FEMS Microbiology Reviews 11 (1993) 297-316
© 1993 Federation of European Microbiological Societies 0168-6445/93/$15.00
Published by Elsevier
297
FEMSRE 00317
Microbial formation and transformation
of organometallic and organometalloid compounds
G.M. G a d d
Department of Biological Sciences, University of Dundee, Dundee, UK
(Received 10 February 1993; accepted 4 May 1993)
Abstract: Microbial formation and transformation of organometallic and organometalloid compounds comprise significant components of biogeochemical cycles for the metals mercury, lead and tin and the metalloids arsenic, selenium, tellurium and germanium.
Methylated derivatives of such elements can arise as a result of chemical and biological mechanisms and this frequently results in
altered volatility, solubility, toxicity and mobility. The major microbial methylating agents are methylcobalamin (CH3CoB12),
involved in the methylation of mercury, tin and lead, and S-adenosylmethionine (SAM), involved in the methylation of arsenic and
selenium. Evidence for the methylation of other toxic metal(loid)s is sparse. Biomethylation may result in metal(loid) detoxification
since methylated derivatives may be excreted readily from cells, are often volatile and may be less toxic, e.g. organoarsenicals.
However, for mercury, low yields of methylated derivatives and the existence of more efficient resistance mechanisms, e.g.
reduction of Hg 2÷ to Hg °, suggest a lower significance in detoxification. Bioalkylation has only been characterised in detail for
arsenic. Microorganisms can accumulate organometal(Ioid)s, a phenomenon relevant to toxicant transfer to higher organisms. As
well as bioaccumulation, many microorganisms are capable of the degradation and detoxification of organometal(loid) compounds
by, e.g. demethylation and dealkylation. Several organometal(loid) transformations have potential for environmental bioremediation.
Key words: Organometal; Organometalloid; Biomethylation; Bioalkylation; Organometal(loid); Degradation
Introduction
An organometallic compound may be simply
defined as a compound containing at least one
metal-carbon bond. Where such compounds contain 'metalloid' elements such as germanium, arsenic, selenium or tellurium, which are less electronegative than carbon but are not considered to
Correspondence to: G.M. Gadd, Department of Biological
Sciences, University of Dundee, Dundee DD1 4HN, UK.
be true metals by chemists, the term 'organometalloid' is frequently used [1]. Simple alkyl
and aryl derivatives are those formed by replacement of one or more carbon-hydrogen bonds by
metal (or metalloid)-carbon bonds, the carbon
atom remaining tetravalent, e.g. (C2Hs)4Pb.
Other kinds of organometallic compounds include those where the organic group is bonded to
the metal through the electrons in carbon-carbon
pi-bonds, e.g. (CsHs)2Fe (ferrocene), the carbon
atom again being tetravalent [1]. Detailed accounts of the chemistry of organometal(loid)
compounds are available [1-6].
298
Organometal(loid) compounds arise in the environment as a result of natural processes, including those mediated by living organisms, or because of accidental or deliberate introduction.
Increasing industrial use of organometallic compounds has lead to increased environmental discharge in effluents a n d / o r other waste products,
e.g. methylmercury compounds, tetraethyllead.
Deliberate environmental introduction results
from the widespread use of many organometal
(loid)s as biocides, e.g. organoarsenicals, organotins and organomercurials [1,7-9]. A characteristic of all biologically important organometallic
compounds is that they have some degree of
stability towards water. This is because of the
crucial role of water in cell metabolism and also
for transport and mobilization in aquatic ecosystems [1,6]. Although a voluminous literature on
the environmental impact of organometal(loids)
now exists in relation to toxicity, biodegradative
pathways and methods of analysis [1,3,4,6,10-13],
there is still a lack of appreciation relating to the
important roles played by microorganisms in organometal(loid) transformations in the environment. Methylated derivatives of several elements
naturally arise in the environment as a result of
chemical and biological methylation, microorganisms playing highly significant roles in the latter
process [1,6]. Such processes are components of
natural global biogeochemical cycles for elements
which include mercury, arsenic, tin, lead, selenium, tellurium, germanium and antimony [5].
Furthermore, microorganisms are capable of the
degradation and detoxification of organometal
(loid)s arising from natural and anthropogenic
activities, some of these processes being relevant
to environmental bioremediation [14].
This review seeks to emphasise the fundamental roles of microorganisms in the formation and
transformation of organometal(loid)s in aquatic
and terrestrial ecosystems. Particular attention is
given to the chemical and biochemical mechanisms involved in organometal(loid) synthesis and
degradation as well as important abiotic influences and their contribution to the overall process. Where appropriate, the potential of microbial organometal(loid) transformations for environmental bioremediation is also discussed.
Microbial formation of organometal(loid) compounds
Biological alkylation (= bioalkylation) refers to
the mechanism by which alkyl groups are linked
to metal or metalloid atoms, thus forming
metal(loid)-carbon bonds, as a result of direct or
indirect biological action. The most common alkyl
group transferred is the methyl group, this process being termed biological methylation
(=biomethylation). The introduction of alkyl
groups other than methyl is much rarer though
h a s been detected for arsenic and selenium. There
appears no evidence for bioarylation [1].
Biomethylation
Intracellular biomethylation of organic molecules, e.g. nucleic acid bases, proteins, fatty acids
and polysaccharides, is an important aspect of the
physiology of pro- and eukaryotic cells [15]. The
major biological methylating agents are S-adenosylmethionine (SAM), methyltetrahydrofolic acid
(THF) and methylcobalamin (CH3CoB12), a
derivative of vitamin B12 (Fig. 1) and these are
also involved in the biomethylation of potentially
toxic metals [1,15,16-20]. The methyl group is
frequently transferred to the metal, e.g. Hg(II), as
a carbanion (CH 3) which, although usually arising from methylcobalamin (which appears to be
the sole carbanion-donating natural methylating
agent), may sometimes arise from other organometallic species present in the external environment, e.g. (CH3)3Sn +, (CH3)3Pb + [7]. Natural
methylating agents like methylcobalamin (involved in biomethylation of Hg, Pb, Sn, Pd, Pt,
Au and TI [17], S-adenosylmethionine (involved
in biomethylation of As and Se) and N-methyltetrahydrofolate may also transfer methyl groups as
radicals (CH3) or as carbonium ions (CH~-) [1,17].
As well as these, many other biomolecules, e.g.
iodomethane (methyl iodide, CH3I), betaine, humic acids, may oxidize and methylate metal
species [7,19,20-22]. It should be mentioned that
considerable quantities of methylated metal compounds are introduced into the environment as a
result of industrial activities and pollution which
can complicate assessment of microbial methylation [23,24]. These include (CH3)3S+I - and
299
CONH2CONH2
H~oc
H3C "'~
.%
N/~-'N ~--=--~ CH3
H~OC/"<
a
H3 CH3
CONH2
r.--')'--H H O
N~CU
0 ~/ ~ 0 - ""OH
b
NH2
[
N~C"-C/N\\
CH3 l
11 CH
0%
HC~N/C'~N /
C-CH-CHrCHrS-CH2 ~O_
I
HO
NH2
H
~
OH
H:jN..,~N~.N~.
H
OH
COOH
CH2
I
CHz
/N..~/~- CH2-NH/--{(~)-'CO-NH-CH-COOH
C
H
~
~H3
Fig. 1. Structural formulae of major biological methylating
agents. (a) methylcobalamin;(b) S-adenosylmethionine;(e)
N S-methyl-tetrahydrofolieacid.
(CH3)3Sn ÷ which can themselves act as aquatic
metal methylating agents [22].
Mercury biomethylation. Mercury methylation
can be catalysed by many bacteria, and certain
other microorganisms, in aquatic and terrestrial
habitats. The process generally reaches an equilibrium with the opposite process of demethyla-
tion and other removal mechanisms [25,26] (see
later). Although encountered in aerobic and
anaerobic environments, maximal rates of methy!ation occur under anaerobic conditions at redox
potentials ranging from -100 to + 150 mV [9];
sulphate-reducing bacteria appear to be important methylators of mercury and tin [27]. Factors
that influence rates of methylation include the
Hg 2÷ concentration and those physico-chemical
factors that govern microbial activity and metal
ion availability in natural habitats, e.g. chemical
speciation, organic and inorganic components,
pH, E,, temperature, etc. [16,23,24,26,28-32].
Clay minerals, Fe and Mn oxides, and humic
materials have a major influence on Hg methylation (and demethylation) and can strongly inhibit
or stimulate Hg transformations, possibly by altering composition of the microbial community
[33]. Sulphide is highly important in sediments
since at S 2- concentrations > 1.8 mg g - ' , intractable mercuric sulphide is formed. Sulphide
may also promote removal of any methylmercury
formed by dismutation reactions. Rapid photolytic decay of methylmercury occurs in the atmosphere while in rivers, 95% of methylmercury
may be bound to solid particulates [9].
As mentioned, the most important naturally
occurring Hge+-methylating molecule is methylcobalamin (CH3CoB12). Two products can form,
salts of the methyimercuric ion, CH3Hg + ('methylmercury') and gaseous dimethylmercury ((CH 3)2
Hg):
CH3CoB12 + Hg 2+ H20) CHsHg++ H2OCoB~2
CH3CoBIe + CH3Hg +
H20
' (CH3)2Hg + H2OCoB~-2
Methylmercury usually predominates, the formation of dimethylmercury proceeding at much
slower rates.
The other main natural methylating agents,
S-adenosylmethionine and N-methyltetrahydrofolate are not involved in mercury biomethylation
since they transfer methyl groups as carbonium
ions (CH~) which are unlikely to transfer to
positive Hg z+ [9,16]. In some cases where there is
extensive pollution by alkyl leads and Hg 2+,
transalkylation from the lead compound may re-
300
of sediments and enter the atmosphere thus resuiting in wider diffusion in the environment (see
later). In the biogeochemical mercury cycle (Fig.
2), there is interchange between methylated mercury compounds, Hg 2÷ and Hg ° which can account for the relatively low persistence of
CH3Hg ÷ in the environment.
It is significant that some end-products of glucose metabolism, e.g. acetate and propionate, may
also donate methyl groups extracellularly to
Hg(II) under UV (sunlight) irradiation [20]:
suit in abiotic Hg 2+ methylation by carbanion
transfer:( C H 3 ) 3 P b + + H g 2+
CH3Hg +
As mentioned previously, CH3Hg + may be
further biologically methylated by CH3CoB12 to
(CH3)2Hg. Further methylation of CH3Hg + may
also occur by disproportionation reactions involving S2- or HeS:, (CH3)2Pb2++
2CH3Hg++
S 2-
,
(CH3Hg)2S
Hg2+ + CH3CO2
, (CH 3) 2Hg + HgS
Light may be required for process completion.
The dimethylmercury so formed may diffuse out
Anthropogenic sources
Volcanic activity etc.
CH3Hg++
(CH3)2Hg
CH3CO 2
hv
Fish
~ Cell 6 + Hg °
Human consumption
air
,. Predation
F
water
Shellfish
CH~Hg"
i~
l
CHj
+ Hg" ~
T
b
Hg2" ~
H2
CH~Hg"
I
a
HgO +
Hg2.
sediment
--- (CH,)2Hg
J~lDMPT
CH,HgSCH~
h~' (CH,):Pb:'),pb.2 C H / H g T f S ~
c
, (CH3)2Hg + CO 2
(CH,):Hg
g
CO 2
h v
CH, + C_,H~
CH~Hg"
-~ Hg"-~
CHJ
k
k
hv , C H 3 H g + +
L H~"
P
CH3HgSCH3
e
Fig. 2. Simplified biogeochemical cycle for mercury in the environment. Main reactions detailed are (a) methylation; (b)
demethylation; (c) reduction; (d) oxidation; (e) disproportionation of mercurous cation; (f) chemical disproportionation by H 2S; (g)
photolytic decay; (h) transalkylation by alkyllead compounds; (i) methylation be methyl iodide; (j) formation of methyl iodide by
possible reaction of dimethyl /3-propiothetin (Fig. 3), a ubiquitous algal metabolite, with aqueous iodide [20]; (k) deposition by
rainfall, pollution. Adapted with permission from Wood and Wang [62] and Craig [9]. Reactions (a)-(d) catalysed by microorganisms, chiefly bacteria.
301
Tin biomethylation. Methyltin compounds are
widely found in natural environments although it
is usually difficult to ascertain the relative proportions that arise from anthropogenic sources,
abiotic or biotic methylation reactions or degradation of more complex organotin molecules
[4,17,18,23,24,34-36].
Biomethylation of inorganic SnC14 • 5H20 has
been demonstrated with pure bacterial cultures,
e.g. Pseudomonas sp., trimethyltin and methylstannanes ((CH3)nSnH4_ n (n = 2-4)) arising as
products [34,37,38]. Other work has used natural
aquatic/sediment samples with biomethylation of
inorganic Sn(II) and Sn(IV) compounds and certain organotins being demonstrated; methylated
derivatives did not arise in sterile or poisoned
controls [34]. Products arising from Sn(IV)
or methyltin compounds included CH3Sn 3+,
(CH3)2Sn 2+, (CH3)3Sn +, (CH3)4Sn and methyltin hydrides of possible composition (CH3) 2
SnH 2 and (CH3)3SnH [4,34,39] although monomethyltin may comprise > 90% of the alkyltin
products formed [40]. On incubation of trimethyltin hydroxide, (CH3)3SnOH, with sediment
samples, tetramethyltin, (CH3)4Sn, was detectable in both non-sterile and sterile preparations (in the ration of 2.7 : 1) [39]. (CH3)48n production was enhanced by S 2- addition (1) and it
has been proposed that some of the (CH3)4Sn is
formed abiotically by disproportionation of the
initially formed bis(trimethyltin)sulphide from
naturally occurring sulphides (2) [41]:
2(CH3)3Sn++ S~-
, ((CHs)3Sn):S
tion reactions, e.g. between organotins and mercury (4) [18]:
3RaSn + 3SnC14 .
+ 2RSnC13
(3)
Hg 2+ + (CH3)3Sn +
.
• CH3Hg++ (CH3)2Sn 2+
(4)
However, a detailed study has failed to find evidence for inorganic tin methylation in the absence of microbial activity and bacterial enrichments and axenic anaerobe cultures retained their
ability to produce methyltins in the absence of
sediment. The majority of methylating bacteria
were anaerobic sulphate reducers [40,43]. It appears some butyltin species can be partially degraded by physico-chemical and microbial action
(see later) and the products then methylated since
di- and tributylmethyltin have been found in organotin-polluted sediments [13,44].
Methylcobalamin (CH3CoB12) has been proposed as the main methylating agent for tin compounds. The proposed mechanism involves a
stannyl radical, Sn(III) arising from the oxidation
of Sn(II), Fe(III) being used as the oxidizing
agent. The stannyl radical reacts with CH3CoBI2
to give homolytic cleavage of the Co-C bond and
produces CH3SnCI 3 (under reaction conditions
of high chloride concentrations) [17]. Other
methyl donors may also be important for extracetlular reactions. Dimethyl /3-propiothetin (DMPT)
is a common intracellular algal metabolite and a
(1)
3(CH3)3SnSSn(CH3) 3
h v ((CH3)2SnS)3 + 3(CH3)4S n
, 2R3SnCI + 2R2SnCI 2
CH 3
I
(2)
Tetramethyltin formation may therefore arise
from biotic and abiotic mechanisms. It is significant that in anaerobic sediment cultures, rapid
methyltin production was correlated with numbers of sulphate-reducing bacteria, e.g. Desulfovibrio, and sulphur-oxidizers; S z- reactions are
likely to be involved here, as with mercury [42].
Such disproportionation reactions (3) may complicate differentiation of abiotic and biotic mechanisms in tin cycling, as do various transmethyla-
S+
CH 3 ,
HO-,.
CH
/CH2
C"
II
O
Fig. 3. Structural formula of dimethyl/3-propiothetin (DMPT).
302
source of methyl halides released into the environment. Cellular DMPT (Fig. 3) reacts with
aqueous iodide to give CH3I at an annual bioproduction rate > 10~2g year-1. Methyl halides, such
as CH3I, can react with SnS (or other metal
elements in minerals or wastes) to yield methyltin
triiodide [20,21,45]:
SnS + 3CH3I
H:O
H~As(V)O,
~,
O:
:As(III)O(OH):~
arsenite
RR'SCIt~"
RR'S
CH~AsCO)(OH):
, CH3SnI 3 + (CH3)2S
Terrestrial wood decay fungi also form CH3I,
and other alkyl halides, which may have a methytatory function [46-48].
Arsenic biomethylation. Original work by Challenger [49] revealed that trimethylarsine was produced by several fungi, e.g. Scopulariopsis brevicaulus, from arsenite. Subsequent work confirmed that other fungi such as Penicillium sp.,
Gliocladium roseum and the yeast Candida humicola were also capable of arsenic biomethylation
[6,50-53]. Methylation occurs by transfer of carbonium ions (CH~) from S-adenosylmethionine
(Fig. 4). Arsenic biomethylation under aerobic
conditions is not restricted to fungi and yeasts.
Bacteria, including Aeromonas sp., Flavobacterium sp. and Escherichia coli can produce a
variety of methylated arsenic derivatives including dimethyl- and trimethylarsine. Methylated arsenic compounds are widely found in aquatic
habitats suggesting that arsenic biomethylation is
a significant process in the environment [52]. The
relative involvement of soil, sediment or aquatic
organisms is not, however, fully ascertained. Reaction rates vary depending on environmental
parameters, e.g. pH, and the organisms involved,
and not all products may be formed [6]. Under
anaerobic conditions, methylation stops at
dimethylarsine which is rapidly oxidized under
aerobic conditions [50].
Arsenic biomethylation appears widespread in
higher animals, including mammals, although
there is uncertainty regarding the possible involvement of intestinal microflora in this process.
Some experiments with 'germ-free' mice suggested that the microflora may have little or no
role in the process, with methylation resulting
from biochemical activities of the animal. However, in vitro experiments with rats showed that
methylarsenicals were only generated by caecal
arsenate
methylarsonic acid
O:
[Ctt,As(OH):I
i ~ RR'SCH~'
RR'S
[(CH~):As(:OI(OH)," ]
(CH0.~As(:O)(OH)
,
cacodylic acid
O~
[(CH,):AsOH]
IHI
+ (CH~):AsH
RR'SCH,'
RR'S
(CH,)~AsO
O:
(CH,LAs,
trimethylarsine
f- RR'SCH,'
v~" RR'S
[(CH0,As']
Fig. 4. Outline scheme for biomethylation of arsenic. SAdenosylmethionine is represented as RR'SCH~. Original
scheme based on work of Challenger [49] and adapted from
Thayer [1,6] and Thayer and Brinckman [18] with permission.
Additional reactions (dashed lines) proposed by Thayer and
Brinckman [18].
contents while in fish, intestinal microorganisms
were implicated in organoarsenical production
[6,501.
Selenium biomethylation. Biomethylation of selenium appears to be by a similar mechanism to
that for arsenic. Several fungi and bacteria can
produce dimethylselenide from inorganic selenium compounds [54,55] while rats fed selenate
or selenite salts exhale dimethylselenide and
dimethyldiselenide and excrete trimethylselenonium ion in urine [56]. The production of dimethylselenide and dimethyldiselenide from soil and
303
:Se(:O)(OH)O~ RR'SCH~"
RR'S
[CH3SeO,H]
OH
IHI
[CH~SeH
[CH,SeO;]
1ol or CH~SeOH]
~ RR'SCH~"
RR'S
(CH~)2SeO2
02
[(CH,)2SeOI
, CH,SeSeCH~
dimethyldiselenide
dimethylselenone
2e
02
~J~
~ RR'SCH3"
* RR'S
(CH,)2Se:
2(CH3)3Pb + + Pb
dimethylselenide
[(CH,hS~-OI
[(CH~)jSe]"
Pb(IV) oxides with CH~CoBI2 has been recorded,
although (CH3)4Pb was not produced after
attempted methylation of (CH3)3Pb ÷ or
(CH3)2Pb 2+ under conditions suitable for Hg(II)
methylation [22]. Iodomethane (methyl iodide)
may be involved in lead methylation and the
formation of (CH3)2Pb 2+ and (CH3)3Pb ÷ by reaction of CH3I with Pb(II) salts has been confirmed by several groups. If Pb(II) was reduced to
Pb(0), or if Pb(0) was used alone, CH3I methylation resulted in (CH3)4Pb as well as (CH3)3Pb +
and (CH3)2Pb 2" as products. (CH3)3Pb + may
disproportionate in the presence of Pb(0) [22,59]:
trimethylselenonium ion
Fig. 5. Outline scheme for biomethylation of selenium. SAdenosylmethionine is represented as RR'SCH~. Pathways
shown are as adapted from Reamer and Zoller [67] and
Thayer and Brinckman [18] with permission. Additional reactions (dashed lines) proposed by Thayer and Brinkman [18].
sediment samples amended with a variety of selenium compounds has been observed, as well as
the production of volatile dimethylselenone
[18,55,56]. The mechanism of selenium biomethylation is a methyl carbonium pathway and is
similar to that originally proposed by Challenger
[49,53] (Fig. 5). Other pathways for volatile selenium production by plants have been proposed
which involve selenomethionine and Se-methytselenomethionine intermediates [57,58].
Biomethylation of lead and other metal(loid)s.
Biomethylation of metal(loid)s other than those
discussed has received less attention. There does
not appear to be any conclusive evidence for
Pb(II) methylation by methylcobalamin in water
[59]. However, although reports from different
groups have been contradictory, the slow formation of CH3(CzH5)2PbCI by reaction of (C2H5) 2
PbCI 2 with CH3CoB12 has been reported [22].
The formation of tetramethyllead by reaction of
, (CH3)zPb2++ (CH3)4Pb
Tetramethyllead may also arise from disproportionation reactions of (CH3)3Pb + with S 2- in
sediments [22]:
2(CH3)3Pb++ S2((CH3)3Pb)2S
, ((CH3)3)Pb)2S
~ (CH3)aPb + (CH02PbS
It has been proposed, from measurements of Pb
concentrations in prehistoric snow and ice, that a
considerable proportion of excess Pb in the environment arises from biotic and abiotic aikylation
processes [59,60].
Microbial transformations of tellurium appear
analogous to those for selenium and include
methylation and reduction. Certain fungi and
bacteria can produce volatile dimethyl telluride
from tellurium salts [56,61].
There are examples of methyl-accumulating
reactions for a variety of other metals in the
literature and methylcobalamin has been implicated in several including TI, Pd, Pt, Au and Cr
[62]. However, the environmental significance of
such processes and the relative involvement of
microbial/abiotic components is unclear, while
results for other metals, e.g. Cd, Ag, Cu, are
often based on scant evidence. Inorganic and
organic Ge species exist in natural waters and,
although sources of methylgermanium are not
conclusively determined, biomethylation is an obvious suggestion [63]. There is evidence for methylation of Sb by marine algae (diatoms) (see later)
[64].
304
Biological importance of biomethylation
Biomethylation is frequently proposed as a
mechanism of metal(ioid) detoxification. For arsenic, methylarsenicats are less toxic and are excreted more readily and bind less to cell walls
and biomolecules than inorganic arsenic compounds [6]. Microbially produced trimethylarsine
(or dimethylarsine) diffuses through cell walls and
is lost to the atmosphere. Aquatic algae may
combine methylation with alkylation to form other
non-toxic organoarsenicals (see later). It should
be mentioned, however, that plasmid-mediated
arsenate resistance in E. coli and Staphylococcus
aureus depends on an effiux mechanism and not
methylation [65,66]. There is little or no detailed
information relating to the genetics and biochemical regulation of metal(loid) methylation in microorganisms. Selenium methylation has been
suggested as a detoxification mechanism since
dimethylselenide is a gas and may be lost from
the medium [67].
For mercury, a detoxification role is not so
clear as biomethylation results in the formation
of methylmercury which is usually considered to
be more toxic than Hg 2+ [68] although it may
diffuse out of cells faster than Hg 2÷ [20]. Furthermore, in most bacterial examples, Hg2+-resis tance involves reduction of Hg 2÷ to Hg ° which
volatilizes from the growth medium, while CH 3
Hg+-resistance involves enzymatic cleavage of the
C-Hg bond followed by Hg 2+ reduction to Hg °.
Not surprisingly, only a few examples exist which
suggest a detoxification role for Hg 2÷ methylation. A Hg2+-resistant strain of Neurospora crassa
produced more methylmercury than the parental
strain [69]. In Clostridium cochlearium, an auxotrophic mutant for vitamin Bt2 was more sensitive to Hg 2+ and did not methylate Hg 2÷ in
contrast to the B12-independent parent strain [70].
However, the parental strain only methylated
0.16% of the Hg 2+ after 48 h of growth. Hg z+
reduction to Hg ° is usually more rapid and
> 50% may be removed after 48 h growth [20].
Indeed, the often extremely low yield of methylated products is another reason for cautious
interpretation of biomethylation as a microbial
detoxification mechanism. Published % conversion values for Hg 2+ methylation range from
0.002-1.2% of added Hg 2+, while for SnC14,
methyltins arose at a conversion rate of 0.013%
after 60 days incubation [20]. Regardless of their
significance in detoxification, methylation reactions (biotic and abiotic) are important in the
natural environment since they alter the volatility,
solubility, toxicity and the mobility of metal(loid)s
[23,24]) and occupy key positions in biogeochemical cycles [62,71].
Bioalkylation
While it is predominantly methyl groups that
are transferred onto metal(loid)s, the transfer of
other alkyl groups has been reported, particularly
for arsenic. Arsenic bioalkylation occurs mainly
HzC-COR'
a
r
He-cow
(R = -H, -COOH)
O+
H~-O-P-O-CHCH~As(CI-tg~
O
CH
[ 3
O = As-CH~
R
O
OCH2CHOHCH2R
b
(R = -OH, -S03H)
OH
OH
CH3
I+
~O
H3C-As'CHz'C ~" OCH 3
CH3
d
I+
H3C-As-CH2-CH2OH
CH3
Fig. 6. Structural formulae of selected bioalkylated organoarsenic compounds. (a) arsenolipid (R = - H , -COOH); (b) arsenosugar (R = -OH, -SO3H); (c) arsenobetaine; (d) arseno choline.
305
in the marine environment with algae playing the
most important role. Arsenates (and arsenites)
may be taken up by algal cells (the marine As
concentration may approach 20 nmol dm -3 [50])
and converted to less toxic alkylarsenicals. In this
case, therefore, bioalkylation may act as a detoxification mechanism, and, as a consequence, inorganic arsenic usually comprises only a small proportion of total arsenic in algae. In some species,
up to 96% of cellular arsenic is present in organic
forms [50]. Alkylated arsenic compounds in algae
include arsenolipids and arsenosugars (Fig. 6).
Arsenobetaine is commonly found in marine
invertebrates and fish with arsenocholine occurring to a lesser degree (Fig. 6). Neither of these
are found in algae. The possible origin of arsenobetaine has been a controversial topic. It is generally accepted that this form of arsenic is accumulated via food chains rather than from the surrounding water which therefore implicates algae
since they are primary producers in the marine
environment [72]. It has been postulated that
arsenobetaine and arsenocholine are derived from
algal arsenosugars [50,73] with a trimethylarsonium derivative being a precursor [74]. Alternatively, arsenobetaine may also arise from the
degradation of As-phosphatidyl choline [73]. Arsenobetaine is apparently an innocuous arsenical
and resistant to metabolic degradation [74]. It is
likely that bioalkylation can occur for selenium,
and certain other elements, though detailed information is lacking.
Degradation and detoxification of organometal
(ioid)s
The ubiquitous mechanism of Hg 2+ resistance
in bacteria involves enzymatic reduction of Hg 2+
by cytoplasmic mercuric reductase (MR) to Hg °
which is of lower toxicity than Hg 2+, volatile and
rapidly lost from the environment [65,66,75,76].
organomercurials are enzymatically detoxit:cd by
organomercurial lyase (OL) which cleaves the
Hg-C bond of, e.g. methyl-, ethyl- and phenylmercury to form Hg 2+ and methane, ethane and
benzene, respectively. The resulting Hg 2+ is
volatilized by mercuric reductase:C6HsHg +
phenylmercury
OL > C 6 H 6 ], + Hg2+
benzene
CH3Hg +
methylmercury
Ot ~ CH 4 1' + Hg z+
methane
C2HsHg +
ethylmercury
Hg2+
oi. ~ C 2 H 6 1" + H g 2+
ethane
NADPH"~.e-a2NADP~ HgO
MR
In addition to the above mechanism, CO z has
been observed as a methylmercury demethylation
product in anaerobic lake sediments. Such an
oxidative demethylation process is now believed
Io widely occur in freshwater, estuarine and hypersaline systems under aerobic and anaerobic
conditions by a variety of bacteria including sulphate-reducers, methanogens and aerobes. It has
been proposed that the biochemical pathway of
demethylation is partly common to mechanisms
used for the metabolism of one-carbon compounds [77].
Several bacteria can demethylate organoarsenicals in aquatic and terrestrial environments,
e.g. Achromobacter, Flavobacteriurn, Nocardia,
Pseudornonas and Alcaligenes species. This may
be the most important loss mechanism for methylarsenicals from aerated soils and CO 2 and arsenate may be final products [50]. Methanogenic
and sulphate-respiring bacteria can demethylate
dimethylselenide to CH 4 and CO 2 [78].
Organotin degradation involves sequential removal of organic groups from the tin atom which
generally results in a reduction of toxicity [35,79]:
RaSn
> R3SnX
, RSnX 3
) R2SnX 2
> SnX 4
Breaking of Sn-C bonds can be accomplished
by abiotic and biotic mechanisms. Ultra-violet
(UV) irradiation and chemical cleavage are the
most significant abiotic mechanisms encountered
under usual environmental conditions and may
occur in aquatic and terrestrial ecosystems.
Degradation of methyltin chlorides, (CH3) ~
SnCi4_ . (n = 1-3) gives hydrated Sn(IV) oxide as
a final product [35]. It has been reported that
306
sunlight degradation of tributyltin oxide may have
a photolytic t~ of 18-89 days [23,80]. Regarding
chemical cleavage, mineral acids, alkalis and other
substances, may attack and break the Sn-C bond.
Transmethylation reactions with other metals, e.g.
Hg 2+, have been described previously. Gamma
(y) irradiation may degrade organotins which is
significant for y-irradiated food: organotin release may occur if organotin-stabilised PVC is
used as packaging [35].
It seems well established that microorganisms
are capable of organotin degradation [81]. Certain Gram-negative bacteria can dealkylate tributyltins [82] although uptake of the monovalent
tributyltin cation may occur without degradation
in some species [83]. Degradation of tributyltin
oxide and tributyltin napthenate, used as wood
preservatives, may be achieved by fungal action,
di- and monobutyltins resulting as breakdown
products [82,84]. Debutylation of tributyltin by
the green alga Ankistrodesmus falcatus gave di-,
monobutyl- and inorganic tin [85]. Soil microorganisms can slowly degrade triphenyltin acetate
to give similar end-products [86]. However, it
should be stressed that there is still a dearth of
information relating to biotic organotin degradation including the relationship between resistance
and degradative ability, the influence of the anionic radical, and the relative importance of microbial and abiotic degradation in natural habitats [23,35,79].
Dealkylation of organoleads, ultimately to inorganic lead, has been postulated as an important
part of the biogeochemical lead cycle [11,59].
Certain airborne organoleads, e.g. tetraethyllead,
are subject to photolytic cleavage of the Pb-C
bonds with the lead returning to the terrestrial
environment [1,59]. Such effects are less significant in aquatic and terrestrial ecosystems because
of poor light penetration.
As previously described for tin, metal-carbon
bonds may be susceptible to attack by environmental/chemical means without microbiological
intervention although in some circumstances, biological activity may enhance the environmental
conditions necessary for abiotic attack, e.g. alteration of pH, redox potential, etc. Some reactions
may result in disappearance of the organometal
(loid) from a given location; other reactions may
give rise to more toxic products and exacerbate
toxic effects. Mercury(II) salts can demethylate
other compounds:
HgCI 2 + (CH3)3SnCI
, CH3HgCI + (CH3)2SnCI z
Clearly, many mechanisms of organometal
(loid) degradation result in detoxification and occupy significant positions in the biogeochemical
cycles of the elements involved. It should also be
mentioned that a variety of additional microbial
mechanisms may be involved in microbial metal
and organometal(loid) detoxification and also that
toxicity can be greatly modified by physico-chemical environmental parameters [14,23,24,28,3133,87].
Microbial formation and transformation of organometal(Ioid)s as significant components of biogeochemical cycles for metals and metalloids
Environmental cycles of organometal(loid)
compounds depend on biotic and abiotic processes and may be greatly affected by anthropogenic processes such as industrial emissions
and pollution. The transformation processes previously discussed are significant in biogeochemical cycles as the toxicity and mobility of organometal(loid)s compounds may be altered. Many
organometal(loid)s are volatile which is important
for atmospheric dispersal but may also render
them subject to photolytic decomposition. The
lipid solubility of organometallic compounds also
plays a major role in environmental cycling since
many organometallic compounds can readily cross
biological membranes, a property that often underlies toxic symptoms [6].
Biological accumulation and excretion of organometal(loid)s
Organometal(loid) compounds can be accumulated by the biota in natural environments and
this can be accompanied by toxicity a n d / o r biological alteration or detoxification of such compounds [81]. Bioaccumulation therefore has a sig-
307
nificant influence on the mobility and bioavailability of organometal(ioid)s. A further consequence of bioaccumulation is the transfer of organometal(loid)s through food chains/webs.
When organisms at the lower levels of food chains,
e.g. algae and bacteria, accumulate toxicants, the
toxicant concentration may increase in predators
('biomagnification') and have serious effects on
organisms at the top of the chain, including humans [1,6].
In comparison with toxic metals, a negligible
amount of information on microbial accumulation of organometal(loid)s is available. It is known
that algae, e.g. Chlorella sp., Chaetoceros muelleri
and Phaeodactylum tricornutum, can accumulate
(and excrete) methylmercury compounds. Bacterial strains isolated from Chesapeake Bay had
bioaccumulation factors (BCF) of 350-850 for
tributyltin while methyltin-resistant isolates exhibited BCF values up to 220 for tributyltin and
SnC14 [8,83]. The algae Isochrysis galbana and
Ankistrodesmus falcatus had BCF values of 5500
[88] and 30000 [85]. Bioaccumulation of tin compounds has been detected in most groups of
organisms including fungi, higher algae, plants,
crustaceans and fish [6,8,89,90]. Most studies on
biomagnification have used methylmercuric compounds because of their ready accumulation and
subsequent retention by organisms [6]. In the
model food chain sequence involving the alga
Chlorella t,ulgaris, Daphnia magna (cladoceran),
Gambusia affinis (mosquito fish) and Salmo
gairdneri (rainbow trout), the overall rate of mercury transfer from water to the terrestrial consumer was about 15% [6]. Biomagnification is
especially significant in the aquatic environment
because of industrial discharge of pollutants.
Rather less work has considered terrestrial
food chains though it is appreciated that soil
microorganisms are involved in sequestration,
transformation a n d / o r mobilization of organometal(loid)s and influence their availabil!ty to
plants which are also capable of bioaccumulation
[58]. Grazing by herbivores may again lead to
biomagnification and ultimate transfer to humans. It should be noted that the physico-chemical properties of the environment will also influence the speciation and bioavailability of organo-
metal(loid)s. This is especially significant in soil
habitats where there may be significant retention
of organometal(loid)s on clays, colloids and other
constituents [6,7,33].
Although 'biomagnification' is a widely used
term, the concept can be misleading in certain
instances. Comparison of organometal(loid) concentrations in water with concentrations in
biomass may be questionable and, further, does
not allow for changes in speciation or excretion.
This is illustrated with organoarsenicals which
may be readily excreted after uptake a n d / o r formation. Thus, there does not appear to be biomagnification of arsenic in aquatic habitats, as
there is for mercury [1,6]. It is more realistic to
compare arsenic/phosphorus ratios at different
levels in the food chain, since it is the chemical
similarity between arsenate and phosphate that
initially leads to arsenate accumulation. The As : P
ratio decreases in the marine food chain which
demonstrates a progressive purification of the
phosphate pool and elimination of arsenic [50].
Organometal(loid) transformations in biogeochemical cycling of metal(loid)s
Mercury. Mercury cycling in the environment
has received considerable attention because of its
potency as an environmental pollutant. The aqueous environment is of central importance despite
most (approx. 98%) of the total mercury and
methylmercury being located in sediments [9]. As
described, interconversions between inorganic
and organic forms of mercury can be catalysed by
microbial activity a n d / o r abiotic mechanisms.
Most of the significant reactions occur in the
water column or in the surface layers of sediments with transfer between biotic and environmental components. Conversion of inorganic
mercury to methylmercury may be microbially
catalysed in the water (or intestinally in fish) or
may occur in the sediments prior to release into
the water. Distribution of inorganic and organic
forms of mercury is mainly by movement of suspended sediments and dissolved mercury with
some contribution made by mercury-accumulating biota [9]. In marine habitats, the estimated
total concentration of methylmercury, arising
from the methylation of inorganic mercury, in
308
open ocean waters is about 88 pg dm-3. Rates of
methylmercury release from sediments to water
are increased with rising temperature, total mer.'ury content or the addition of nutrients, presumably reflecting increases in microbial activity. A
reduction in pH also favours methylmercury release. The influence of acid rain may therefore
have important effects on the mobilization of
organic mercury and elevated methylmercury levels often occur in fish from lakes of low pH [9].
However, other studies have demonstrated decreased Hg 2+ methylation at reduced pH possibly meaning that other mechanisms such as altered partitioning of methylmercury between sediment and water may account for elevated levels
of methylmercury at low pH [26].
Sulphate deposition may stimulate methylmercury production by enhancing the activity of sulphate-reducing bacteria in sediments [27,91]. A
simplified biogeochemical cycle for mercury is
shown in Fig. 2. As mentioned previously the
microbial oxidation and reduction of inorganic
mercury, as well as breakdown of organomercurials, are important parts of this cycle. Detailed accounts of microbial mercury transformations are available [65,66,76,92-96].
Lead. The role of organolead compounds in
the environmental cycling of lead is still a matter
of debate [6,60]. A considerable input is via anthropogenic sources, e.g. vehicle emissions,
though it now seems clear that biomethylation
also plays a significant role in organolead cycling.
Anthropogenic sources
j
R,Pb
R3PbX
RzPbXz
" Pbi,o,z
/I
"
air
water
s2~
decay
~sediment
-" R4Pb
" R3PbX
" RzPbX2
/ / P i ( ~
b i ~- ' ' ~
(
~ PbS
S 2.
Me PbS
Me(Pb
J
Fig. 7. Simplified biogeochemical cycle for organolead compounds in the environment. An important role is attributed to
bioalkylation in this scheme. Also represented are lead methylation by methyl iodide (Mel) and sulphide-mediated disproportionation. Most environmentally significant organolead compounds are tetraalkylleads with R being a methyl or ethyl group and X being
the counter-ion, often halide. Modified from De Jonghe and Adams [60], with permission.
309
c o n c e r n b e c a u s e o f t h e e n t r y o f o r g a n o t i n s into
the e n v i r o n m e n t f r o m i n d u s t r i a l a n d a g r i c u l t u r a l
a p p l i c a t i o n s [8,10-13,36,98]. T h e l a r g e s t c o m m e r cial use o f o r g a n o t i n s is as P V C s t a b i l i z e r s a n d
such c o m p o u n d s m a y e n t e r the e n v i r o n m e n t as a
result o f l e a c h i n g / w e a t h e r i n g , l a n d b u r i a l a n d
incineration, although the majority of organotins
e n t e r i n g the e n v i r o n m e n t arise f r o m b i o c i d a l app l i c a t i o n s [3,23,35,79]. A n t i b i o l o g i c a l uses o f org a n o t i n s i n c l u d e a n t i h e l m i n t h i c s for poultry,
This m e a n s t h a t i n o r g a n i c l e a d in s e d i m e n t s m a y
b e r e m o b i l i z e d to r e - e n t e r t h e cycle as o r g a n i c
l e a d [59]. It s h o u l d be n o t e d t h a t d i s p r o p o r t i o n a tion a n d r e d i s t r i b u t i o n r e a c t i o n s a r e c h a r a c t e r i s tic o f a l k y l l e a d c o m p o u n d s a n d t h e s e m a t e r i a l s
m a y have a p o t e n t i a l role as i n t e r m e d i a t e s in t h e
f o r m a t i o n o f o t h e r l e a d derivatives [6,60,97]. A
s i m p l i f i e d b i o g e o c h e m i c a l cycle for o r g a n o l e a d s is
shown in Fig. 7.
Tin. T h e cycling o f tin is a t o p i c o f i n c r e a s i n g
A n t h r o p o g e n i c sources
f R , Sn ~
- R, S n .
f R3SnX ~
" x h$
fR2SnX_, ~ ' ~
H u m a n consumption
d ).
f RSnX3 ~ - . "d
_ "..
"x]
Fish, seafood etc.
:~n.*.,
_
R,SnX
air
- "
water
kl---~
( C H ) Sn 2'
RR
<. . . . .
SnS
."/c
I N
- SnX,
f~.x"
~
(C
r,.~,A,
,),
,..
f
(
i , / e \ ,",_~.
,),Sn
(
3)>
"x
\
H3 S
/
\
i
nX,
:
S 2....
sediment
((CH3),SnS)3 ~
((CH3)3Sn)2S
SnX:
SnX,
Fig. 8. Simplified biogeochemical cycle for organotins in the environment. At usual environmental pH values, organotins of general
formula R n S n X 4 _ n (n = 0 - 3 ) exist in aqueous solution as simple neutral hydroxides. Little is known of the effect of the anionic
radical (X) on breakdown. In the environment, organotins usually exist as, or are converted to, oxides, hydroxides, carbonates or
hydrated cations. As a result of biomethylation, methylstannanes ((CH3)nSnH4_ n) may also be produced [35]. Main reactions
detailed are (a) bioaccumulation; (b) deposition and/or release from biota on death or other processes; (c) organotin degradation
(biotic and/or abiotic); (d) photolytic degradation of organotins and resulting free radical production [171; (e) biomethylation; (f)
methyltin degradation (demethylation); (g) disprolx~rtionation reactions; (h) sulphide-mediated disproportionation of bis(trimethyltin) sulphide (i) SnS formation; (j) formation of methyl iodide by reaction of dimethyl fl-propiothetin (DMPT) (Fig. 3) with aqueous
iodide [20]; (k) CH31 methylation of SnX2; (1) oxidative methylation of SnS by CH3I to form methyltin triiodide; (m)
transmethylation between organotins and mercury providing a link with the mercury cycle (Fig. 2). Scheme constructed with
reference to a number of sources including Ridley et al. [17], Thayer [6l, Blunden and Chapman [35], Cooney [23] and Brinckman
and Olson [201.
310
fungicides, acaricides, disinfectants, marine antifouling paints, molluscicides and preservatives
of natural and synthetic materials, e.g. wood,
textiles, paper [2-4,8,12,13,23,24]. The application of such compounds can lead to contamination of aerial, terrestrial and aquatic habitats
[36,99]. With reference to airborne particulates,
tin is a major pollutant (as are lead and tellurium) [8] while aquatic systems are contaminated by run-off water and overspray and leaching from antifouling paints and preparations [3].
In soils and aquatic sediments, a considerable
proportion of organotins may be absorbed by the
constituent particles from which slow leaching is
possible. A simplified biogeochemical cycle for
organotins is shown in Fig. 8. Whether effected
by physical, chemical or biological action, environmental transformations of tin compounds can
alter their volatility, solubility and absorptivity,
and therefore their toxicity.
Arsenic. In relation to the formation of environmentally stable organic derivatives, arsenic has
an extensive chemistry and has been widely studied [52]. The arsenic cycle in terrestrial organisms
is, as mentioned, characterized by simple methylated species (methanearsonic and dimethylarsinic acids, trimethylarsine oxide and methylated arsines). In aquatic systems, more complex
organoarsenic compounds dominate, and methylated compounds generally appear to be degra-
H ~ "CH~CI.~OH
~tlwnob~ine
CH,
3CH,
CH~-,As'CH,CO0"
CH3
3H"
H~.CH2CH20 R
k.~ ~.~
" (CH~),AsCHiCH,OR ~
arsenobetaine
As-phosphatidylethanolamine
SAM
As-phosphatidyl choline
fOx.
I.f~
1~
'
/" O= ,As".CH, 0
N.C~c.Nx~
,
CH3 x ' ~
o
"N''N"
CH~O-P-O-C-C-As'-CH2 0 I
/
ell,
"7/
OH
OH, H H CH3 Hi('H H'~
I
H C - - O - C - - R2
I
~
CH,-As'CH~CH,OH
OC-'H,CHOHCH,R
)_/
/
R= SOiH
arscnocholinc
/
OH
R= OH
/Red/M.,h
R= OSOH
-
--c..-.o'~.---"--..~
O
R= O-P-OCH,CH(OH)CH~OH
OH OH
4
,CH,
0 =,~ICH=CHzOH
CH3
dimethyloxarsylcthanol
o
]
CH~-O-C --R,
O~
arsenoribosidcs
o
~
c.,
O-AsCH,COOH
Meth.
Red./Metl~
CH,
HAsO~
~- CHraLsO(OH),
l (CH3),AsOOH e
l
antenite
monomethylarsonic acid
dimethylat~inicacid
/ Red./Meth.
dimethyloxarsylacetic acid
Red.
~o~
arsenate
1 Red./Meth.
v
(CH,),~ •
trimethylarsine
Red.
(C~,~-O
~~
trimethylamineo x i d e
c~. ,~'c~coo
CH,
amcnobctainc
Fig. 9. Postulated biological transformations and biogeochemical cycling of arsenic (after Maher [73], with permission). Spatial
separation of main reactions in the biosphere is not represented since several reactions may occur in both aquatic and terrestrial
habitats, while certain areas remain a subject of controversy, e.g. doubts over the environmental significance of arsenic
volatilization [50]. Two postulated routes for arsenobetaine formation are shown [73]. Note that methylarsenicals, used as
herbicides, may be a significant anthropogenic input into the cycle [6]. See text for details of organisms involved.
311
dation products of more complex compounds [50].
The application of methylarsenicals as herbicides
is an anthropogenic source of these materials in
the arsenic cycle although there is high retention
of arsenic species in soils. It therefore appears
that algal excretion of methylarsenic acids is the
main mechanism which leads to the presence of
organoarsenicals in natural waters [50]. Removal
processes include bacterial-mediated demethylation and oxidation of arsenite to arsenate. Arsenic speciation in water is therefore determined
by the balance between reduction and methylation by marine algae and by bacterial demethylation. It has been suggested that arsenic methylation in the terrestrial environment may represent
a significant loss of arsenic to the atmosphere.
However, only a minor proportion of atmospheric
arsenic is gaseous and most of this appears to be
inorganic. Volatilization of arsenic from oceans
by biomethylation has also been dismissed as a
source of atmospheric arsenic [50]. A simplified
biogeochemical cycle for arsenic is shown in
Fig. 9.
Selenium, tellurium, germanium, thallium, antimony and manganese. Anthropogenic inputs of
selenium and tellurium into the environment can
result in adverse effects on the biota. Both elements have widespread commercial uses while
selenium can additionally arise in the environment as a result of the burning of fossil fuels [56].
Selenium has use in glass manufacture, electronics, photocopying, pigment production and as additives to poultry and other livestock feedstuffs,
antimicrobial and chemotherapeutic agents and
dietary supplements. Main commercial selenium
compounds are inorganic whereas organoselenium compounds can arise in organisms as the
protein derivatives selenocystine ((-SeCH2CH
(NH2)COOH) 2) and selenomethionine (CH 3
SeCH 2CH 2CH(NH 2)COOH) [56]. Tellurium has
applications in electronics and, increasingly, in
the semiconductor industry. Some microorganisms can reduce tellurite to elemental tellurium
which appears as a black deposit [100]. Analogous
reduction of selenate or selenite results in amorphous selenium formation which can impart a red
colour to colonies in pure culture due to deposition, chiefly in walls and membranes [61]. Dissim-
ilatory reduction of selenate to Se°(s) can occur
independently of sulphate; Se°(s) appears to be
the main form of selenium in salt-marsh sediments [10]. Ecological interrelationships and the
potential of such processes for bioremediation of
selenium oxyanion-contaminated wastewaters has
been described since the capacity for dissimilatory selenium reduction appears to be a widespread phenomenon [101-103]. Furthermore, microbial Se volatilization also appears to have potential for the decontamination of selenium-contaminated soils, sediments and waters [58,104]. It
therefore has applications for evaporation pond
disposal sites as well as drainage and other waters
[58,104]. The presence of naturally occurring Semethylating microorganisms in soils and waters
means that Se-volatilization can be significantly
enhanced by appropriate amendments, e.g. addition of fertilizers, nutrients, ploughing and irrigation, which stimulate the activity of microbial
populations [58]. Microbial volatilization of Se is
being applied to the removal of selenium from
the contaminated San Joaquin valley and Kesterton reservoir, CA, with the main parameters enhancing Se volatilization being carbon source,
aeration, moisture and elevated temperature [58].
The technology for selenium volatilization can be
operated through evaporation pond management
in series or by primary pond application [105].
Incoming water is evaporated to dryness until
sediment Se is around 100 mg kg -~ prior to
volatilization treatment. The pond is placed back
in operation when the Se content has been decreased to an acceptable level. Atmospheric Se
concentrations were considerably lower than recommended safety standards with the impact of
volatilized Se on neighbouring lands being minimal and, in fact, mainly blowing towards Se-deficient areas [58]. Future research is concerned
with the initiation of a full scale Se volatilization
operation for contaminated soils and sediments
[58].
Methylgermaniums are found in rainfall and
surface waters although their origin remains uncertain. Inorganic germanium enters the aquatic
environment as a result of natural leaching and
industrial effluents with some also being released
into the atmosphere from the burning of fossil
312
o
O-
H2~;O-- P--O~H 2
H
3 HO--C--H
O-+Sb --CH_
]
2
I
O.
OCH 2
H--C--O--CO--R
I
H2CO--CO--R'
CH3 ~
OH
OH
Fig. 10. Structural formula of the stibnolipid of the marine
diatom Thalassiosira nana (from Benson [64], with permission).
fuels. Little is known of the biogeochemical cycling of germanium.
Thallium has been investigated for methylation
reactions though there are no reports to date of
the occurrence of methylthallium species in the
environment, despite methyl species being less
toxic then T1 ÷ [5]. Methylantimony(V) species
can be detected in natural waters at the ng dm -3
level [7]. The occurrence of methyl forms is ascribed to biological activity and, as with arsenate,
a variety of antimony metabolites may be found
in marine algae such as protein-bound antimony
and a stibnolipid in the diatom T h a l a s s i o s i r a n a n a
[64] (Fig. 10). Algal metabolism of oceanic antimony species appears to parallel that of arsenate,
and the reduction, methylation, and adenosylation of antimony may serve to mediate detoxification and excretion. Some early work has indicated
possible antimony methylation by fungal cultures
though detailed information is lacking [5]. Since
there are similarities between arsenate and antimonate species, and similar concentrations of arsenic and antimony in aquatic habitats ( < ng g in unpolluted waters), it is thought that a similar
environmental chemistry may operate, though detailed biotic information on antimony is still
scarce.
Conclusions
The significance of microorganisms as important agents of biogeochemical change is appreciated in several areas of microbiology although
their involvement in organometal(loid) formation
and degradation, component parts of biogeochemical cycles for several metals and metalloids,
is less widely appreciated. Furthermore, anthropogenic entry of organometai(loid)s into the environment continues to increase with many being
accumulated by the biota and capable of causing
toxic effects. Those microbial transformations that
result in immobilization, volatilization a n d / o r
detoxification of organometal(loid)s may have potential for the bioremediation of contaminated
habitats. While Se has received most attention in
this regard in relation to reduction [101,103] and
biomethylation [58,104,105], a similar application
of biomethylation to As decontamination can also
be envisaged [51,52]. Biosorption, and those organometal(loid) degradative reactions which result
in the liberation of ionic metal species, may also
interact with other microbial biotechnologies for
the treatment of toxic metals and radionuclides
[106]. It is hoped that this review highlights the
global importance of microbial organometal(loid)
transformations and emphasises the need for work
on pure and applied aspects of organometal
(loid)-microbe interactions.
References
1 Thayer, J.S. (1988) Organometallic Chemistry, An
Overview. VCH Verlagsgesellschaft,Weinheim.
2 Davies, A. and Smith, P.J. (1980) Recent advances in
organotin chemistry. Adv. Inorg. Chem. Radiochem. 23,
1-77.
3 Blunden, SJ., Hobbs, L.A. and Smith, P.J. (1984) The
environmental chemistry of organotin compounds. In:
Environmental Chemistry (H.J. Bowen, Ed.), pp. 49-77.
Royal Society of Chemistry, London.
4 Thompson, J.AJ., Sheffer, M.G., Pierce, R.C., Chau,
Y.K., Cooney, J.J., Cullen, W.R. and Maguire, J.R. (1985)
Organotin compounds in the aquatic environment. NRCC
Association Committee for Scientific Criteria Environmental Quality Publication No. NRCC 22494, National
Research Council of Canada, Ottawa, Ontario.
5 Craig, P.J. (1986) Organometallic Compounds in the Environment. Longman, Harlow.
6 Thayer, J.S. (1984) OrganometaUic Compounds and Living Organisms. Academic Press, New York, NY.
7 Craig, PJ. (1986) Occurrence and pathways of organometallic compounds in the environment - general considera-
313
tions. In: organometallic Compounds in the Environment
(P.J. Craig, Ed.), pp. 1-64. Longman, Harlow.
8 Cooney, J.J. and Wuertz, S. (1989) Toxic effects of tin
compounds on microorganisms. J. Ind. Microbiol. 4, 375402.
9 Craig, P.J. (1986) Organomercury compounds in the environment. In: Organometallic Compounds in the Environment (P.J. Craig, Ed.), pp. 65-110. Longman, Harlow.
10 Hall, L.W. and Pinkney, A.E. (1985) Acute and sublethal
effects of organotin compounds on aquatic biota: an
interpretative literature evaluation. CRC Crit. Rev. Toxicol. 14, 159-209.
11 Chau, Y.K. (1986) Occurrence and speciation of organometallic compounds in freshwater systems. Sci. Total
Environ. 49, 305-323.
12 Saxena, A.K. (1987) Review: organotin compounds toxicology and biomedicinal applications. Appl. Organometal. Chem. 1, 39-56.
13 Maguire, R.J. (1987) Environmental aspects of tributyltin.
Appl. organometal. Chem. 1,475-498.
14 Gadd, G.M. (1992) Molecular biology and biotechnology
of microbial interactions with organic and inorganic heavy
metal compounds. In: Molecular Biology and Biotechnology of Extremophiles (R.A. Herbert and R.J. Sharp,
Eds.), pp. 225-257. Blackie and Son, Glasgow.
15 Williams, R.J.P. (1988) The transfer of methyl groups: a
general introduction. In: The Biological Alkylation of
Heavy Elements (P.J. Craig and F. Glockling, Eds.), pp.
5-19. Royal Society of Chemistry, London.
16 Wood, J.M. (1974) Biological cycles for toxic elements in
the environment. Science 183, 1049-1052.
17 Ridley, W.P., Dizikes, L.J. and Wood, J.M. (1977)
Biomethylation of toxic elements in the environment.
Science 197, 329-332.
18 Thayer, J.S. and Brinckman, F.E. (1982) The biological
methylation of metals and metalloids. Adv. Organometal.
Chem. 20, 313-356.
19 Thayer, J.S. (1989) Methylation: its role in the environmental mobility of heavy elements. Appl. Organometal.
Chem. 3, 123-128.
20 Brinckman, F.E. and Olson, G.J. (1988) Global biomethylation of the elements: its role in the biosphere translated to new organometallic chemistry and biotechnology.
In: The Biological Alkylation of Heavy Elements (P.J.
Craig and F. Glockling, Eds.), pp. 168-196. Royal Society
of Chemistry, London.
21 Brinckman, F.E., Olson, G.J. and Thayer, J.S. (1985)
Biological mediation of marine metal cycles: the case of
methyl iodide. In: Marine and Estuarine Chemistry (A.C.
Sigleo and A. Hattori, Eds.), pp. 227-238. Lewis Publishers, Chelsea, MI.
22 Rapsomanikis, S. and Weber, J.H. (1986) Methyl transfer
reactions of environmental significance involving naturally occurring and synthetic reagents. In: Organometallic
Compounds in the Environment (P.J. Craig, Ed.), pp.
279-307. Royal Society of Chemistry, London.
23 Cooney, J.J. (1988) Microbial transformations of tin and
tin compounds. J. Ind. Microbiol. 3. 195-204.
24 Cooney, J.J. (1988) Interactions between microorganisms
and tin compounds. In: The Biological Alkylation of
Heavy Elements (P.J. Craig and F. Glockling, Eds.), pp.
92-104. Royal Society of Chemistry, London.
25 Korthals, E.T. and Winfrey, M.R. (1987) Seasonal and
spatial variations in mercury methylation and demethylation in an oligotrophic lake. Appl. Environ. Microbiol. 53,
2397-2404.
26 Steffan, R.J., Korthals, E.T. and Winfrey, M.R. (1988)
Effects of acidification on mercury methylation, demethylation and volatilization in sediments from an acid-susceptible lake. Appl. Environ. Microbiol. 54, 2003-2009.
27 Gilmour, C.C., Henry, E.A. and Mitchell, R. (1992) Sulfate stimulation of mercury, methylation in freshwater
sediments. Environ. Sci. Technol. 26, 2281-2287.
28 Gadd, G.M. and Griffiths, A.J. (1978) Microorganisms
and heavy metal toxicity, Microb. Ecol. 4, 303-317.
29 Babich, H. and Stotzky, G. (1980) Environmental factors
that influence the toxicity of heavy metal and gaseous
pollutants to microorganisms. CRC Crit. Rev. Microbiol.
8, 99-145.
30 Craig, P.J. and Moreton, P.A. (1985) The role of speciation in mercury methylation in sediments and waters.
Environ. Pollut. 10, 141-158.
31 Hughes, M.N. and Poole, R.K. (1991) Metal speciation
and microbial growth - - the hard (and soft) facts. J. Gen.
Microbiol. 137, 725-734.
32 Gadd, G.M. (1990) Metal tolerance. In: Microbiology of
Extreme Environments (C. Edwards, Ed.), pp. 178-210.
Open University Press, Milton Keynes.
33 Jackson, T.A. (1989) The influence of clay minerals,
oxides, and humic matter on the methylation and
demethylation of mercury by micro-organisms in freshwater sediments. Appl. Organometal. Chem. 3, 1-30.
34 Hallas, L.E., Means, J.C. and Cooney, J.J. (1982) Methylation of tin by estuarine microorganisms. Science 215,
1505-1507.
35 Blunden, S.J. and Chapman, A. (1986) Organotin compounds in the environment. In: Organometallic Compounds in the Environment (P.J. Craig, Ed.), pp. 111-159.
Longman, Harlow.
36 Makkar, N.S., Kronick, A.T. and Cooney, J.J. (1989)
Butyltins in sediments from Boston Harbour, USA.
Chemosphere 18, 2043-2050.
37 Brinckman, F.E. and Iverson, W.P. (1975) Chemical and
bacterial cycling of heavy metals in the estuarine system.
In: Marine Chemistry in the Coastal Environment (T.M.
Church, Ed.), pp. 314-342. American Chemical Society,
Washington, D.C.
38 Jackson, J.A., Blair, W.R., Brinckman, F.E. and lverson,
W.P. (1982) Gas-chromatographic speciation of methylstannanes in the Chesapeake Bay using purge and trap
sampling with a tin-selective detector. Environ. Sci. Technol. 16, 110-119.
314
39 Guard, H.E., Cobet, A.B. and Coleman, W.M. (1981)
Methylation of trimethyltin compounds by estuarine sediments. Science 213, 770-771.
40 Gilmour, C.C., Tuttle, J.H. and Means, J.C. (1987)
Anaerobic microbial methylation of inorganic tin in estuarine sediment slurries. Microb. Ecol. 14, 233-242.
41 Craig, P.J. and Rapsomanikis, S. (1984) Formation of
tetramethyltin from trimethyltin precursors in sediment
environments. Environ. Technol. Lett. 5, 407-416.
42 Gilmour, C.C., Tuttle, J.H. and Means, J.C. (1985) Tin
methylation in sulphide bearing sediments. In Marine
and Estuarine Geochemistry (A.C. Sigles and A. Hattori,
Eds.), pp. 239-258. Lewis Publishers, Chelsea, MI.
43 Compeau, G. and Bartha, G. (1985) Sulfate-reducing
bacteria: principal methylators of mercury in anoxic estuarine sediment. Appl. Environ. Microbiol. 50, 498-502.
44 Maguire, R.J. (1984) Butyltin compounds and inorganic
tin in sediments in Ontario. Environ. Sci. Technol. 18,
291-294.
45 Manders, W.F., Olson, G.J., Brinckman, F.E. and Bellama, J.M. (1984) A novel synthesis of methyltin triodide
with environmental implications. J. Chem. Soc. Chem.
Commun. 8, 538-540.
46 Harper, D.B. (1985) Halomethane from halide ion - - a
highly efficient fungal conversion of environmental significance. Nature 315, 55-57.
47 Harper, D.B. and Kennedy. J.T. (1986) Effect of growth
conditions on halomethane production by Phellinus
species: biological and environmental implications. J.
Gen. Microbiol. 132, 1231-1246.
48 Harper, D.B. and Hamilton, J.T.G. (1988) Biogenesis of
halomethanes by fungi, ln: The Biological Alkylation of
Heavy Elements (P.J. Craig and F. Glockling, Eds.), pp.
197-200. Royal Society of Chemistry, London.
49 Challenger, F. (1945) Biological methylation. Chem. Rev.
36, 315-361.
50 Andreae, M.O. (1986) Organoarsenic compounds in the
environment. In: Organometallic Compounds in the Environment (P.J. Craig, Ed.), pp. 198-228. Longman, Harlow.
51 Davis Huysmans, K. and Frankenberger, W.T. (1991)
Evolution of trimethylarsine by a Penicillium sp. isolated
from agricultural evaporation pond water. Sci. Total.
Environ. 105, 13-28.
52 Tamaki, S. and Frankenberger, W.T. (1992) Environmental biochemistry of arsenic. Rev. Environ. Contam. Toxicol. 124, 79-110.
53 Challenger, F. (1978) Biosynthesis of organometallic and
organometalloidal compounds. In: Organometals and
Organometalloids. Occurrence and Fate in the Environment (F.E. Brinckman and J.M. Bellama, Eds.), pp. 1-22.
American Chemical Society, Washington D.C.
54 Thompson-Eagle, E.T. and Frankenberger, W.T. (1991)
Selenium biomethylation in an alkaline, saline environment. Water Res. 25, 231-240.
55 Thompson-Eagle, E.T., Frankenberger, W.T. and Karl-
son, V. (1989) Volatilization of selenium by Alternaria
alternata. Appl. Environ. Microbiol. 55, 1406-1413.
56 Chau, Y.K. and Wong, P.T.S. (1986) Organic Group V1
elements in the environment. In: Organometallic Compounds in the Environment (P.J. Craig, Ed.), pp. 254-278.
Longman, Harlow.
57 Lewis, B.A.G. (1976) Selenium in biological systems and
pathways for its volatilization in higher plants. In: Environmental Biochemistry, Vol. 1. Carbon, Nitrogen, Phosphorus, Sulphur and Selenium Cycles (J.O. Nriagu, Ed.),
pp. 389-409. Ann Arbor Sci., MI.
58 Thompson-Eagle, E.T. and Frankenberger, W.T. (1992)
Bioremediation of soils contaminated with selenium. Adv.
Soil Sci. 17, 261-310.
59 Hewitt, C.N. and Harrison, R.M. (1986) Organolead
compounds in the environment. In: Organometallic Compounds in the Environment (P.J. Craig, Ed.), pp. 160-197.
Longman, Harlow.
60 De Jonghe, W.R.A. and Adams, F.C (1986) Biogeochemical cycling of organolead compounds. Adv. Environ. Sci.
Technol. 17, 561-594.
61 Konetzka, W.A. (1977) Microbiology of metal transformations. In: Microorganisms and Minerals (E.D. Weinberg, Ed.), pp. 317-342. Marcel Dekker, New York, NY.
62 Wood, J.M. and Wang, K.-K. (1983) Microbial resistance
to heavy, metals. Environ. Sci. Technol. 17, 582-590.
63 Lewis, B.L., Andreae. M.O. and Froelich, P.N. (1988)
The biogeochemistry of methylgermanium species: a review. In: The Biological Alkylation of Heavy Elements
(P.J. Craig and F. Glockling, Eds.), pp. 77-91. Royal
Society of Chemistry, London.
64 Benson, A.A. (1988) Antimony metabolites in marine
algae. In: The Biological Alkylation of Heavy Elements
(P.J. Craig and F. Glockling, Eds.), pp. 135-137. Royal
Society of Chemistry, London.
65 Silver, S. and Walderhaug, M. (1992) Ion transport. In:
Encyclopedia of Microbiology, Vol. 2 (J. Lederberg, Ed.),
pp. 549-560. Academic Press, San Diego, CA.
66 Misra, T.K. (1992) Heavy metals, bacterial resistances. In:
Encyclopedia of Microbiology, Vol. 2 (J. Lederberg, Ed.),
pp. 361-369. Academic Press, San Diego, CA.
67 Reamer, D.C. and Zoller, W.H. (1980) Selenium
biomethylation products from soil and sewage sludge.
Science 208, 500-502.
68 Trevors, J.T. (1986) Mercury methylation by bacteria. J.
Basic Microbiol. 8, 499-504.
69 Landner L. (1971) Biochemical model for the biologfcal
methylation of mercury suggested from methylation studies in vivo with Neurospora crassa. Nature 230, 452-453.
70 Pan-Hou, H.S. and Imura, N. (1982) Involvement of
mercury methylation in microbial mercury detoxification.
Arch. Microbiol. 131, 176-177.
71 Ochiai, E.-I. (1987) General Principles of Biochemistry of
the Elements. Plenum Press, New York, NY.
72 Edmonds, J.S. and Francesconi, K.A. (1988) The methylation of arsenic by marine macro-algae. In: The Biologi-
315
cal Alkylation of Heavy Elements (P.J. Craig and F.
Glockling. Eds.), pp. 138-141. Royal Society of Chemistry, London.
73 Maher, W.A. (1988) Arsenic in the marine environment
of South Australia. In: The Biological Alkylation of Heavy
Elements (P.J. Craig and F. Glockling, Eds.), pp. 120-126.
Royal Society of Chemistry, London.
74 Benson, A.A. (1988) Biological methylation of the algal
trialkylarsine oxides. In: The Biological Alkylation of
Heavy Elements (P.J. Craig and F. Glockling, Eds.), pp.
127-131. Royal Society of Chemistry, London.
75 Belliveau, B.H. and Trevors, J.T. (1989) Mercury resistance and detoxification in bacteria. Appl. Organometal.
Chem. 3, 283-294.
76 Brown, N.L., Lund, P.A. and Ni Bhriain, N. (1989) Mercury resistance in bacteria. In: The Genetics of Bacterial
Diversity (D.A. Hopwood and K.F. Chater, Eds.), pp.
175-195. Academic Press, London.
77 Oremland, R.S., Culbertson, C.W. and Winfrey, M.R.
(1991) Methylmercury decomposition in sediments and
bacterial cultures: involvement of methanogens and sulfate reducers in oxidative demethylation. Appl. Environ.
Microbiol. 57, 130-137.
78 Oremland, R.S. and Zehr, J.P. (1986) Formation of
methane and carbon dioxide from dimethylselenide in
anoxic sediments and by a methanogenic bacterium. Appl.
Environ. Microbiol. 52, 1031-1036.
79 Blunden, S.J. and Chapman, A.H. (1982) The environmental degradation of organotin compounds - - a review.
Environ. Technol. Lett. 3, 267-272.
80 Maguire, R.J., Carey, J.H. and Hale, E.J. (1983) Degradation of the tri-n-butyltin species in water. J. Agric.
Food Chem. 31, 1060-1065.
81 Brinckman, F.E., Olson, G.J., Blair, W.R. and Parks, E.J.
(1988) Implications of molecular speciation and topology
of environmental metals: uptake mechanisms and toxicity
of organotins. In: Aquatic Toxicology and Hazard Assessment, Vol. 10 (W.J. Adams, G.A. Chapman and W.G.
Landis, Eds.), pp. 219-232. American Society for Testing
and Materials, Philadelphia, PA.
82 Barug, D. (1981) Microbial degradation of bis(tributyltin)
oxide. Chemosphere 10, 1145-1154.
83 Blair, W.R., Olson, G.J., Brinckman, F.E. and Iverson,
W.P. (1982)Accumulation and fate of tri-n-butyltin cation
in estuarine bacteria. Microb. Ecol. 8, 241-251.
84 Orsler, R.J. and Holland. G.E. (1982) Degradation of
tributyltin oxide by fungal culture filtrates. Int. Biodeteriot. Bull. 18, 95-98.
85 Maguire, R.J., Wong, P.T.S. and Rhamey, J.S. (1984)
Accumulation and metabolism of tri-n-butyltin cation by
a green alga Ankistrodesrnusfalcatus. Can. J. Fish. Aquat.
Sci. 41,537-540,
86 Barnes, R.D., Bull, A.T. and Poller, R.C. (1973) Studies
on the persistence of the organotin fungicide fentin acetate (triphenyltin acetate) in the soil and on surfaces
exposed to light. Pestic. Sci. 4, 305-317.
87 Laurence, O.S., Cooney, J.J. and Gadd, G.M. (1989)
Toxicity of organotins towards the marine yeast Debaryomyces hansenii. Microb. Ecol. 17. 275-285.
88 Laughlin, R.B. (1986) Bioaccumulation of tributyltin: the
link between environment and organism. In: Oceans 86
Proceedings, Volume 4, pp. 1206-1209. 1EEE Service
Center, Piscataway, NJ.
89 McDonald, L. and Trevors, J.T. (1988) Review of tin
resistance, accumulation and transformations by microorganisms. Water Air Soil Pollut. 40, 215-221.
90 Gadd, G.M., Gray, D.J. and Newby, P.J. (1990) Role of
melanin in fungal biosorption of tributyltin chloride. Appl.
Microbiol. Biotechnol. 34, 116-121.
91 Gilmour, C.C. and Henry, E.A. (1991) Mercury methylation in aquatic systems affected by acid deposition. Environ. Pollut. 71, 131-169.
92 Jeffries, T.W. (1982) The microbiology, of mercury. Progr.
Ind. Microbiol. 16, 21-75.
93 Robinson, J.B. and Tuovinen, O.H. (1984) Mechanisms
of microbial resistance and detoxification of mercury and
organomercury compounds: physiological, biochemical
and genetic analyses. Microbiol. Rev. 48, 95-124.
94 Foster, T.J. (1987) Genetics and biochemistry of mercury
resistance. CRC Crit. Rev. Microbiol. 15, 117-140.
95 Silver, S. and Misra, T.K. (1988) Plasmid-mediated heavy
metal resistances. Ann. Rev. Microbiol. 42, 717-743.
96 Silver, S., Laddaga, R.A. and Misra, T.K. (1989) Plasmiddetermined resistance to metal ions. In: Metal-Microbe
Interactions (R.K. Poole and G.M. Gadd, Eds.). pp.
49-63. IRL Press, Oxford.
97 Macaskie, L.E. and Dean, A.C.R. (1987) Trimethyllead
degradation by an alkyllead-tolerant yeast. Environ.
Technol. Lett. 8, 635-640.
98 Boyer, I.J. (1989) Toxicity of dibutyltin, tributyltin and
other organotin compounds to humans and to experimental animals. Toxicology 55,253-298.
99 Hallas, L.E. and Cooney, J.J. (1981) Tin and tin-resistant
microorganisms in Chesapeake Bay. Appl. Environ. Microbiol. 41,466-471.
100 Smith, D.G. (1974) Tellurite reduction in Schizosaccharornyces pornbe. J. Gen. Microbiol. 83. 389-392.
101 Oremland, R.S., Hollibaugh, J.T., Maest, A.S., Presser,
T.S., Miller, L.G. and Culbertson, C.W. (1989) Selenate
reduction to elemental selenium by anaerobic bacteria in
sediments and culture: biogeochemical significance of a
novel, sulphate-independent respiration. Appl. Environ.
Microbiol. 55, 2333-2343.
102 Oremland, R.S., Steinberg, N.A., Maest, A.S., Miller,
L.G. and Hollibaugh, J.T. (1990) Measurement of in situ
rates of selenate removal by dissimilatory bacterial reduction in sediments. Environ. Sci. Technol. 24, 1157-1164.
103 Steinberg, N.A. and Oremland, R.S. (1990) Dissimilatory
selenate reduction potentials in a diversity of sediment
types. Appl. Environ. Microbiol. 56, 3550-3557.
104 Frankenberger, W.T. and Karlson, U. (1991) Bioremediation of seleniferous soils. In: On-site Reclamation (R.E.
316
Hinchee and R.F. Olfenbuttel. Eds.), pp 239-254. Butterworth-Heinemann, Stoneham, MA.
105 Thompson-Eagle, E.T., Frankenberger, W.T. and Longley, K.E. (1991) Removal of selenium from agricultural
drainage water through soil microbial transformations.
In: The Economics and Management of Water and
Drainage in Agriculture (A. Dinar and D. Zilberman,
Eds.), pp. 169-186. Kluwer Academic Publishers, Norwell, MA.
106 Gadd, G.M. (1992) Microbial control of heavy metal
pollution. In: Microbial Control of Pollution (J.C. Fry,
G.M. Gadd, R.A. Herbert, C.W. Jones and I. WatsonCraik, Eds.), pp. 59-88. Cambridge University Press,
Cambridge.