soluble arsenicals in
marine mammals, seabirds, and sea turtles before we conducted studies on the
blubber of marine mammals. Lipid-soluble arsenic was found in ringed seal liver,
kidney, muscle, and gonad tissues, but proportions were very low (Ebisuda et al.
2002). In contrast, lipid-soluble arsenic accounted for about 90% of arsenic in
ring seal blubber (Ebisuda et al. 2002). Hence, lipid-soluble arsenicals in the
ringed seal blubber were characterized using the method of Edmonds et al. (1992).
Results showed at least two lipid-soluble arsenicals: one was tetraethylammonium
hydroxide (TEAH) hydrolyzable, and the other was TEAH stable but NaOH
labile. Both these released DMA(V) after hydrolysis (Ebisuda et al. 2003). Thus,
it is suggested that marine mammals accumulate arsenic mostly in blubber as
DMA-containing lipid-soluble arsenicals. Structural identification of these lipid-
soluble arsenicals will be critical to understanding how arsenic is metabolized
in marine mammals. Insights into how DMA(V) is incorporated into lipid- soluble
arsenicals in marine animals would also be useful, considering the genotoxic
potential of DMA(V).
6.3 Promising Analytical Methods for Lipid-Soluble Arsenicals
Miyajima et al. (1988) purified and analyzed, with nuclear magnetic resonance
(NMR), a lipid-soluble arsenical from the tiger shark (Galeocerdo cuvier) without
digestion that revealed the presence of (CH
3
)
2
As(O)CH
2
−
in this arsenical. Devalla
and Feldmann (2003) characterized lipid-soluble arsenicals by an enzymatic hydrolytic
method. Water-soluble arsenicals, released by phospholipase D treatment of lipid-soluble
ones from a marine alga, was mainly an arsenosugar. In contrast, DMA(V) and
MA(V) were the major arsenicals released from lipid-soluble arsenicals in kidney,
and muscle, and DMA(V) from lipid-soluble ones in feces of marine algae-eating
sheep. Also, the digestion of the lipid-soluble arsenicals by phospholipase D indi-
cated that the released arsenicals were bound not to the simple lipid (comprising
91% of total lipid) but to a complex lipid (e.g., phospholipid and sphingolipid)
present in sheep tissues. Importantly, limited numbers of water-soluble arsenicals
released from lipid-soluble arsenicals may result from binding of an arsenical moiety to
various lipid-soluble species (Schmeisser et al. 2006a). Ninh et al. (2007) sug-
gested the possible presence of two DMA(V)-containing lipid-soluble arsenicals,
phosphatidyldimethylarsinic acid and DMA(V)-containing sphingomyelin, in the
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 59
Japanese flying squid (Todarodes pacificus) using a combination of chemical and
enzymatic hydrolysis techniques.
Little is known about the toxicity of lipid-soluble arsenicals in organisms
(Francesconi 2005). In humans, after metabolism, DMA(V) and trace amounts of
oxo-DMAP, thio-DMAP, oxo-DMAB, and thio-DMAB were excreted in the urine
(Schmeisser et al. 2006a,b). The toxicological evaluation of lipid-soluble arsenicals
and their metabolites will be the subject of future research.
Because it has been difficult to directly analyze for lipid-soluble arsenicals by
HPLC-ICP-MS, very little progress has been made with lipid-soluble arsenicals in
contrast to water-soluble arsenicals. Very recently, it became feasible to analyze
for lipid-soluble arsenicals directly by HPLC-ICP-MS (Schmeisser et al. 2005).
These authors found that problems associated with the introduction of organic sol-
vent to the plasma were considerably reduced by using a low column flow rate, a
cooled spray chamber (−5°C), and addition of oxygen directly to the plasma (20%
in argon). The authors applied this method to successfully observe the presence of
at least 10 lipid-soluble arsenicals in fish oil. This method would be useful to char-
acterize lipid-soluble arsenicals in marine mammals, seabirds, and sea turtles.
7 Future Areas of Study
More than 30 arsenicals have been identified in marine environments (Francesconi
and Kuehnelt 2004), and it is expected that more will be identified as analytical tech-
niques advance. For example, 15 unknown trace arsenicals, which were not detected
by conventional methods such as HPLC-ICP-MS and HPLC-hydride generation
(HG)-ICP-MS, were found in human urine using LC-high-efficiency photooxidation
(HEPO)-HG-ICP-MS (Nakazato and Tao 2006). The development of such new ana-
lytical techniques is important because such tools improve our understanding of
arsenics behavior in geochemical cycles, ecosystems, organisms, and transformation
pathways, as well as their toxicity. If progress is to be made, convenient methods of
synthesis for standard compounds of recently identified arsenicals and commercial
production thereof are badly needed. As recently identified arsenicals (e.g., DMAA
and TMAP) have been quantified in some CRMs such as DORM2 (dogfish muscle)
and BCR CRM627 (tuna fish tissue) (Sloth et al. 2003), these CRMs can be utilized
to evaluate the accuracy of methods for these arsenicals.
Studies on the interaction between arsenic and other elements in marine organ-
isms are also necessary. It is well known that mercury is detoxified through binding
to selenium and/or sulfur in marine mammals and seabirds (Shibata et al. 1992; Ng et
al. 2001; Arai et al. 2004; Ikemoto et al. 2004). Similarly, interaction between
arsenic and selenium and/or sulfur may also occur (Gailer 2007). A metabolite
containing glutathione (GSH) and equimolar amounts of As and Se, [(GS)
2
AsSe]
−
,
was identified in bile from rabbits injected with selenite [Se(IV)] and As(III)
(Gailer et al. 2000). This metabolite is thought to be synthesized in hepatocytes
(Gailer et al. 2002a) and erythrocytes (Manley et al. 2006) with high endogenous
60 T. Kunito et al.
concentrations of GSH. Furthermore, [(CH
3
)
2
AsSe
2
]
−
can be synthesized chemi-
cally by reacting DMA(V) with GSH and Se(IV) (Gailer et al. 2002b). Even though
these metabolites are present in marine mammals, seabirds, and sea turtles, the lev-
els may be low. Small granules (diameter about 3 nm) of As
2
Se were observed in
the kidney of rats injected with inorganic arsenic and selenium (Berry and Galle
1994), and thus it will be interesting to examine whether such granules are present
in marine animals. Also, Kanaki and Pergantis (2007) recently synthesized seleno-
arsenosugars and seleno-DMA(V) by reacting arsenosugars and DMA(V) with
H
2
Se, respectively; these reaction products are unstable and, if present, are not
expected to occur in organisms at significant levels.
DMA(V), which is widely distributed in marine animals, was thought to be
a metabolite formed during the detoxification of inorganic arsenic; however,
the metabolites of the methylation process for inorganic arsenic (see Fig. 3)
include MA(III) and DMA(III), which have recently been reported to be highly
toxic; hence, this process is no longer regarded as one of detoxification for
inorganic arsenic. The possible carcinogenicity of DMA(V) should induce
additional interest in conducting safety evaluations on marine animals.
Furthermore, arsenosugars, which are generally considered to be practically non-
toxic, exhibit genotoxicity by forming reactive oxygen species when present as
trivalent arsenicals (Andrewes et al. 2004). Such trivalent arsenosugars are readily
formed by the reaction of pentavalent arsenosugars with thiols (Andrewes et al.
2004). Additional evaluations of the in vivo toxicity of arsenosugars are there-
fore needed.
Nonextractable arsenic in the tissues of marine organisms should be studied. For
example, an average of 0.73 µg g
−1
dry wt of arsenic was present as nonextractable
residue after extraction with methanol:water (9:1 v/v) in the liver of green turtles
(Kubota et al. 2003a). This residue constituted about 25% of total arsenic. Thus,
nonextractable arsenic cannot be ignored if the complete pattern of arsenic metabo-
lism in marine organisms is to be understood. Arsenic-binding proteins are known
to exist in experimental animals, and such proteins may be involved in the trans-
formation and detoxification of arsenic (Aposhian and Aposhian 2006). For exam-
ple, it was reported that most of the As(III) added to rat liver cytosol became
protein bound, and this binding affected the extent of its subsequent methylation
(Styblo et al. 1996; Styblo and Thomas 1997). Trivalent methylated arsenicals,
MA(III) and DMA(III), also bind to proteins (Naranmandura et al. 2006). Further
work is needed to clarify the role arsenic-bound proteins play in the metabolism of
arsenic by marine mammals, seabirds, and sea turtles.
8 Summary
Although there have been numerous studies on arsenic in low-trophic-level marine
organisms, few studies exist on arsenic in marine mammals, seabirds, and sea turtles.
Studies on arsenic species and their concentrations in these animals are needed to
Arsenic in Marine Mammals, Seabirds, and Sea Turtles 61
evaluate their possible health effects and to deepen our understanding of how arsenic
behaves and cycles in marine ecosystems. Most arsenic in the livers of marine mam-
mals, seabirds, and sea turtles is AB, but this form is absent or occurs at surprisingly
low levels in the dugong. Although arsenic levels were low in marine mammals, some
seabirds, and some sea turtles, the black-footed albatross and hawksbill and loggerhead
turtles showed high concentrations, comparable to those in marine organisms at low
trophic levels. Hence, these animals may have a specific mechanism for accumulat-
ing arsenic. Osmoregulation in these animals may play a role in the high accumula-
tion of AB. Highly toxic inorganic arsenic is found in some seabirds and sea turtles,
and some evidence suggests it may act as an endocrine disruptor, requiring new and
more detailed studies for confirmation. Furthermore, DMA(V) and arsenosugars,
which are commonly found in marine animals and marine algae, respectively, might
pose risks to highly exposed animals because of their tendency to form reactive oxygen
species. In marine mammals, arsenic is thought to be mainly stored in blubber as lipid-
soluble arsenicals. Because marine mammals occupy the top levels of their food
chain, work to characterize the lipid-soluble arsenicals and how they cycle in marine
ecosystems is needed. These lipid-soluble arsenicals have DMA precursors, the exact
structures of which remain to be determined. Because many more arsenicals are
assumed to be present in the marine environment, further advances in analytical
capabilities can and will provide useful future information on the transformation and
cycling of arsenic in the marine environment.
Acknowledgments We are grateful to Dr. Y. Shibata (National Institute for Environmental
Studies, Japan), Prof. N. Miyazaki (the University of Tokyo, Japan), Dr. J. Yang (Freshwater
Fisheries Research Center, China), Prof. H. Ogi (Hokkaido University, Japan), Dr. H. Tanaka
(National Research Institute of Fisheries and Environment of Inland Sea, Japan), and Mr. K. Ebisuda
(Ehime University, Japan) for their help and discussion. We also thank Prof. A. Subramanian
(Ehime University, Japan) for critical reading of the manuscript. Work on these topics in our laboratory
was supported by the “21st Century COE Program” and “Global COE Program” from the
Ministry of Education, Culture, Sports, Science and Technology (MEXT), Japan and Japan
Society for the Promotion of Science (JSPS), respectively.
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