Can CBD Help With Psoriasis?

Psoriasis is a common skin condition which accelerates the life cycle of skin cells(1). It causes cells to build up fast on the skin surface. The extra skin cells create red patches and scales that are itchy and often painful.

Psoriasis is a chronic disease that sometimes comes and goes. The primary objective of a psoriasis treatment plan is to prevent skin cells from overgrowing.

There is no cure for psoriasis. However, its symptoms are manageable. Lifestyle changes, such as frequent moisturizing of the skin, quitting smoking, and managing stress, may help alleviate symptoms of psoriasis.

According to the American Academy of Dermatology Association, stress is a common psoriasis trigger that can cause psoriasis flare-ups(2). 

Why People Are Using CBD for Psoriasis 

Research has shown that cannabidiol (CBD), which can come in the form of CBD products, such as oils or tinctures, balms, patches, face masks, creams, or salves, possesses health benefits. 

With CBD’s numerous purported benefits, there has been much interest in whether or not this cannabinoid can also help with skin issues linked to psoriasis, like inflammation, itch, dryness, and pain.

Some studies have been encouraging, including one study published in the Journal of Dermatological Science, which found that cannabinoids, such as CBD, slowed down the cell growth and division of keratinocytes involved in psoriatic skin rashes(3).

Keratinocytes are cells of the epidermis (outermost layer of the skin) that produce keratin, a type of protein that makes up the hair, skin, and nails.

Psoriasis is an inflammatory disease also characterized by the overproliferation of epidermal keratinocyte.

The significant purported benefits of CBD in skin care seem to be linked to its anti-inflammatory characteristics, as a 2017 study published in the Journal of the American Academy of Dermatology suggested(4).

In a study that examined cannabinoids, particularly CBD and THC (tetrahydrocannabinol), in the management of difficult-to-treat pain, the authors found that cannabinoid analgesics (pain-relievers) have generally been well-tolerated in clinical trials with acceptable adverse event profiles(5). 

CBD might also be useful in reducing itch and skin dryness, as shown in a 2017 review published in the Journal of the American Academy of Dermatology(6). 

Researchers of a 2019 study published in the journal Clinical Therapeutics examined individuals with skin diseases, either psoriasis, eczema, or scarring(7). 

Conducted by researchers in Italy, the study showed that applying a topical CBD-enriched ointment on psoriatic skin helped reduce psoriasis symptoms. 

Results suggested that CBD might help in hydrating and moisturizing the skin, enhancing skin elasticity.

The researchers concluded that the ointment was not only safe and effective but also improved the quality of life for individuals with psoriasis.

How CBD Oil Works to Help With Psoriasis 

A study published in the journal Trends in Pharmacological Sciences has suggested the existence of a functional endocannabinoid system (ECS) in the skin and implicated it in various biological processes(8). 

In the said study, the authors noted that the primary physiological function of the ECS in the skin seemed to be the regulation of the well-balanced proliferation, survival, and tolerance of skin cells. 

They said that the disruption of this delicate balance might facilitate the development of multiple skin problems, such as acne, seborrhea (red, itchy rash and white scales), allergic dermatitis, psoriasis (painful, dry, raised, and red skin lesions), and cancer.

The ECS, which is a network of cannabinoids and cannabinoid receptors, regulates most functions in the human body. The ECS also gathers and interprets signals from cannabinoids.

In one study published in the Journal of Dermatological Science, researchers isolated THC, CBD, and other cannabinoids from the cannabis plant(9).

The researchers found that, when applied to human skin cells, the cannabinoids inhibited the overproduction of keratinocytes commonly seen in psoriasis.

In addition to its impact on CB1 and CB2 receptors, there is evidence that CBD impacts the TRPV-1 and GPR55 receptors(10). Both receptors play a role in pain signaling and inflammation.


Studies have shown that CBD may help with psoriasis symptoms. However, more longitudinal research is needed to provide conclusive evidence of this cannabinoid’s efficacy. 

Also, the side effects of long-term CBD use is unknown, like the interactions that it may have with other pharmaceuticals currently taken.

Thus, before using CBD as supplemental psoriasis treatment, consult with a dermatologist experienced in cannabis use for advice.



  1. Mayo Clinic. (2019, March 13). Psoriasis. Retrieved from
  2. AAD. Are Triggers Causing Your Psoriasis Flare-Ups? Retrieved from
  3. Wilkinson JD, Williamson EM. Cannabinoids inhibit human keratinocyte proliferation through a non-CB1/CB2 mechanism and have a potential therapeutic value in the treatment of psoriasis. J Dermatol Sci. 2007;45(2):87–92. DOI:10.1016/j.jdermsci.2006.10.009.
  4. Mounessa BS et al. The role of cannabinoids in dermatology. Journal of the American Academy of Dermatology. Volume 77, Issue 1, July 2017, Pages 188-190.
  5. Russo EB. Cannabinoids in the management of difficult to treat pain. Ther Clin Risk Manag. 2008;4(1):245–259. doi:10.2147/tcrm.s1928.
  6. Mounessa J et al. The role of cannabinoids in dermatology. Journal of the American Academy of Dermatology, Volume 77, Issue 3, September 2017, Pages e87-e88.
  7. Palmieri B, Laurino C, Vadalà M. A therapeutic effect of cbd-enriched ointment in inflammatory skin diseases and cutaneous scars. Clin Ter. 2019 Mar-Apr;170(2):e93-e99. DOI: 10.7417/CT.2019.2116. DOI: 10.7417/CT.2019.2116.
  8. Bíró T, Tóth BI, Haskó G, Paus R, Pacher P. The endocannabinoid system of the skin in health and disease: novel perspectives and therapeutic opportunities. Trends Pharmacol Sci. 2009;30(8):411–420. DOI:10.1016/
  9. Wilkinson J, Williamson E.  Cannabinoids inhibit human keratinocyte proliferation through a non-CB1/CB2 mechanism and have a potential therapeutic value in the treatment of psoriasis. Journal of Dermatological Science Volume 45, Issue 2, February 2007, Pages 87-92.
  10. Sharir H, Abood ME. Pharmacological characterization of GPR55, a putative cannabinoid receptor. Pharmacol Ther. 2010;126(3):301–313. DOI:10.1016/j.pharmthera.2010.02.004; Costa B, Giagnoni G, Franke C, Trovato AE, Colleoni M. Vanilloid TRPV1 receptor mediates the antihyperalgesic effect of the nonpsychoactive cannabinoid, cannabidiol, in a rat model of acute inflammation. Br J Pharmacol. 2004;143(2):247–250. DOI:10.1038/sj.bjp.0705920.

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Gentamicin sulfate
1.1 Substance
1.2 Group
1.3 Synonyms
1.4 Identification numbers
1.4.1 CAS number
1.4.2 Other numbers
1.5 Brand names, Trade names
1.6 Manufacturers, Importers
2.1 Main risks and target organs
2.2 Summary of clinical effects
2.3 Diagnosis
2.4 First aid measures and management principles
3.1 Origin of the substance
3.2 Chemical structure
3.3 Physical properties
3.3.1 Properties of the substance
3.3.2 Properties of the locally available formulation
3.4 Other characteristics
3.4.1 Shelf-life of the substance
3.4.2 Shelf-life of the locally available formulation
3.4.3 Storage conditions
3.4.4 Bioavailability
3.4.5 Specific properties and composition
4.1 Indications
4.2 Therapeutic dosage
4.2.1 Adults
4.2.2 Children
4.3 Contraindications
5.1 Oral
5.2 Inhalation
5.3 Dermal
5.4 Eye
5.5 Parenteral
5.6 Other
6.1 Absorption by route of exposure
6.2 Distribution by route of exposure
6.3 Biological half-life by route of exposure
6.4 Metabolism
6.5 Elimination by route of exposure
7.1 Mode of action
7.1.1 Toxicodynamics
7.1.2 Pharmacodynamics
7.2 Toxicity
7.2.1 Human data Adults Children
7.2.2 Relevant animal data
7.2.3 Relevant in vitro data
7.3 Carcinogenicity
7.4 Teratogenicity
7.5 Mutagenicity
7.6 Interactions
7.7 Main adverse effects
8.1 Material sampling plan
8.1.1 Sampling and specimen collection Toxicological analyses Biomedical analyses Arterial blood gas analysis Haematological analyses Other (unspecified) analyses
8.1.2 Storage of laboratory samples and specimens Toxicological analyses Biomedical analyses Arterial blood gas analysis Haematological analyses Other (unspecified) analyses
8.1.3 Transport of laboratory samples and specimens Toxicological analyses Biomedical analyses Arterial blood gas analysis Haematological analyses Other (unspecified) analyses
8.2 Toxicological Analyses and Their Interpretation
8.2.1 Tests on toxic ingredient(s) of material Simple Qualitative Test(s) Advanced Qualitative Confirmation Test(s) Simple Quantitative Method(s) Advanced Quantitative Method(s)
8.2.2 Tests for biological specimens Simple Qualitative Test(s) Advanced Qualitative Confirmation Test(s) Simple Quantitative Method(s) Advanced Quantitative Method(s) Other Dedicated Method(s)
8.2.3 Interpretation of toxicological analyses
8.3 Biomedical investigations and their interpretation
8.3.1 Biochemical analysis Blood, plasma or serum Urine Other fluids
8.3.2 Arterial blood gas analyses
8.3.3 Haematological analyses
8.3.4 Interpretation of biomedical investigations
8.4 Other biomedical (diagnostic) investigations and their interpretation
8.5 Overall Interpretation of all toxicological analyses and toxicological investigations
8.6 References
9.1 Acute poisoning
9.1.1 Ingestion
9.1.2 Inhalation
9.1.3 Skin exposure
9.1.4 Eye contact
9.1.5 Parenteral exposure
9.1.6 Other
9.2 Chronic poisoning
9.2.1 Ingestion
9.2.2 Inhalation
9.2.3 Skin exposure
9.2.4 Eye contact
9.2.5 Parenteral exposure
9.2.6 Other
9.3 Course, prognosis, cause of death
9.4 Systematic description of clinical effects
9.4.1 Cardiovascular
9.4.2 Respiratory
9.4.3 Neurological CNS Peripheral nervous system Autonomic nervous system Skeletal and smooth muscle
9.4.4 Gastrointestinal
9.4.5 Hepatic
9.4.6 Urinary Renal Other
9.4.7 Endocrine and reproductive systems
9.4.8 Dermatological
9.4.9 Eye, ear, nose, throat: local effects
9.4.10 Haematological
9.4.11 Immunological
9.4.12 Metabolic Acid-base disturbances Fluid and electrolyte disturbances Others
9.4.13 Allergic reactions
9.4.14 Other clinical effects
9.4.15 Special risks
9.5 Other
9.6 Summary
10.1 General principles
10.2 Relevant laboratory analyses
10.2.1 Sample collection
10.2.2 Biomedical analysis
10.2.3 Toxicological analysis
10.2.4 Other investigations
10.3 Life supportive procedures and symptomatic/specific treatment
10.4 Decontamination
10.5 Elimination
10.6 Antidote treatment
10.6.1 Adults
10.6.2 Children
10.7 Management discussion
11.1 Case reports from literature
11.2 Internally extracted data on cases
11.3 Internal cases
12. Additional information
12.1 Availability of antidotes
12.2 Specific preventive measures
12.3 Other

1.1 Substance
Gentamicin sulfate
1.2 Group
Aminoglycoside antibiotic
1.3 Synonyms
1.4 Identification numbers
1.4.1 CAS number
1405-41-0 (Gentamicin sulfate)
1.4.2 Other numbers
89 – c (Martindale Online monograph link number)
LY2450000 (Gentamycin) RTECS
1.5 Brand names, Trade names
Gentamicin injection B.P.
Gentamicin sulphate injection (U.S.P.)
Cidomycin (Roussel)
Genticin (Nicholas), Lugacin (Lagap)
Sintepul, Gentamina, Glevomicina, Rovixida (Argentina)
Refobacin (Australia)
Geomycin (Belgium)
Gentalline, Gentalyn (France)
Refobacin, Sulmycin (Germany),
Gentacin (Japan)
Genalfa, Gentalyn, Gentibioptal, Genticol, Ribomicin (Italy)
Biogen, Espectrocina, Genta-Gobens, Gentadavur, Gentallenas,
Gentamival, Gento, Gentoma, Nuclogen, Sulgemicin (Spain)
Garamicina (Sweden)
Genoptic (USA)
Garamycin (Kirby – Warrick, UK)
Gentamicin BDH (E. Merck, UK)
Gentigan (E. Merck, UK)
Cidomycin Intrathecal Injectable (Roussel, UK)
Gentamicin L – BDH (E. Merck, UK) (Nicholas, UK)
Cidomycin Injectable Paediatric (Roussel, UK)
Garamycin Paediatric Injection (Kirby – Wanick, UK)
Genticin Paediatric Injection (Nicholas, UK)
Cidomycin Sterile Powder (Roussel, UK)
Genticine Pure Powder (Nicholas, UK)
Gentamicin Powder
Gentamicin eye drops (B.P.) (0.3% sulphate) (for both eye and
Minims Gentamicin (S & N Pharma UK),
Gentamicin sulphate ophthalmic solution (U.S.P.)
Alcomicin (Alcon, UK:Farillon, UK)
Gentamicin eye ointment (0.3% sulphate)
Gentamicin ear drops (0.3% sulphate)
Gentisone HC (Nicholas)
Gentamicin cream
Genticin HC (Nicholas)
Gentamicin skin ointment
Cidomycin Topical (Roussel)
Gentamicin Ointment,

(Merck, 1989)
1.6 Manufacturers, Importers
Chinoin (Hungary), Roussel (UK), Kirby-Warrick (UK), Nicholas
(UK), E. Merck (UK)

Local importers: Hoechst (Ceylon) Ltd., and Chemical
Industries (Colombo) Ltd.
2.1 Main risks and target organs
The main toxic effects are vestibular damage, deafness and
renal dysfunction. The damage on the vestibular portion of
the eighth cranial nerve appears to be greater than that on
the cochlear portion. The main target organs are the eighth
cranial nerves and the kidneys.
2.2 Summary of clinical effects
Damage to eighth cranial nerve (both divisions) resulting in
tinnitus, deafness, nausea, vomiting, vertigo, dizziness and
nystagmus, and nephrotoxicity causing acute tubular necrosis
resulting in renal failure.
2.3 Diagnosis
Loss of hearing, dizziness, vertigo, ataxia, nausea, vomiting
and renal impairment developing in a patient on gentamicin
therapy suggests a diagnosis of gentamicin toxicity. Other
toxic features are muscular paralysis and respiratory

Monitoring serum drug concentration is useful both to detect
toxicity and in the management of acute poisoning.
Assessment of renal function (blood urea, serum creatinine,
creatinine clearance) is helpful to detect renal damage.
Audiometry may help in the detection of ototoxicity.
2.4 First aid measures and management principles
No first-aid measures are available.
Stop further administration of gentamicin.

Gentamicin can be removed from the body by haemodialysis,
peritoneal dialysis and charcoal haemoperfusion.
Approximately 50% of the administered dose can be removed in
12 hours by haemodialysis.
3.1 Origin of the substance

Gentamicin was isolated in 1963 by Weinstein and colleagues
from the soil fungus Micromonospora purpura (of the
Actinomycete group) (Sande and Mandell, 1985). It was
introduced in the USA in 1969. It is a “complex” of
gentamicins C1, C1a and C2 and also gentamycin A which differs
from the other members of the complex but is similar to
Kanamycin C (Merck, 1989). Within the aminoglycoside family
the suffix “- mycin” is used in the name when the antibiotic
is produced by Streptomyces species and “micin”, when produced
by Micromonospora species.
3.2 Chemical structure
Gentamicin C
In Gentamicin C1, R1 = R2 = CH3
Gentamicin C2, R1 = CH3; R2 = H
Gentamicin C1a, R1 = R2 = H
Gentamicin C1, C21H43N5O7
Gentamicin C2, C20H41N5O7
Gentamicin C1a, C19H39N5O7

Gentamicin has two amino sugars joined in a glycosidic linkage
to a hexose nucleus. The hexose in this case is 2-
deoxystreptamine. Hence the compound is an aminoglycosidic

Gentamicin C1a, C19H39N2O7,
diamino-2,3,4,6-tetradeoxy- -D-erythro-hexopyranosyl-(1 4)-2-
Also known as gentamicin D.

Gentamicin A, C18H36N4O10, 0-2-amino-2-deoxy- -D-
glucopyranosyl-(1 4)-0-[3-deoxy-
3-(methylamino)- -D-xylopyranosyl-(1 6)]-2-deoxy-D-
3.3 Physical properties
3.3.1 Properties of the substance
Highly water soluble polar cation with an
optimum pH of 6 to 8. Antibiotic activity is
inhibited by acid pH and divalent cations.

Gentamicin sulphate is a white to buff coloured,
odourless, hygroscopic powder containing not
more than 15% water. Melts with decomposition
between 200 and 250°C. Moderately soluble in
ethanol, methanol and acetone. Practically
insoluble in benzene and halogenated
hydrocarbons. A solution in water is
dextrorotatory. A 4% solution in water has a pH
of 3.5 to 5.5.
3.3.2 Properties of the locally available formulation
3.4 Other characteristics
3.4.1 Shelf-life of the substance
Refer to the manufacturers instructions.
Stable in light, air and heat
3.4.2 Shelf-life of the locally available formulation

3.4.3 Storage conditions
Store in air-tight containers.
The potency of gentamicin sulphate is lost in plastic
disposable syringes and a brown precipitate is formed.
Storage in glass disposable syringes should not exceed
30 days.
3.4.4 Bioavailability
The bioavailability of gentamicin by intramuscular and
intravenous injection is assumed to be complete.
3.4.5 Specific properties and composition
No data available.
4.1 Indications
Gentamicin is widely used in the treatment of severe
infections. It is active against many strains of Gram-
negative bacteria and Streptococus aureus. It is
inactive against anaerobes and poorly active against
Streptococus haemolyticus and pneumococcus. When used as
presumptive treatment prior to microbiological
identification of the pathogen, gentamicin is combined
with penicillin and/or metronidazole.

Examples of use: septicaemia; neonatal sepsis; neonatal
meningitis; biliary tract infection; pyelonephritis,
prostatitis; endocarditis.
4.2 Therapeutic dosage
4.2.1 Adults

In patients with normal renal function: 2 to 5 mg/kg
daily in divided doses eight-hourly.

In patients with impaired renal function: serum
concentrations of gentamicin must be measured during
therapy, and dosage adjusted to give peak concentrations
below 10 mg/l and trough concentrations below 2 mg/l.
The dose interval should be increased to 12 hours when
creatinine clearance (CCr) is 30 to 70 ml/minute; to 24
hours when CCr is 10 to 30 ml/minute; to 48 hours when
CCr is 5 to 10 ml/minute; and three to four days after
dialysis at 5 ml/minute.

In a usual ten-day course, the peak and trough
gentamicin level in serum must be measured at least
once. If no facilities for measurement are available it
is better not to continue treatment beyond 72 hours.

Intrathecal administration: 1 to 5 mg/day in divided
doses every eight hours, with concurrent intramuscular
administration of 2 to 4 mg/kg/day.
4.2.2 Children
Serum concentrations should be monitored daily.

In children with normal renal function:
Up to 2 weeks of age: 3 mg/kg every 12 hours
2 weeks to 12 years: 2mg/kg every 8 hours.

Neonates with severe infections
3 mg/kg 12-hourly

Impaired renal function: Serum concentrations of
gentamicin must be measured during therapy, and dosage
adjusted to give peak concentrations below 10 mg/l and
trough concentrations below 2 mg/l.
4.3 Contraindications
Pregnancy; myasthenia gravis

Cautions in use: pre-existing renal disease; advanced age;
simultaneous use of loop diuretics, cephalosporins, or
neuromuscular blocking agents
5.1 Oral
Gentamicin is not absorbed after oral administration.
5.2 Inhalation
No data available.
5.3 Dermal
In ointment form, gentamicin is absorbed slowly. However
topical creams may give rise to rapid absorption sufficient to
cause adverse effects and substantial systemic absorption may
occur after application to large areas of denuded skin (as in
5.4 Eye
The pyrogen-free solution (intrathecal preparation) may also
be used for ocular administration (subconjunctival or
5.5 Parenteral
Ototoxicity and nephrotoxicity are seen principally with this
form of administration.
5.6 Other
Administration by intrathecal and intraperitoneal routes can
also result in poisoning.
6.1 Absorption by route of exposure
Parenteral (IM, SC and IV):
Less than 1% of a dose of gentamicin is absorbed following
oral or rectal administration; it is well absorbed after
subcutaneous and intramuscular injection. Serum levels after
intramuscular administration are slightly lower than after
intravenous administration.

After intramuscular administration, peak plasma concentrations
are reached in 30 to 90 minutes. A dose of 1 mg/kg produces
average peak plasma concentration of about 4 µg/ml although
there may be considerable inter-individual variation and
higher concentrations in patients with renal impairment
(Reynolds, 1982). In critically ill patients (especially with
shock) the absorption from intramuscular sites is poor.
6.2 Distribution by route of exposure
Gentamicin is not significantly protein bound. Peak serum
levels are observed 30 to 60 minutes after an intramuscular
dose and immediately after an intravenous dose.

The volume of distribution is approximately 25% of lean body

weight and equal to extracellular fluid volume.

The concentration of gentamicin in secretions and most tissues
is low. High concentrations are found in renal cortical
tissue, endolymph and perilymph of the inner ear. The
concentration in bile approaches close to 30% of serum
concentration. Penetration into respiratory secretions is
poor. Diffusion into pleural and synovial fluid is slow but
considerable concentrations may be achieved with repeated

The concentration in cerebrospinal fluid (CSF) is less than
10% of the concomitant plasma concentration, though in the
presence of meningeal inflammation the concentration may reach
20% of plasma levels. In neonates with meningitis, therapeutic
concentrations may be reached in CSF with parenteral dosing
alone; in adults, intrathecal administration is necessary.

Similarly, penetration to ocular tissue is minimal and
periocular injection of gentamicin is necessary in the
treatment of bacterial endophthalmitis.
6.3 Biological half-life by route of exposure
The elimination half-life in patients with normal renal
function is 2 hours. When creatinine clearance is halved the
t1/2 in serum doubles; adjustments are therefore necessary in
patients with renal failure (Cutler, 1972). The elimination
half-life is 20 to 40 times greater in anephric patients than
in patients with normal renal function and it is also
prolonged in neonates.
6.4 Metabolism
Gentamicin is not metabolized. It is excreted by glomerular
filtration in an active, unchanged form.
6.5 Elimination by route of exposure
Extremely high urinary concentrations may be achieved: after
an intravenous dose of 1 mg/kg, the urinary concentration may
reach 100 mg/l and, after 2 mg/kg, 300 mg/l. Gentamicin
accumulates in renal cortical tissue, from which it is
released over two to three weeks after cessation of therapy.
Extremely sensitive assay techniques can detect very small
quantities of gentamicin in renal cortical tissue for several
months after administration (Schentag et al, 1977).

High concentrations of gentamicin are reached in the liver,
but it is not excreted in bile. Drug levels are low in
biliary obstruction.
7.1 Mode of action
7.1.1 Toxicodynamics
As gentamicin accumulates in the renal cortex, a
critical concentration is reached when the concentrating
ability of the kidney becomes impaired.
Nephrotoxicity appears to be related to the duration for
which the trough serum concentration exceeds 2 µg/ml (De
Broe et al, 1986). The exact mechanism of toxicity is

Ototoxicity and vestibular toxicity seem most highly
correlated with elevated peak concentrations (greater
than 10 µg/ml) of gentamicin. Gentamicin accumulates in
endolymph and perilymph (Huy et al, 1983) and
progressive destruction of ventricular and cochlear
cells occurs. Repeated courses of gentamicin may
produce progressive destruction of cells leading to
deafness. Gentamicin appears to damage the vestibular
portion more than the cochlear portion.

Neuromuscular blockade with acute muscular paralysis and
apnoea may occur rarely. Most episodes have occurred in
association with anaesthesia or administration of other
neuromuscular blockers but may also occur after
intrapleural or intraperitoneal instillation of large
doses of gentamicin or other aminoglycosides. This
phenomenon may occur after intravenous or intramuscular
administration (Pittinger et al, 1970).
7.1.2 Pharmacodynamics
The bactericidal effect of gentamicin is due to
inhibition of protein synthesis in susceptible bacteria.
7.2 Toxicity
7.2.1 Human data Adults
Plasma concentrations above 10 mcg/ml are
associated with high risk of toxicity, although
individual tolerance may vary. Acute renal
failure has developed in patients after
treatment with 1.7 to 2.9 g given over 12 to 18
days (Reynolds, 1982). Children
No data available.
7.2.2 Relevant animal data
No data available.
7.2.3 Relevant in vitro data
No data available.
7.3 Carcinogenicity
No data available.
7.4 Teratogenicity
Foetal auditory and vestibular nerve damage may occur. The
foetus is at greatest risk during the second and third
7.5 Mutagenicity
No data available.
7.6 Interactions
Concomitant use of ethacrynic acid, frusemide, piretanide and
vancomycin have been reported to increase the ototoxicity of
gentamicin. The nephrotoxicity is increased by concurrent use
of some cephalosporins, cisplatin and indomethacin (Reynolds,

The aminoglycoside molecule (gentamicin) is inactivated
chemically by physical contact with penicillins and to a
lesser extent by cephalosporins (Farchione, 1981). Although
this is a laboratory phenomenon with no great clinical
importance, these antibiotics must not be mixed together.

7.7 Main adverse effects
Ototoxicity, which may be irreversible.

Both vestibular and auditory dysfunction may occur.

Nephrotoxicity may precipitate impairment of renal function or
overt renal failure.

A syndrome of acute neuromuscular paralysis may occur after
administration of gentamicin and general anaesthesia or
neuromuscular blocking agents. Although rare, neuromuscular
blockade may also occur after intracavitory installation or
parenteral injections of gentamicin.

Intrathecal/intraventricular administration may cause local
inflammation leading to radiculitis and other complications.
Fever and pleocytocis of CSF may occur after intrathecal
administration of gentamicin.
8.1 Material sampling plan
8.1.1 Sampling and specimen collection Toxicological analyses Biomedical analyses Arterial blood gas analysis Haematological analyses Other (unspecified) analyses
8.1.2 Storage of laboratory samples and specimens Toxicological analyses Biomedical analyses Arterial blood gas analysis Haematological analyses Other (unspecified) analyses
8.1.3 Transport of laboratory samples and specimens Toxicological analyses Biomedical analyses Arterial blood gas analysis Haematological analyses Other (unspecified) analyses
8.2 Toxicological Analyses and Their Interpretation
8.2.1 Tests on toxic ingredient(s) of material Simple Qualitative Test(s) Advanced Qualitative Confirmation Test(s) Simple Quantitative Method(s) Advanced Quantitative Method(s)
8.2.2 Tests for biological specimens Simple Qualitative Test(s) Advanced Qualitative Confirmation Test(s) Simple Quantitative Method(s) Advanced Quantitative Method(s) Other Dedicated Method(s)
8.2.3 Interpretation of toxicological analyses
8.3 Biomedical investigations and their interpretation
8.3.1 Biochemical analysis Blood, plasma or serum Urine Other fluids

8.3.2 Arterial blood gas analyses
8.3.3 Haematological analyses
8.3.4 Interpretation of biomedical investigations
8.4 Other biomedical (diagnostic) investigations and their
8.5 Overall Interpretation of all toxicological analyses and
toxicological investigations
8.6 References
9.1 Acute poisoning
9.1.1 Ingestion
Not relevant.
9.1.2 Inhalation
Not relevant.
9.1.3 Skin exposure
No data available.
9.1.4 Eye contact
No data available.
9.1.5 Parenteral exposure
Poisoning from this route causes acute renal failure due
to acute tubular necrosis. It also causes tinnitus,
impairment of hearing, nausea, vomiting, dizziness,
vertigo and nystagmus.
9.1.6 Other
No data available.
9.2 Chronic poisoning
9.2.1 Ingestion
Not relevant.
9.2.2 Inhalation
Not relevant.
9.2.3 Skin exposure
Systemic absorption may occur following application to
large denuded areas of the body (especially in cream
form). Serum concentrations of 1 µg/ml may be achieved
and 2 to 5% of the applied dose is excreted in the
urine. Cases of total deafness following inadvertent
systemic absorption of neomycin, another aminoglycoside,
have been reported (Greenberg & Momary, 1965).
9.2.4 Eye contact
No data available.
9.2.5 Parenteral exposure
Prolonged administration can cause both renal and 8th
cranial nerve toxicity.
9.2.6 Other
No data available.
9.3 Course, prognosis, cause of death
Acute renal failure caused by gentamicin is reversible.
Deafness may be permanent.
9.4 Systematic description of clinical effects
9.4.1 Cardiovascular
No significant effects.
9.4.2 Respiratory
Respiratory depression may occur as a result of
neuromuscular blockade.
9.4.3 Neurological CNS

There may be convulsions, encephalopathy,
lethargy, confusion, hallucinations, mental
depression and delirium (Pleocytocis can be
observed in the cerebrospinal fluid). Peripheral nervous system
Cochlear damage: Hearing loss, initially, high

Vestibular damage: This is more common than
hearing loss. Onset of labyrinthine dysfunction
may be preceded by a headache for one to two
days. This is followed by an acute phase during
which nausea, vomiting and difficulty in
balancing develops and persists for one to two
weeks; other features include dizziness, vertigo,
tinnitus and roaring in the ears. Vertigo is
worse on standing. Inability to perceive
termination of movement (“mental past pointing”)
may occur. Drifting of eyes occurs at the
cessation of movement so that focusing and
reading are difficult; positive Romberg’s sign
and nystagmus are also observed.

Chronic labyrinthitis: Ataxia, inability to walk
or make sudden movements. Sensory symptoms such
as numbness, tingling may occur. Autonomic nervous system
No information available. Skeletal and smooth muscle
Muscle twitching may occur. Neuromuscular
paralysis is a rare but important toxic effect.
9.4.4 Gastrointestinal
Nausea and vomiting.
9.4.5 Hepatic
Increased serum aminotransferase and serum bilirubin
levels have been reported rarely (Reynolds, 1982).
9.4.6 Urinary Renal
A reversible, mild renal impairment may occur in
8 to 26% of patients who are given gentamicin.

Proximal tubular cells (PCT) may be damaged due
to retention of gentamicin. This may be manifest
as increased excretion of renal PCT tubular
brush border enzymes such as alanine
aminopeptidase, alkaline phosphatase and beta-D-
glucosaminidase. This stage is followed by the
appearance of granular and hyaline casts in
urine, proteinuria and a defect in renal
concentrating ability. The glomerular
filtration rate is reduced. Acute tubular
necrosis and interstitial nephritis may occur. Other
No data available.
9.4.7 Endocrine and reproductive systems
No data available.

9.4.8 Dermatological
Skin rashes have been reported. Loss of hair and
eyebrows has been reported in a patient during treatment
with gentamicin (Yoshioka & Matsuda, 1970).
9.4.9 Eye, ear, nose, throat: local effects
Subconjunctival infection: pain, hyperaemia and
conjunctival oedema (Reynolds, 1993).
9.4.10 Haematological
Rarely, anaemia and purpura have been reported
(Reynolds, 1982).
9.4.11 Immunological
Hypersensitivity to gentamicin can occur rarely (see
9.4.12 Metabolic Acid-base disturbances
No information available. Fluid and electrolyte disturbances
Hypomagnesaemia may occur with prolonged
therapy. There may be hypocalcaemia and
hypokalaemia. Others
No data available.
9.4.13 Allergic reactions
Hypersensitivity reactions, and very rarely anaphylaxis
can occur. Collapse with tachycardia, hypotension and
apnoea has occurred within a minute of starting to
inject 80 mg of gentamicin IV (Hall, 1977).
9.4.14 Other clinical effects
No data available.
9.4.15 Special risks
Pregnancy: gentamicin is contraindicated due to the
risk of auditory and ves-tibular damage. Accumulation
in foetal plasma and amniotic fluid occurs after
administration during late pregnancy.

Breast feeding: no data available.

Enzyme deficiencies: no data available
9.5 Other
No data available.
9.6 Summary
10.1 General principles
Make a proper assessment of airway, breathing,circulation
and neurological status of the patient.

Discontinue gentamicin therapy when early signs of vestibulo-
cochlear toxicity, such as tinnitus and impairment of
hearing are observed. Careful observation of renal function
is necessary and when renal failure is established. Toxic
concentrations of gentamicin can be reduced by either
haemoperfusion or dialysis.
10.2 Relevant laboratory analyses
10.2.1 Sample collection
Collect blood for serum gentamicin levels (peak or
trough samples or random sample) and biomedical

10.2.2 Biomedical analysis
Biochemical profile with blood urea nitrogen,
creatinine, and electrolytes should be obtained.
Creatinine clearance should be determined.

Elevation of serum creatinine, blood urea nitrogen
and serum potassium is observed in renal failure.
10.2.3 Toxicological analysis
Monitor plasma gentamicin levels. Peak levels over
10 mcg/ml or trough levels over 2 µg/ml are
associated with adverse effects.
10.2.4 Other investigations
Not relevant
10.3 Life supportive procedures and symptomatic/specific
Make a proper assessment of airway, breathing,circulation
and neurological status of the patient.

If respiration is impaired maintain clear airway, aspirate
secretions, and administer oxygen. Support ventilation
using appropriate mechanical device. Assess renal function
regularly. Monitor and correct fluid and electrolyte
balance. Control convulsions by appropriate drugs such as
diazepam. Regular audiometric examinations may help in
follow up.
10.4 Decontamination
Not relevant.
10.5 Elimination
Forced diuresis – Chronic diuretic therapy and loop
diuretics increase the toxicity of gentamicin. Therefore,
the use of these are contraindicated.

Dialysis: peritoneal dialysis and haemodialysis are useful
for removing gentamicin from the body. Approximately 50% of
the administered dose is removed in 12 hours by
haemodialysis. Frequent monitoring of the plasma gentamicin
concentration is advised. Peritoneal dialysis is less
effective than haemodialysis. Clearance rates are
approximately 5 to 10 ml/minute, but is highly variable. In
one case, haemoperfusion through acrylic resin coated
charcoal combined with haemodialysis removed about 70% of
the gentamicin given in excess to an anuric patient.
10.6 Antidote treatment
10.6.1 Adults
No specific antidote is available.
10.6.2 Children
No specific antidote is available.
10.7 Management discussion
Gentamicin is not absorbed from the gastrointestinal tract
and no serious adverse effects would be seen with deliberate
oral ingestion. Toxic doses may be administered
inadvertently to normal subjects or toxicity may occur due
to undetected renal disease.

Dialysis or haemoperfusion is indicated in cases of massive

overdose or when serum concentrations of gentamicin are
extremely high.

Calcium salts given intravenously have been used to counter
the neuromuscular blockade caused by gentamicin. The
effectiveness of neostigmine has been variable (Reynolds,
11.1 Case reports from literature
Acute nephrotoxicity has resulted after inadvertent
administration of 25mg to 152mg intramuscularly or
intravenously (Fuquay et al, 1981; Bolam et al, 1982; Smith
1982; Koren et al, 1986).
11.2 Internally extracted data on cases
To be completed by the centre.
11.3 Internal cases
To be completed by the centre.
12. Additional information
12.1 Availability of antidotes
Not relevant.
12.2 Specific preventive measures
Gentamicin should not be used in patients with a known
history of allergy to aminoglycosides. Gentamicin should be
given with care and in reduced dosage to patients with
impaired renal function and plasma concentrations of
gentamicin should be checked frequently. Gentamicin should
be withdrawn immediately if symptoms of ototoxicity occur.
12.3 Other
No data available.
Bolam DL, Jenkins SA, Nelson RM (1982). Aminoglycoside overdose
in neonates. J Ped 100: 835.

Cutler RE et al (1972). Correlations of serum creatinine
concentrations and gentamicin half life. J Am Med Assoc 219:

DeBroe M et al (1986). Choice of drug and dosage regimen: Two
important risk factors for aminoglycoside nephrotoxicity. Am J
Med 80(6B): 115.

Farchione LA (1981). Inactivation of aminoglycosides by
penicillins. J Antimicr Chemotherapy 8: 27.

Fuquay D, Koup J, Smith AL (1981). Management of neonatal
gentamicin overdosage. J Pediatr 99: 473 – 476.

Greenberg JH, Momary H (1965). Audiotoxicity and nephrotoxicity
due to orally administered neomycin. J Am Med Assoc 194: 827.

Hall FJ (letter) (1977). Lancet 2: 455.

Huy PTB, Meulemans A, Wassef M, Manuel C, Sterkess O, Amiel C
(1983). Gentamicin persistence in rat endolymph and perilymph
after a two day constant infusion. Antimicrob Agents Chemother
23: 344 – 346.

Koren G, Barzilay Z, Greenwald M (1986). Tenfold errors in
administration of drug doses: a neglected iatrogenic disease in
paediatrics. Paediatrics 77: 848 – 849.

Pittinger CB, Eryasa Y, Adamson R (1970). Antibiotic-induced
paralysis. Anesth Analg 49: 482 – 501.

Reynolds JEF (1982). Martindale, The Extrapharmacopoeia. 28th
Edition. London. The Pharmaceutical Press. 1166 – 1173.

Reynolds JEF (1993). Martindale, The Extrapharmacopoeia. 30th
Edition. London. The Pharmaceutical Press. 170 – 172.

Sande MA, Mandell GL (1985). In: Rall TW, Murad F, eds.
Antimicrobial Agents. The Pharmacological Basis of Therapeutics.
7th Edition, New York. Macmillan Publishing Company. 1151-1169.

Schentag JJ et al (1977). Tissue persistence of gentamicin in
humans. J Am Med Assoc 238: 327.

Smith AL (1982). Aminoglycoside overdose in neonates (letter). J
Pediatr 100: 835

Yoshioka H, Matsuda I (letter) (1970). J Am Med Assoc 211: 123.
Authors: Dr Ravindra Fernando and Dr R.L. Jayakody
National Poisons Information Centre,
General Hospital
Sri Lanka

Date: August 1990

Update: Dr R. Fernando

Date: June 1993

Review: IPCS, May 1994

ARSENIC ACID (80% in water)ICSC: 1625
Date of Peer Review: October 2005

Arsenic acid hemihydrate

ortho-Arsenic acid solution

CAS #7778-39-4H3AsO4
RTECS #CG0700000Molecular mass: 141.94
UN #1553
EC Index #033-005-00-1


FIREIn case of fire in the surroundings: use appropriate extinguishing media.


InhalationCough. Shortness of breath. Further see Ingestion.Closed system and ventilation.Fresh air, rest. Refer for medical attention.
SkinRedness. Pain. Burning sensation.Protective gloves. Protective clothing.Remove contaminated clothes. Rinse skin with plenty of water or shower. Refer for medical attention.
EyesRedness. Pain.Safety spectacles .First rinse with plenty of water for several minutes (remove contact lenses if easily possible), then take to a doctor.
IngestionSore throat. Nausea. Vomiting. Diarrhoea. Convulsions.Wash hands before eating. Do not eat, drink, or smoke during work.Rinse mouth. Induce vomiting (ONLY IN CONSCIOUS PERSONS!). Rest. Refer for medical attention.


Evacuate danger area! Consult an expert! Personal protection: complete protective clothing including self-contained breathing apparatus. Do NOT let this chemical enter the environment. Collect leaking liquid in sealable plastic containers. Absorb remaining liquid in sand or inert absorbent and remove to safe place.Unbreakable packaging; put breakable packaging into closed unbreakable container. Do not transport with food and feedstuffs.

EU Classification

Symbol: T, N

R: 45-23/25-50/53

S: 53-45-60-61

Note: [A, E]

UN Classification

UN Hazard Class: 6.1

UN Pack Group: I

Transport Emergency Card: TEC (R)-61S1553 or 61GT4-IStore in an area without drain or sewer access. Separated from strong oxidants, strong bases, metals strong reducing agents, food and feedstuffs. Do not store or transport in aluminium, copper, iron or zinc.


Programme on

Chemical Safety

Prepared in the context of cooperation between the International Programme on Chemical Safety and the Commission of the European Communities © IPCS, CEC 2005



ARSENIC ACID (80% in water)ICSC: 1625





The substance decomposes on heating producing toxic and corrosive fumes. The substance is a strong oxidant and reacts with combustible and reducing materials. The substance is a medium strong acid. Attacks metals to produce toxic and flammable arsine (see ICSC 0222).


TLV: (as As) 0.01 mg/m³ as TWA; A1 (confirmed human carcinogen); BEI issued; (ACGIH 2005).

MAK: Carcinogen category: 1; Germ cell mutagen group: 3A; (DFG 2005).


The substance can be absorbed into the body by inhalation of its vapour, through the skin and by ingestion.


A harmful contamination of the air can be reached very quickly on evaporation of this substance at 20°C on spraying.


The substance is irritating to the eyes, the skin and the respiratory tract. The substance may cause effects on the blood, cardiovascular system, gastrointestinal tract, liver and peripheral nervous system. The effects may be delayed. See Notes.


The substance may have effects on the peripheral nervous system and skin, resulting in polyneuropathy and skin lesions. The substance may have effects on the cardiovascular system. This substance is carcinogenic to humans.

Boiling point: 120°C

Solubility in water: at 20°C very good

The substance is very toxic to aquatic organisms.
Depending on the degree of exposure, periodic medical examination is suggested. The symptoms of acute poisoning do not become manifest until hours. Do NOT take working clothes home.
LEGAL NOTICENeither the CEC nor the IPCS nor any person acting on behalf of the CEC or the IPCS is responsible for the use which might be made of this information
© IPCS, CEC 2005

INTOX Home Page



SM Bradberry BSc MB MRCP
WN Harrison PhD CChem MRSC
ST Beer BSc

National Poisons Information Service
(Birmingham Centre),
West Midlands Poisons Unit,
City Hospital NHS Trust,
Dudley Road,
B18 7QH

This monograph has been produced by staff of a National Poisons
Information Service Centre in the United Kingdom. The work was
commissioned and funded by the UK Departments of Health, and was
designed as a source of detailed information for use by poisons
information centres.

Peer review group: Directors of the UK National Poisons Information


Toxbase entry

Type of product

Used in wood preservatives and insecticides. In water arsenic acid
forms the arsenate ion.


Small ingestions of dilute (<3%) arsenate solutions usually are
without serious adverse effects. A patient has survived the deliberate
ingestion of 10 g arsenate (Mathieu et al, 1992).


Systemic toxicity may follow arsenic acid ingestion, inhalation or
topical exposure.


– Causes skin burns. Systemic arsenic poisoning may occur
after substantial exposure.


Minor ingestions (small amounts of dilute (<3%) solutions):
– Usually no serious effects. Mild gastrointestinal upset may

Substantial ingestions:
– Rapid onset (within 1-2 hours) of burning of the mouth and
throat, hypersalivation, dysphagia, nausea, vomiting,
abdominal pain and diarrhoea.
– In severe cases gastrointestinal haemorrhage, cardiovascular
collapse, renal failure, seizures, encephalopathy and
rhabdomyolysis may occur.
– Other features: facial and peripheral oedema, ventricular
arrhythmias (notably torsade de pointes), ECG abnormalities
(QT interval prolongation, T-wave changes), muscle cramps.
– Investigations may show anaemia, leucopenia,
thrombocytopenia or evidence of intravascular haemolysis.
– Death may occur from cardiorespiratory or hepatorenal
failure. The adult respiratory distress syndrome (ARDS) has
been reported.
– Survivors of severe poisoning may develop a peripheral
neuropathy and skin lesions typical of chronic arsenical


– Rhinitis, pharyngitis, laryngitis and tracheobronchitis may
occur. Tracheal and bronchial haemorrhage may complicate
severe cases.

Chronic arsenic exposure

– may occur following ingestion, inhalation or topical
exposure. Features include weakness, lethargy,
gastrointestinal upset, dermal manifestations
(hyperkeratosis and “raindrop” pigmentation of the skin), a
peripheral (motor and sensory) neuropathy and psychological

– Also reported: peripheral vascular disease (cold sensitivity
progressing to ulceration and gangrene), renal tubular or
cortical damage and haematological abnormalities (notably



1. Irrigate with copious volumes of water.
2. Treat burns conventionally.
3. Consider the possibility of systemic arsenic poisoning after
significant exposure.


Minor ingestions:
1. Gastrointestinal decontamination is unnecessary.
2. Symptomatic and supportive care only.

Substantial ingestions:
1. Most patients will vomit spontaneously but in those who do not,
gastric lavage should be considered only if the patient presents
within one hour.
2. Supportive measures are paramount. Intensive resuscitation may be
required. Ensure adequate fluid replacement and close observation
of vital signs including cardiac monitoring.
3. Diarrhoea can be controlled with loperamide.
4. Monitor blood urea, creatinine, electrolytes, liver function and
full blood count.
5. Collect blood and urine for arsenic concentration measurements.
6. ECG evidence of QT prolongation may precede atypical ventricular
arrhythmias (notably torsade de pointes). Avoid drugs which
prolong the QT interval e.g. procainamide, quinidine or
disopyramide. Isoprenaline is effective with phenytoin,
lignocaine or propranolol as alternatives.

7. Antidotes – chelation therapy with either dimercaprol, DMSA or
DMPS should be considered in symptomatic patients where there is
analytical confirmation of the diagnosis, but only after
specialist advice from the NPIS.


Armstrong CW, Stroube RB, Rubio T, Siudyla EA, Miller GB.
Outbreak of fatal arsenic poisoning caused by contaminated drinking
Arch Environ Health 1984; 39: 276-9.

Donofrio PD, Wilbourn AJ, Albers JW, Rogers L, Salanga V, Greenberg
Acute arsenic intoxication presenting as Guillain-Barré-like syndrome.
Muscle Nerve 1987; 10: 114-20.

Engel RR, Hopenhayn-Rich C, Receveur O, Smith AH.
Vascular effects of chronic arsenic exposure: a review.
Epidemiol Rev 1994; 16: 184-209.

Goldsmith S, From AHL.
Arsenic-induced atypical ventricular tachycardia.
N Engl J Med 1980; 303:1096-7.

Greenberg C, Davies S, McGowan T, Schorer A, Drage C.
Acute respiratory failure following severe arsenic poisoning.
Chest 1979; 76: 596-8.

Kersjes MP, Maurer JR, Trestrail JH, McCoy DJ.
An analysis of arsenic exposures referred to the Blodgett Regional
Poison Center.
Vet Hum Toxicol 1987; 29: 75-8.

Kingston RL, Hall S, Sioris L.
Clinical observations and medical outcome in 149 cases of arsenate ant
killer ingestion.
Clin Toxicol 1993; 31: 581-91.

Kosnett MJ, Becker CE.
Dimercaptosuccinic acid as a treatment for arsenic poisoning.
Vet Hum Toxicol 1987; 29: 462.

Mathieu D, Mathieu-Nolf M, Germain-Alonso M, Neviere R, Furon D,
Wattel F.
Massive arsenic poisoning – effect of hemodialysis and dimercaprol on
arsenic kinetics.
Intensive Care Med 1992; 18: 47-50.

McWilliams ME.
Accidental acute poisoning by a concentrated solution of arsenic acid
from percutaneous absorption: a case report.
Vet Hum Toxicol 1989; 31: 354.

Peterson RG, Rumack BH.
D-penicillamine therapy of acute arsenic poisoning.
J Pediatr 1977; 91: 661-6.

Substance name

Arsenic acid

Origin of substance

Reaction of arsenic trioxide with nitric acid
(HSDB, 1995)


Orthoarsenic acid (DOSE, 1992)

Chemical group

A compound of arsenic, a group VA element

Reference numbers

CAS 7778-39-4 (DOSE, 1992)
RTECS CG0700000 (RTECS, 1995)
UN 1554 (solid) 1553 (liquid) (DOSE, 1992)

Physicochemical properties

Chemical structure
H3AsO4 (DOSE, 1992)

Molecular weight
141.94 (DOSE, 1992)

Physical state at room temperature
Solid (HSDB, 1995)

White (HSDB, 1995)



Weakly acidic (HSDB, 1995)

3020 g/L in water at 12.5°C (HSDB, 1995)

Autoignition temperature

Chemical interactions
Aqueous solutions emit arsine gas when in contact with active
metals such as arsenic, iron, aluminium and zinc.
(HSDB, 1995)

Major products of combustion
Produces arsine gas (HSDB, 1995)

Explosive limits

Not flammable (HSDB, 1995)

Boiling point

2.2 at 20°C (HSDB, 1995)

Vapour pressure

Relative vapour density

Flash Point

Arsenic acid is a strong oxidizer. (HSDB, 1995)


Preparation of arsenate salts
Manufacture of pesticides (DOSE, 1992)

Hazard/risk classification

Index no. 033-005-00-1
Risk phrases
Carc. Cat. 1; R45 – May cause cancer
T: R23/25 – Also toxic by inhalation and if swallowed
Safety phrases
S53-45 – Avoid exposure-obtain special instructions before use.
In case of accident or if you feel unwell seek medical advice
immediately (show label where possible).
EEC No: NIF (CHIP2, 1994)


Arsenic acid is one of the most commercially important pentavalent
compounds of arsenic. It is formed from the treatment of arsenic
trioxide with nitric acid and used in the manufacture of arsenate
salts and pesticides. It is also formed by the slow reaction of
arsenic pentoxide in water (Fielder et al, 1986).

Aqueous solutions of arsenic acid can liberate arsine gas, the most
acutely toxic form of arsenic, when in contact with active metals
(HSDB, 1995).

It has been suggested that soluble arsenic compounds such as arsenic
acid represent a much more acute toxic hazard than insoluble, poorly
absorbed forms (Done and Peart, 1971).


The main source of arsenic exposure in the world population is
drinking water with an high inorganic arsenic concentration (Chiou et
al, 1995; Das et al, 1995). In water arsenic acid forms the arsenate
ion. Arsenate is the most thermodynamically stable oxide of arsenic in
water and as such is believed to be the predominant species,
especially under aerobic conditions (IPCS, 1981).

Arsenic acid and its salts have been used widely in industry (Fielder
et al, 1986; McWilliams 1989) and in pesticide formulations (Done and
Peart, 1971; Miller et al, 1980) where they have been the source of
acute and chronic, accidental and deliberate intoxication.


Once absorbed pentavalent arsenic is reduced in vivo to trivalent
arsenic. The principle mechanism of arsenic intoxication is disruption
of thiol proteins. For example, trivalent arsenic inactivates pyruvate
dehydrogenase by complexing with the sulphydryl groups of a lipoic
acid moiety (6,8-dithiooctanoic acid) of the enzyme (Jones, 1995).

Enhanced cellular destruction of damaged thiol proteins may produce
toxic oxygen radicals. Arsenic-induced reduced lymphocyte
proliferation (Gonsebatt et al, 1994) and impaired macrophage function
also have been described (Lantz et al, 1994).

Dong and Luo (1994) have suggested that while arsenic can directly
damage DNA, a more important mechanism in arsenic-induced
carcinogenicity is enhanced mutagenicity of other compounds via
increased DNA-protein crosslinks.

The affinity of arsenic for sulphydryl groups is utilized in chelation



Soluble arsenical compounds such as arsenic acid are well absorbed
after ingestion, when absorption is almost complete. Following
inhalation there is significant mucociliary clearance and
gastrointestinal absorption of respired particles (Fielder et al,

From very limited animal data arsenic acid appears to be well absorbed
through the lungs (Fielder et al, 1986).

Although direct evidence of transcutaneous arsenic absorption in man
is scarce (Fielder et al, 1986) there are reports of systemic arsenic
toxicity following presumed dermal exposure (Garb and Hine, 1977;
McWilliams, 1989).


Absorbed arsenic is distributed to all body tissues (Fielder et al,
1986). Once absorbed pentavalent arsenic is reduced in vivo to
trivalent arsenic. Trivalent arsenic is methylated in the liver to
methylarsonic acid and dimethylarsinic acid (IPCS, 1996). Short-term
studies on humans indicate that daily intake in excess of 0.5 mg
progressively, but not completely, saturates the capacity to methylate
inorganic arsenic (IPCS, 1996).


The half-life of arsenic in blood is about 60 hours with renal
excretion predominantly as mono- and dimethyl-derivatives (Waldron and
Scott, 1994). The whole body half-life of arsenic in six human
volunteers fitted a three compartment system, with 65.9 per cent of
orally administered arsenic acid having a half-life of 2.1 days, 30.4
per cent a half-life of 9.5 days and 3.7 per cent a half-life of 38.4
days (mean values) (Pomroy et al, 1980).


Dermal exposure

Severe foot burns occurred in a patient exposed to concentrated
arsenic acid-saturated clothing for eight hours (McWilliams, 1989).
Soft tissue deposits believed to be metallic arsenic were noted on
X-ray. The patient was transiently encephalopathic and developed a
chronic painful foot motor neuropathy. Twenty four hour urine arsenic
elimination ranged from 2500 mg falling to 160 mg during an eight week
course of penicillamine.

Garb and Hine (1977) reported a worker splashed down the left side of
his body with arsenic acid in an industrial accident. He immediately
removed a contaminated glove and washed his hands but did not notice
arsenic acid in his left shoe for some ten minutes. After washing the
affected leg and removing a soiled sock he continued to work for a
further four hours by which time he had sustained second degree burns
in the acid-exposed areas. Eleven hours after the incident he
developed pain and swelling at the contact sites plus nausea,
vomiting, diarrhoea and abdominal pain necessitating hospital
admission. During the next three days his vision became “foggy” and he
complained of a sore tongue and “aching teeth”. After seven days he
developed a burning sensation in the unexposed foot accompanied by
paraesthesiae in all extremities. He became progressively weaker and
was confined to a wheelchair. Two years after the accident diminished
touch sensation and tendon reflexes persisted and he could walk only
with a left leg brace and crutches. Chelation therapy was not employed
at any time.

Ocular exposure

Arsenic acid is an eye irritant and may cause burns. Most injuries
result from exposure to dusts, causing conjunctivitis, lacrimation,
photophobia and chemosis (Grant and Schuman, 1993).


The oral toxicity of arsenic acid is dependent on the amount ingested
and the concentration of the preparation. Tallis (1989) noted that
arsenate salts can react with hydrochloric acid in the stomach to
liberate arsenic acid and the corresponding metal chloride. In this
way poorly soluble salts such as lead arsenate can be a source of
arsenic acid which is soluble and well absorbed.

Although pentavalent arsenic is reduced in vivo to the generally
more toxic trivalent arsenic (Waldron and Scott, 1994) ingestion of
dilute arsenic acid salt solutions (less than three per cent) usually
are without serious adverse effects (Kingston et al, 1993). In 149
such cases involving sodium arsenate (2.28 per cent) ant killer, 97
per cent of patients were asymptomatic and only one required hospital
admission (Kingston et al , 1993).

Gastrointestinal toxicity

Of 57 cases of ant killer ingestion involving arsenic acid salts
(maximum arsenate concentration three per cent) only seven patients
were symptomatic. All these vomited, with abdominal pain, diarrhoea
and nausea also reported (Kersjes et al 1987).

Armstrong et al (1984) reported eight family members (two of whom
died) poisoned by well-water which contained arsenic 108 mg/L (form
unknown). All had gastrointestinal symptoms including dry mouth,
vomiting, dysphagia and/or diarrhoea. Urine arsenic concentrations
were directly related to the amount of water consumed. The estimated

daily dose of arsenic ingested by the surviving family members ranged
from 26-127 mg. An 11 year-old girl and a 27 year-old man died after
ingesting approximately 77 mg and 166 mg arsenic daily respectively
(duration unknown). Autopsy of the most severely poisoned patient
showed diffuse gastrointestinal tract inflammation.

A 32 year-old man ingested 900 mg of an arsenic acid salt, vomited
within one hour and developed diarrhoea three hours later. His
clinical course was complicated by hypotension and renal failure but
after 82 days chelation therapy with N-acetylcysteine he fully
recovered (Martin et al, 1990).

Another patient survived the deliberate ingestion of 10 g of an
arsenic acid salt (Mathieu et al, 1992). Severe nausea, vomiting and
abdominal tenderness developed within three hours with cardiovascular
collapse and subsequent acute renal failure requiring haemodialysis.
The patient made a full recovery over three months.

Other gastrointestinal features of arsenic poisoning include burning
of the mouth and throat with dysphagia (Heyman et al, 1956) and
hypersalivation (Schoolmeester and White, 1980).


Armstrong et al (1984) reported a 27 year-old man poisoned by well
water containing 108 mg/L arsenic. He presented after feeling unwell
for six days and on examination was jaundiced with a bilirubin
concentration of 120 µmol/L. He collapsed and died later that day and
autopsy showed a liver arsenic concentration of 86 mg/kg. He had
ingested an estimated 166 mg arsenic daily (duration unknown) and had
a urine arsenic concentration of 1.6 mg/L at post mortem. Seven other
members of his family with lower arsenic exposure showed transiently
elevated serum hepatic transaminase activities and total bilirubin
concentrations (values not given).

Schoolmeester and White (1980) reported a 16 year-old female who
ingested 300 mg sodium arsenate in a suicide attempt. She developed
severe abdominal pain and vomiting within 30 minutes. A 24 hour urine
collection had an arsenic concentration of 14.2 mg/L (time of
collection not stated). Forty-eight hours later serum liver
transaminase and alkaline phosphatase activities were elevated (values
not given) but these abnormalities resolved within six months.


Hypotension (Martin et al, 1990; Mathieu et al, 1992) or
rhabdomyolysis following substantial arsenic acid ingestion may
precipitate renal failure. Renal cortical necrosis has been described
(Gerhardt et al, 1978). Haematuria was reported in one patient in a
series of 57 cases of sodium arsenate ingestion (Kersjes et al, 1987).

Pyuria, proteinuria and elevated serum creatinine concentrations were
reported in members of a family poisoned with arsenic-contaminated
well-water (Armstrong et al, 1984). The most severely poisoned
patient, who died after a six day illness, developed gross haematuria,
a serum creatinine concentration of 390 µmol/L and heavy proteinuria.
He had consumed an estimated 166 mg arsenic daily (duration unknown)
and had a urine arsenic concentration at post mortem of 1.6 mg/L. No
chelation therapy was given. The other family members were treated
with dimercaprol and penicillamine. Six eventually recovered.

Cardiovascular toxicity

Tachycardia is reported frequently following ingestion of arsenic acid
salts and is contributed to by anxiety, intravascular fluid depletion
and possibly direct arsenic-induced cardiotoxicity (Le Quesne and
McCleod, 1977; Martin et al, 1990; Cullen et al, 1995).

Ventricular arrhythmias, notably torsade de pointes (Beckman et al,
1991) have been observed. Other ECG abnormalities include prolongation
of the QT interval (Goldsmith and From, 1980; Schoolmeester and White,
1980), idioventricular rhythm (Armstrong et al, 1984) and non-specific
T wave changes. Sudden onset bradycardia then asystole has been
reported following massive acute arsenic ingestion despite vigorous
resuscitation and no earlier arrhythmia.

Armstrong et al (1984) reported a 27 year-old man who had consumed an
unstated quantity of well-water containing 108 mg/L arsenic. After a
six day illness (diagnosed as an upper respiratory tract infection) he
collapsed and sustained a respiratory arrest and seizures. An
idioventricular rhythm was noted. He was resuscitated but remained
comatose and died a few hours later.

Pericardial effusion with tamponade was reported in another member of
the same family who had also drunk arsenic-contaminated water; the
outcome in this case was not stated (Armstrong et al, 1984).


In 57 sodium arsenate ingestions involving solutions containing
1.5-3.0 per cent arsenate, headache, dizziness, lethargy and
somnolence were each reported in 2 per cent of cases; 88 per cent of
patients were asymptomatic (Kersjes et al 1987).

More substantial arsenic ingestions have caused muscle cramps, a
sensorineural hearing deficit (Goldsmith and From, 1980),
encephalopathy (Jenkins, 1966) and seizures.

A peripheral sensory and/or motor neuropathy has been described in
survivors of severe acute arsenic poisoning although this is more
typical following chronic exposure.

Armstrong et al (1984) reported eight family members poisoned with
well-water containing 108 mg/L arsenic. All developed gastrointestinal
symptoms with “altered mental status” and seizures noted in four. Coma
developed in three patients and a peripheral neuropathy in two.

Goebel et al (1990) demonstrated acute wallerian degeneration of
myelinated nerve fibres in a patient who developed a symmetrical
polyneuropathy after attempting suicide by ingesting arsenic. Clinical
improvement was associated with microscopic evidence of neurological

A 46 year-old man developed feet numbness ten days after drinking a
solution of sodium arsenate (concentration unknown) in attempted
suicide. Two months after ingestion neurological examination
demonstrated distal muscle weakness bilaterally, absent knee and ankle
reflexes and reduced position and vibration sense with a high-stepping
gait. Sixteen months later there was improvement in both sensory and
motor deficits although residual disability was evident at eight year
follow-up (Le Quesne and McCleod, 1977).

Dermal toxicity

Le Quesne and McCleod (1977) described a patient who developed a
papular erythematous rash and generalized epidermal desquamation one
week after drinking 10 mL of an arsenate solution (concentration

Striate leukonychia (Mees’ lines) and hyperkeratotic or hyperpigmented
skin lesions are characteristic of chronic arsenic intoxication but
have been described also following substantial acute ingestion (Heyman
et al, 1956; Kyle and Pease, 1965; Jenkins, 1966).

Facial and peripheral oedema have also been reported following arsenic
ingestion (Heyman et al, 1956; Kyle and Pease, 1965).


In moderate or severe arsenic poisoning investigations typically show
anaemia, leucopenia or pancytopenia (Kyle and Pease, 1965; Armstrong
et al, 1984). There may be evidence of intravascular haemolysis and
the blood film often shows basophilic stippling (Kyle and Pease,

Mathieu et al (1992) described a 30-year-old male who ingested 10 g
sodium arsenate with suicidal intent. He developed severe
gastrointestinal features of arsenic poisoning within hours and
required haemodialysis for management of acute renal failure. Five
days after ingestion he developed thrombocytopenia and anaemia. Bone
marrow examination showed maturation arrest but recovery ensued over
10 days.

Multi-organ toxicity

Severe acute arsenic poisoning may result in death from
cardiorespiratory or hepatorenal failure (Jenkins, 1966; Armstrong et
al, 1984; Campbell and Alvarez, 1989; Morton and Dunnette, 1994). The
adult respiratory distress syndrome (ARDS) has been described
(Bolliger et al, 1992).


Inhalation of arsenic compounds causes rhinitis, pharyngitis,
laryngitis and tracheobronchitis (Morton and Dunnette, 1994).


Dermal exposure

Occupational exposure may lead to chronic arsenical toxicity.

Contact dermatitis has been reported in workers exposed to arsenic
acid salts used in crystal manufacture (Barbaud et al, 1995).


Ingestion of arsenic-contaminated drinking water (Feinglass, 1973;
Chiou et al, 1995), illicit whisky (Moonshine) (Gerhardt et al, 1980)
“tonics” or traditional remedies have caused chronic arsenical


Occupational exposure may lead to chronic arsenical poisoning.
Perforation of the nasal septum has been reported.

Systemic arsenic acid toxicity

The systemic features observed are similar for each source of exposure
which are considered together.

General toxic effects

Patients with chronic arsenic acid poisoning may present with general
debility, progressive weakness (Feinglass, 1973; Gerhardt et al, 1980)
fever and sweats (Heyman et al, 1956).

Dermal toxicity

The characteristic dermal manifestations are hyperkeratosis and
“raindrop” pigmentation of the skin (Heyman et al, 1956; Kyle and
Pease, 1965; Shannon and Strayer, 1989). Hyperkeratoses appear as
multiple small nodules which may coalesce to form plaques and are
found most commonly on the palms and soles.

By contrast, hyperpigmentation is more prominent in the axilla, groin,
areola and around the waist, typically with mucosal sparing (Shannon
and Strayer, 1989). These changes seem to be exacerbated by poor
nutritional status (Das et al, 1995).

Hyperkeratotic lesions may develop into squamous cell carcinomas which
are notable for their occurrence on non light-exposed areas of the
upper extremities and trunk (Shannon and Strayer, 1989).

The fingernails may become brittle with transverse white striae (Mees’
lines) (Mees, 1919; Heyman et al, 1956; Kyle and Pease, 1965; Gerhardt
et al, 1980).

Exfoliative dermatitis (Nicolis and Helwig, 1973) has been reported.

Neuropsychological toxicity

A symmetrical peripheral neuropathy is typical. Sensory symptoms
predominate with paraesthesiae, numbness and pain, particularly of the
soles of the feet, extending in a “glove and stocking” distribution
(Jenkins, 1966; Gerhardt et al, 1980).

Motor involvement with symmetrical distal limb weakness, muscle
atrophy and loss of deep tendon reflexes is recognized (Heyman et al,
1956; Gerhardt et al, 1980; Bansal et al, 1991).

Complete respiratory muscle paralysis (Greenberg et al, 1979; Gerhardt
et al, 1980), a phrenic neuropathy (Bansal et al, 1991) and cranial
nerve involvement (Schoolmeester and White, 1980) have been reported.
The neuropathy may be confused with the Guillain-Barré syndrome (Kyle
and Pease, 1965; Donofrio et al, 1987). Gastrointestinal symptoms and
skin manifestations suggest arsenic poisoning, while a high CSF
protein concentration and cranial nerve involvement are more typical
of the Guillain-Barré syndrome.

Electromyelography may show reduced peripheral nerve conduction
velocities in the absence of symptoms.

Psychological impairment is widely reported in chronic arsenical
poisoning with defects of verbal learning ability and memory and
personality changes (Heyman et al, 1956; Schoolmeester and White,

Hutton et al (1982) described a case of chronic self-intoxication with
sodium arsenate ant poison. The patient was initially admitted with
gastrointestinal symptoms and pancytopenia. He subsequently developed
severe peripheral neuropathy and myelopathy. Urinalysis revealed an
arsenic concentration of 3600 mg/L. The patient eventually admitted
self-administering arsenic in order to secure early retirement on
medical grounds.

Gastrointestinal toxicity

Nausea is common in chronic arsenical poisoning. In the German
literature Reinl (1970) reported a 53 year-old man who had been
exposed to arsenic acid used as an oxidizing agent in chemical
manufacture. During his ten year employment he developed symptoms
including diarrhoea and vomiting. He died of lung cancer believed to
be related to arsenic exposure.


Abnormal liver enzyme activities (Schoolmeester and White, 1980) have
been observed in chronic arsenic poisoning.

Arsenic-induced cirrhosis has been described but may be explained by
concomitant excess ethanol consumption (Morton and Dunnette, 1994).

Narang (1987) suggested increased arsenic consumption as a
contributing factor in the aetiology of liver disease in the Indian
population when he found significantly increased hepatic arsenic
concentrations at autopsy in 178 patients dying from cirrhosis, non
cirrhotic portal fibrosis, fulminant hepatitis, Wilson’s disease or
alcoholic liver disease.

In the German literature cirrhosis and splenomegaly were reported at
autopsy of a 53 year-old man occupationally exposed to arsenic acid
for ten years. Dermal and gastrointestinal symptoms occurred during
his employment and arsenic induced lung cancer was cited as the cause
of death (Reinl, 1970).


Renal manifestations probably reflect capillary damage and include
haematuria, proteinuria with casts and acute tubular or cortical
necrosis (Morton and Dunnette, 1994).

Peripheral vascular and cardiovascular toxicity

“Black foot disease” refers to a severe form of peripheral vascular
disease seen in Taiwan in those who drink artesian well water with an
high arsenic concentration. Initial paraesthesiae and cold sensitivity
progress to ulceration and gangrene (Chiou et al, 1995). It has been
suggested that mortality due to all vascular diseases may be increased
in these populations (Chen and Lin, 1994; Engel et al, 1994).

Raynaud’s syndrome has also been described in those chronically
exposed to arsenic dust.

Several authors refer to the myocardial toxicity of arsenic
(Schoolmeester and White, 1980; Hall and Harruff, 1989) which has been
attributed to impaired oxidative metabolism of myocardial tissue plus
a direct arsenic-induced inflammatory process. A 42 year-old
agricultural worker presented with systemic features of chronic

arsenic poisoning (neuropathy and skin lesions) and had a 24 hour
urine arsenic excretion of 7000 µg (Hall and Harruff, 1989). He
received a 15 day course of dimercaprol with some improvement in motor
function. On the 26th day of hospital admission he suddenly collapsed
and died following a cardiac arrest. At post-mortem he had a diffuse
interstitial myocarditis which was assumed to have triggered a fatal


Pancytopenia has been reported in cases of chronic intoxication by
arsenic acid salts (Schoolmeester and White, 1980; Hutton et al,
1982). Anaemia, neutropenia (Heyman et al, 1956; Kyle and Pease,
1965), or evidence of haemolysis (Kyle and Pease, 1965) have also been
reported as have macrocytosis without anaemia (Heaven et al, 1994) and
a myelodysplastic syndrome (Rezuke et al, 1991).

Chronic arsenic exposure complicated by aplastic anaemia may
predispose to acute myeloid leukaemia (Kjeldsberg and Ward, 1972).

Disrupted haem metabolism with altered urinary porphyrin excretion
(Garcia-Vargas et al, 1994) has been reported.

Pulmonary toxicity

An irritating cough and haemoptysis have been reported in chronic
arsenic intoxication (Heyman et al, 1956).

Endocrine toxicity

Epidemiological evidence from Taiwan (Lai et al, 1994) has recently
associated chronic arsenic exposure with the development of diabetes


Dermal exposure

Surface decontamination should be attempted where necessary. Treat
burns conventionally. Consider the possibility of systemic arsenic
poisoning and the need for chelation therapy (see below).

Ocular exposure

Irrigate the eye with copious lukewarm water. A topical anaesthetic
may be necessary for pain relief. Seek an ophthalmic opinion if
symptoms persist or examination is abnormal.



After acute ingestion of a substantial quantity of arsenic acid most
patients will vomit spontaneously but, in those who do not, gastric
lavage should be considered only if it is possible to undertake the
procedure within the first hour.

Supportive measures

Severe acute arsenic acid poisoning requires prompt intensive
resuscitation with adequate fluid replacement and close observation of
vital signs including cardiac monitoring. Diarrhoea may be treated
symptomatically with loperamide. Chelation therapy should be
considered in symptomatic cases. Obtain blood and urine for arsenic
concentration determination.

Electrocardiographic evidence of QT prolongation in arsenic poisoning
may precede atypical ventricular arrhythmias, notably torsade de
pointes, and in these circumstances drugs which themselves prolong the
QT interval, such as procainamide, quinidine or disopyramide, should
be avoided. Isoprenaline is effective; phenytoin, lignocaine or
propranolol are alternatives (Goldsmith and From, 1980).


Immediate management involves removal from exposure and administration
of supplemental oxygen if necessary. Evidence of systemic arsenic
uptake should be sought and chelation therapy considered as discussed


Chelating agents used in the treatment of arsenic poisoning are
dithiol compounds which can remove arsenic from endogenous sulphydryl
groups, the targets of arsenic toxicity (Jones, 1995).

Traditionally, dimercaprol (British anti-lewisite, BAL) has been the
recommended chelator in arsenic intoxication (Jenkins, 1966; Greenberg
et al, 1979; Roses et al, 1991). However, dimercaprol may produce
unpleasant adverse effects and must be administered by deep
intramuscular injection. There is increasing evidence that
dimercaptosuccinic acid (DMSA, Succimer) (Aposhian et al, 1984;
Graziano, 1986; Fournier et al, 1988; Inns et al, 1990) and
dimercaptopropane sulphonate (DMPS, Unithiol) (Aposhian, 1983;
Aposhian et al, 1984; Hruby and Donner, 1987; Inns et al, 1990) are
less toxic and may be preferable. DMSA and DMPS are more effective in
reducing the arsenic content of tissues, they increase biliary as well
as urinary arsenic elimination and, unlike dimercaprol, do not appear
to cause arsenic accumulation in the brain (Kreppel et al, 1990; Moore
et al, 1994). On the other hand, arsenic mercaptide (the chelation
complex of dimercaprol and arsenic) is dialysable and hence

dimercaprol may be preferred in the presence of renal failure (Sheabar
et al, 1989; Mathieu et al, 1992)

The importance of an increased urine arsenic concentration in
determining the need for chelation therapy is disputed. Kersjes et al
(1987) suggested a spot urine concentration greater than 200 µg/L
should be taken as an indication of “significant” arsenic exposure but
Kingston et al (1993) emphasised that arsenic concentrations
significantly higher than this (3500 µg/24 h and 5819 µg/24 h in two
of their patients) may be observed in the acute phase following
pentavalent arsenic ingestion without severe sequelae.

Dimercaprol (British anti-lewisite; BAL)

Dimercaprol was developed during the Second World War as an antidote
for lewisite (dichloro(2-chlorovinyl) arsine) poisoning (Peters et al,
1945). It possesses two sulphydryl groups and forms a stable
mercaptide ring with arsenic. The alcohol group on dimercaprol confers
some degree of water solubility, thereby enhancing excretion from the
body. As the chelation complex tends to dissociate it is necessary to
maintain a constant excess of dimercaprol. Unlike DMSA and DMPS,
dimercaprol is also lipid soluble and increases the brain arsenic
concentration in arsenic-intoxicated animals (Jones, 1995).

Though increasingly superseded by the less toxic thiol chelating
agents, intramuscular dimercaprol remains useful in severe arsenic
poisoning where vomiting prevents oral antidote administration,
supplies of DMSA or DMPS are not rapidly available (Jolliffe et al,
1991) or renal failure requires haemodialysis; dimercaprol but not
DMSA chelates can cross the dialysis membrane (Sheabar et al, 1989;
Mathieu et al, 1992).

Animal studies

Stocken and Thompson (1946) demonstrated increased urine arsenic
excretion (up to 33.5 per cent of the amount applied) in the 24 hours
following cutaneous application of lewisite to rodents, when
dimercaprol (dose not stated) was spread over the affected area up to
one hour later. Dimercaprol also prevented arsenic-induced diarrhoea
observed in control animals.

Intravenous injection of dimercaprol glucoside 1.5 g/kg prevented
death in two rabbits poisoned with cutaneous lewisite (12 mg/kg).
Eleven control animals died, as did two treated with subcutaneous
dimercaprol 0.07 g/kg (Danielli et al, 1947).

A recent study has demonstrated that intramuscular dimercaprol
protects rabbits against the lethal systemic effects of intravenously
administered lewisite. No appreciable difference was found between the
protective effect of dimercaprol and that of water soluble analogues
DMPS and DMSA (Inns et al, 1990).

Clinical studies

In a case series, 12 men were exposed to smoke containing
diphenylcyano-arsenic (1.6 mg/m3), “other forms of organic arsenic”
(0.5 mg/m3) and “inorganic arsenic” (1.8 mg/m3) for six minutes.
They were treated with 3.5 mg/kg intramuscular dimercaprol 6.5-78
hours post exposure. Urine arsenic excretion increased by an average
of 40 per cent between two and four hours after the injection. The
largest increase, both absolute and relative, was observed in those
treated earliest (6.5 hours after exposure) (Wexler et al, 1946).

Giberson et al (1976) described the treatment of a 44 year-old male
who ingested 400 mg sodium arsenite. Intramuscular dimercaprol 250 mg
was administered every four hours. Haemodialysis was initiated in
response to renal failure with 3.3 mg arsenic removed over four hours.
By the sixth day, when renal function had recovered, arsenic excretion
had reached 75 mg/24h with at least 115 mg arsenic excreted between
days two and six.

A four year-old boy who had ingested an unknown amount of arsenic
trioxide rat poison was treated with dimercaprol 5 mg/kg every four
hours for 16 hours. The urine contained 2,120 µg arsenic over the
first 12 hours. He developed an urticarial rash over the lower
extremities which subsided with the discontinuation of dimercaprol.
The urine arsenic concentration decreased gradually during
d-penicillamine treatment (Peterson and Rumack, 1977).

Schoolmeester and White (1980) reported a 16 year-old female who
ingested 300 mg sodium arsenate in a suicide attempt. She received
intramuscular dimercaprol 125 mg every four hours for the first 24
hours, then twice daily for 24 hours. A 24 hour urine arsenic
concentration (starting time not specified) was 14,200 µg/L. The
effect of chelation therapy on arsenic excretion is not known but the
patient fully recovered.

Mahieu et al (1981) described a 44 year-old male who ingested an
unknown amount of arsenic trioxide which had been mistaken for sugar.
The dose “certainly exceeded 1000 mg”. Intramuscular dimercaprol 2.5-4
mg/kg tds was administered for 21 days. Initial arsenic excretion was
low due to renal insufficiency but increased to 10 mg/24h from three
to seven days post ingestion. The patient excreted a total of 129 mg
arsenic during his 26 days in hospital. A 40 year-old woman poisoned
at the same time and treated with the same regimen for 17 days
excreted 16.7 mg arsenic on the first day, the amount decreasing on
subsequent days. Seventy three milligrams arsenic were eliminated over
three weeks.

A 32 year-old man who ingested 900 mg sodium arsenate in a suicide
attempt commenced treatment with intramuscular dimercaprol 5 mg/kg
four hourly five hours later (Bansal et al, 1991). Dimercaprol was
stopped on day four. This patient also received oral d-penicillamine
and intravenous then oral N-acetylcysteine between days two and 82
post ingestion. The urine arsenic concentration rose on the second

hospital day then declined progressively during the next week although
the data were incomplete and uninterpretable.

A 22 month-old female who developed diarrhoea, vomiting and lethargy
after ingesting approximately 0.7 mg sodium arsenate was treated
initially with one intramuscular dose of dimercaprol 3 mg/kg nine
hours post ingestion (Cullen et al, 1995). Three hours later the
infant was asymptomatic and dimercaprol therapy discontinued although
she subsequently received oral d-penicillamine then oral DMSA to treat
persisting high urine arsenic concentrations (4880 µg/L in the first
24 hours after admission). On the third hospital day the urine arsenic
concentration (from a 24 hour collection) was 1355 µg/L and fell
progressively to 96 µg/L on day 12. These data do not enable any
conclusions to be drawn regarding enhanced arsenic elimination.

No benefit from dimercaprol was reported by McCutchen and Utterback
(1966) in the treatment of severe chronic arsenic poisoning. Other
authors have reported disappointing results with dimercaprol in the
management of arsenic neuropathy (Heyman et al, 1956) although Jenkins
(1966) described “no detectable disability” 18 months after acute
sodium arsenite ingestion in a patient who developed a peripheral
neuropathy and received “a full course of dimercaprol” (details not

Marcus (1987) described a 16 year-old male who survived ingestion of
56 mg arsenic trioxide following treatment with intramuscular
dimercaprol 4 mg/kg every four hours (duration not stated). The
maximum urine arsenic excretion was “over 50 mg/day” falling to 20
µg/day by day 31. At twelve month follow-up neurological effects

Mahieu et al (1981) suggested that a high (greater than 90 per cent)
proportion of methylated arsenic in the urine of poisoned patients
could be used to indicate a late presentation with less likelihood of
benefit from chelation therapy.

Treatment protocol

Dimercaprol must be given by deep intramuscular injection. After
injection 90 per cent of an administered dose is absorbed and Cmax is
attained within one hour (Peters et al, 1947). Dimercaprol is
distributed throughout the intracellular space and metabolic
degradation and excretion is complete in less than four hours.
Depending on severity, 2.5-5 mg/kg should be administered four hourly
for two days. This is to ensure that a constant excess of dimercaprol
is always present as the chelation complex dissociates. Traditionally,
this initial treatment is followed by 2.5 mg/kg bd intramuscularly for
one to two weeks. However, this is an empirical recommendation and may
be insufficient in severe cases. Dosage and duration should be
adjusted therefore, depending on urine arsenic removal.

Adverse effects

The most common adverse effect of dimercaprol is dose-related
hypertension (with an increase in systolic pressure of up to 50 mmHg)
which usually resolves within three hours of administration (Dollery,
1991) but may be associated with nausea, headache, sweating and
abdominal pain. Gastrointestinal disturbance may also occur without
hypertension. Conjunctivitis, paraesthesiae and fever have been

Dimercaprol is contraindicated in severe liver disease since it is
metabolized by glucuronidation with subsequent biliary excretion.


DMSA is commercially available in some countries (though not the UK)
mainly as meso-DMSA, although a DL-form also exists.

Animal studies

Aposhian et al (1984) demonstrated that DMSA was moderately more
effective than DMPS (and substantially more effective than
dimercaprol) in protecting mice from the lethal effects of sodium
arsenite. DMSA mobilizes arsenic from tissues, increasing urine
arsenic excretion without a rise in brain arsenic concentrations
(Aposhian et al, 1984).

Mice administered subcutaneous arsenic trioxide (5 mg/kg) followed
immediately by intraperitoneal DMSA 100 mg/kg, showed significantly
increased urine arsenic excretion (p<0.01) in the first 12 hours post
chelation although the 48 hour urine arsenic elimination was not
significantly different between DMSA-treated mice and controls
(Maehashi and Murata, 1986).

In animal studies DMSA protected against the embryotoxic effects of
sodium arsenite but only when given within one hour of exposure
(Bosque et al, 1991).

Recent experiments suggest that oral monoester DMSA analogues may
offer renal protection in arsenic poisoning by increasing the enteral
arsenic content to enhance faecal rather than renal elimination
(Hannemann et al, 1995). In other animal studies lipophilic DMSA
analogues were inferior to the parent compound as arsenic antidotes
(Kreppel et al, 1993).

Clinical studies

Lenz et al (1981) described a 46 year-old man who ingested 200 mg
arsenic and survived following treatment with oral DMSA 300 mg qds for
three days.

Kosnett and Becker (1987) reported an increase in the 24 hour urine
arsenic excretion from 26 µg to a maximum of 340 µg on the second day
of oral DMSA 660 mg tds in a patient who presented more than 30 days
after malicious acute arsenic ingestion.

Nine days after ingesting approximately 0.7 mg of a soluble arsenic
salt a 22 month-old female was treated with oral DMSA 30 mg/kg/day for
at least four days (Cullen et al, 1995). The child had already
received chelation therapy with dimercaprol and d-penicillamine, but
further treatment was instituted because of a persistently raised
urine arsenic concentration (650 µg/L on day five). Four days later
the urine arsenic concentration had fallen to 96 µg/L. The authors
reported an overall urine arsenic half-life of 2.6 days. Although the
child initially experienced vomiting, diarrhoea and lethargy these
features resolved within 12 hours and renal and hepatic function
remained normal throughout (Cullen et al, 1995).

There was no objective improvement in the neurological manifestations
of chronic arsenic poisoning in a man poisoned by an ethnic remedy
despite two weeks therapy with oral DMSA 400 mg tds (Kew et al, 1993).
No urine arsenic excretion data were given.

A 33 year-old woman with acute-on-chronic lead and arsenic poisoning
from a herbal remedy clinically recovered following two one-week
courses of oral DMSA 270 mg tds, though the effect of chelation
therapy on urine arsenic excretion is difficult to interpret
(Mitchell-Heggs et al, 1990).

Treatment protocol

DMSA is given orally in a dose of 30 mg/kg body weight per day; an
intravenous preparation is available in some countries and may be
preferable if the patient is vomiting (Hantson et al, 1995).

Adverse effects

Side-effects following treatment with DMSA are rare but include skin
rashes, gastrointestinal disturbance, elevation of serum transaminase
activities and flu-like symptoms (Reynolds, 1993). DMSA should be used
with caution in patients with impaired renal function or a history of
hepatic disease (Reynolds, 1993).


Animal studies

DMPS is commercially available as a racemic mixture of the
dextro-rotatory and levo-rotatory forms which appear to be equally
effective arsenic chelators (Aposhian, 1983), though animal studies
suggest DMSA may be superior to either (Aposhian et al, 1984).

Urine arsenic elimination of arsenic-poisoned rats in the 48 hours
post treatment with DMPS 100 mg/kg intraperitoneally was significantly
lower (p<0.05) than in either control (5 mg/kg subcutaneous arsenic
trioxide only) or DMSA-treated mice (Maehashi and Murata, 1986).
However DMPS significantly increased (p<0.01) faecal arsenic
elimination in the 24 hours post chelation compared to control or DMSA
treated mice, suggesting biliary excretion of the DMPS-arsenic chelate
(Maehashi and Murata, 1986).

Other authors have noted enhanced biliary but not faecal arsenic
excretion following parenteral DMPS administration to arsenic-poisoned
experimental animals. This suggests enterohepatic circulation of the
chelate, which Reichl et al (1995) attempted to block using oral
cholestyramine. They demonstrated enhanced faecal arsenic elimination
(p<0.05) when intraperitoneal DMPS 0.1 mmol/kg and subcutaneous
arsenic trioxide (0.02 mmol/kg) administration was followed by an oral
combination of cholestyramine (0.2 g/kg) and DMPS 0.1 mmol/kg (Reichl
et al, 1995).

Domingo et al (1992) demonstrated a protective effect of DMPS 150-300
mg/kg, but not dimercaprol, against experimental arsenite-induced
embryotoxicity and teratogenicity as judged by the incidence of foetal
malformation or death in mice administered intraperitoneal sodium
arsenite (12 mg/kg) on day nine of gestation.

Clinical studies

Two men inadvertently ingested 1 g and 4 g arsenic trioxide
respectively (Moore et al, 1994). The more severely poisoned patient
developed acute renal failure and 26 hours post ingestion had a blood
arsenic concentration of 400 µg/L. He received intravenous DMPS 5
mg/kg every four hours for six days then oral DMPS 400 mg every four
hours for one week. The other patient had a blood arsenic
concentration of 98 µg/L, 36 hours post ingestion and received a
shorter course of intravenous then oral DMPS. Both patients recovered
fully but quantitative data showing the effect of chelation therapy on
urine arsenic elimination were documented poorly.

In another report there was no objective improvement in the
neurological manifestations of chronic arsenic poisoning in a patient
treated with oral DMPS 100 mg tds for three weeks (Kew et al, 1993).

Treatment protocol

DMPS is given orally or parenterally in a dose of 30 mg/kg body weight
per day.

Adverse effects

Side effects following treatment with DMPS are infrequent but have
included allergic skin reactions, nausea, vertigo and pruritis
(Aposhian, 1983).


Animal studies

d-Penicillamine has been reported to be as effective as dimercaprol
and NAC in prolonging the survival time of mice injected with a lethal
dose of sodium arsenite (Shum et al, 1981). Other studies have
disputed the validity of these results and have failed to demonstrate
d-penicillamine as a useful chelator (Aposhian, 1982; Kreppel et al,

Clinical studies

Peterson and Rumack (1977) described three children who shared a
bottle of rat poison containing arsenic trioxide 1.75 per cent. One
died within hours following a rapidly deteriorating course of coma,
convulsions and cardiac arrhythmias. The second, a four year-old male,
presented with lethargy, a sinus tachycardia and tachypnoea. Oral
d-penicillamine 25 mg/kg qds replaced dimercaprol treatment after 16
hours when the patient developed an urticarial rash over the lower
extremities. The first twelve-hour urine collection during dimercaprol
treatment contained 2,120 µg arsenic with the urine arsenic
concentration decreasing during the five days d-penicillamine therapy.
The child made a full recovery.

The third patient (Peterson and Rumack, 1977) had no severe features
of toxicity at presentation. He received the same chelation therapy
regimen as patient 2. On the second day post ingestion the 24 hour
urine arsenic excretion was 300 µg, increasing in the next 24 hours
(the second day of d-penicillamine therapy) to approximately 800 µg.
This patient also recovered fully.

A one year-old child ingested 15-20 mg sodium arsenate (as ant poison)
and was treated within six hours with 5 mg/kg intramuscular
dimercaprol (Peterson and Rumack, 1977). The chelating agent was then
changed to oral d-penicillamine 100 mg/kg/day and continued for five
days. An initial 12 hour urine collection (commenced approximately six
hours post ingestion) contained 192 µg arsenic, increasing to 2000 µg
arsenic in the next 24 hours before falling to approximately 200 µg/24
h on day two. These authors advocated d-penicillamine 100 mg/kg/day as
the treatment of choice in arsenic poisoning (where oral therapy is
possible). They recommended d-penicillamine should be continued until
the 24 hour urine arsenic excretion is less than 50 µg (Peterson and
Rumack, 1977).

A 16 month-old child was given a five day course of oral
d-penicillamine 250 mg qds 14 hours after ingesting 9-14 mg arsenic
trioxide (Watson et al, 1981). Clinical features of toxicity
(diarrhoea, vomiting and lethargy) resolved within 24 hours and the
child was discharged on day three. The arsenic concentration in urine
collected during the first day of treatment was 560 µg/L. However, no
earlier urine arsenic concentrations were measured and prior to

d-penicillamine therapy the patient had received 185 mg dimercaprol
over 18 hours (Watson et al, 1981).

DiNapoli et al (1989) instituted d-penicillamine therapy in a patient
unable to tolerate intramuscular dimercaprol following intravenous
sodium arsenite injection. d-Penicillamine 500 mg tds was administered
and after ten days a 24 hour urine arsenic excretion of 2 mg was
reported. There were no symptoms of bone marrow depression, haemolysis
or peripheral neuropathy. After a further ten days treatment the urine
arsenic concentration was 20 µg/L.

Bansal et al (1991) described a 35 year-old man with severe arsenic
polyneuropathy involving the phrenic nerves bilaterally, who recovered
following d-penicillamine therapy 250 mg tds for two weeks (route of
administration was not stated). However, the 24 hour urine arsenic
excretion only rose to 82.4 µg/g creatinine in the first 72 hours of
chelation compared to a pretreatment value of 73.5 µg/g creatinine.

Cullen et al (1995) reported a 22 month-old child who ingested some
0.7 mg sodium arsenate. Following a single dose of dimercaprol 3
mg/kg, oral d-penicillamine therapy was commenced, 250 mg qds for nine
doses. By day four the 24 hour urine arsenic concentration had dropped
from 4880 to 682 µg/L. The child was discharged on day six on oral
d-penicillamine therapy (dose not stated) but readmitted three days
later due to a persistently high urine arsenic excretion (650 µg/L on
day five). At this stage d-penicillamine was replaced by DMSA since
the child had developed an erythematous rash.

Oral d-penicillamine 250 mg qds for seven days failed to increase
urinary arsenic elimination in a patient with chronic arsenic
poisoning whose initial 24 hour urine arsenic excretion was 342 µg
(normal <5 µg/24 h) (Heaven et al, 1994).

In another report the urine arsenic concentration in a 67 year-old man
with arsenic-associated aplastic anaemia had risen to 20,246 µg/L
after four days penicillamine therapy 500 mg qds compared to a
pretreatment concentration of 7840 µg/L (Kjeldsberg and Ward, 1972).
The patient died from acute myeloid leukaemia some six months later.


Animal studies

The survival time of mice injected subcutaneously with a lethal dose
of sodium arsenite (25 mg/kg) was increased significantly (p<0.05) if
intraperitoneal N-acetylcysteine (NAC) 100 mg/kg was administered 30
minutes later. There was no significant difference between this dose
of NAC, dimercaprol 5 mg/kg and d-penicillamine 50 mg/kg as an
antidote under these conditions (Shum et al, 1981).

Clinical studies

Martin et al (1990) reported “remarkable clinical improvement” in a 32
year-old man with severe arsenic poisoning following ingestion of a
soluble salt when he was administered intravenous NAC 70 mg/kg four
hourly after dimercaprol had “failed to improve his condition”.
However urinary arsenic excretion data were poorly documented and
dimercaprol was continued during treatment with NAC.

Antidotes: Conclusions and recommendations

1. There are no controlled clinical trials of chelation therapy in
arsenic poisoning and no conclusive evidence that dithiol
antidotes reverse arsenic-induced neurological damage. On the
present evidence it is difficult to recommend a single preferred
antidote, though in the absence of renal failure DMSA may offer
some advantages over other agents; if renal failure supervenes
dimercaprol and haemodialysis should be employed.

2. Chelation therapy should be considered in symptomatic patients
where there is analytical confirmation of the diagnosis.

3. Although urine arsenic concentrations are useful to confirm the
diagnosis of arsenic poisoning chelation therapy should not be
instituted on the basis of an increased urine arsenic
concentration alone.


Haemodialysis removes arsenic from the blood but achieves less
effective arsenic clearance than chelation therapy when normal renal
function is present. It is indicated therefore only in the presence of
renal failure.

Giberson et al (1976) reported an arsenic dialysis clearance of 87
mL/min. During four hours of dialysis 3360 µg arsenic was removed in a
patient with acute arsenic poisoning complicated by renal failure who
was also receiving 250 mg intramuscular dimercaprol six times daily.
The 24 hour urine arsenic excretion on the same day was 2030 µg though
this increased to 75,000 µg/24 h on the sixth hospital day when renal
function had recovered.

A similar haemodialysis arsenic clearance of 76-87 mL/min was
demonstrated in another patient with acute sodium arsenite
intoxication complicated by acute renal failure (Vaziri et al, 1980).

Levin-Scherz et al (1987) instituted haemodialysis promptly in a
patient who presented 26 hours after ingesting 2 g arsenic trioxide.
The patient also received intramuscular dimercaprol, 300 mg initially
then 180 mg every four hours, but died within 72 hours of ingestion.
The maximum amount of arsenic removed in the dialysate was 2.9 mg.

Mathieu et al (1992) demonstrated a haemodialysis clearance comparable
to some 40-77 per cent of the daily arsenic renal elimination on the
day following diuresis recovery. In this case the total blood
haemodialysis clearance (210 mL/min) exceeded the instantaneous plasma
haemodialysis clearance (mean 85 mL/min), suggesting that some arsenic
removed by haemodialysis originated in erythrocytes. These authors
showed similar haemodialysis arsenic clearance with or without prior
administration of intramuscular dimercaprol 250 mg, and advocated
dimercaprol as the chelating agent of choice in arsenic poisoning
complicated by renal failure, since it does not impair arsenic
dialysis clearance.

Experimental evidence in dogs (Sheabar et al, 1989) suggests DMSA-
arsenic chelates do not pass through the dialyser membrane.


A 37 year-old man presented within four hours of ingesting 90 mL of a
1.5 per cent arsenic trioxide solution (Smith et al, 1981). Although
initially only tachycardic he subsequently became hypotensive and
oliguric. For the first 48 hours he received 200 mg intramuscular
dimercaprol four hourly then d-penicillamine 500 mg qds. Charcoal
haemoperfusion was instituted 11 hours after admission followed by two
hours haemodialysis. These therapies were repeated over the next four
days but “discontinued because of continued good renal function and
lack of clinical response”. Serum arsenic concentrations immediately
post haemoperfusion were slightly higher than pre-haemoperfusion
values, suggesting no benefit.


Blood arsenic concentrations correlate poorly with exposure but may be
useful in chronic poisoning (Morton and Dunnette, 1994).

Arsenic concentrations in hair and nails have been used to indicate
chronic systemic absorption, although their use as biological monitors
of occupational exposure to airborne arsenic is limited by difficulty
in excluding external contamination (Yamamura and Yamauchi, 1980).

Urine arsenic concentrations are the most useful biomonitoring tool,
ideally as 24 hour urine arsenic excretion collection, although spot
urine arsenic concentrations have been proposed in screening
asymptomatic patients with a history of possible acute arsenic
ingestion (Grande et al, 1987).

Since certain marine organisms (especially mussels) may contain large
amounts of organoarsenicals, it is advisable that workers refrain from
eating seafood for at least 48 hours before urine collection (Buchet
et al, 1994). Analytical speciation methods capable of separating
inorganic arsenic and its methylated derivatives from dietary
organoarsenicals partially overcome this problem (Farmer and Johnson,
1990; Buchet et al, 1994). However, Vahter (1994) has suggested that

under certain circumstances these compounds are released from seafood
which can invalidate assessment of inorganic arsenic exposure.

Farmer and Johnson (1990) found that high urine concentrations of
inorganic arsenic plus its mono- and dimethyl derivatives corresponded
to the possible workplace atmospheric arsenic concentrations for those
involved in arsenic production or glass manufacture. Increased urine
arsenic concentrations have also been noted in timber treatment
workers using an arsenic-based wood preservative (Gollop and Glass,

Telolahy et al (1993) suggested a potential role for increased urine
coproporphyrins as an indicator of chronic occupational arsenic
exposure since arsenic is known to disrupt haem metabolism.

Regular examination of the skin should be included in an occupational
health surveillance programme. Workers with evidence of excessive
arsenic exposure should be offered long-term monitoring for the
development of skin, bladder or lung cancer, though in practise this
may be difficult to execute.


Maximum exposure limit

Long-term exposure limit (8 hour TWA reference period) 0.1 mg/m3
(Health and Safety Executive, 1995).



Individuals who chronically ingest arsenic have an increased risk of
developing skin cancer, usually squamous cell carcinoma but also basal
cell carcinomas (Schoolmeester and White, 1980; Chen et al, 1988;
Shannon and Strayer, 1989; Chiou et al, 1995). Squamous cell
carcinomas may arise in areas of arsenic-induced Bowen’s disease
(Shannon and Strayer, 1989).

Hsueh et al (1995) demonstrated a significant dose-response relation
between skin cancer prevalence and arsenic exposure from artesian well
water. These authors identified chronic hepatitis B carriage and
malnutrition as risk factors for arsenic-induced dermatological

Skin cancer has also been documented among vineyard workers and
farmers exposed to inhaled inorganic arsenic in pesticides (Chen and
Lin, 1994) although skin and gastrointestinal absorption probably
contributed to arsenic toxicity in these cases.

There is an association between chronic arsenic exposure and cancer of
the urinary tract (Chen et al, 1988; Chen and Lin, 1994), lung (Chen
and Lin, 1994) and liver, both hepatic angiosarcoma and hepatocellular
carcinoma (Chen and Lin, 1994).

Reinl (1970) reported the death of a 53 year-old man who had been
exposed to arsenic acid for ten years. The compound had been used an
as oxidizing agent in a chemical manufacturing process. He suffered
gastrointestinal and dermal symptoms of arsenic toxicity during his
employment and at autopsy arsenic induced metastatic bronchial
carcinoma was diagnosed as the cause of death.

Smoking exerts a synergistic effect with ingested and inhaled arsenic
in the development of pulmonary malignancy. There is limited evidence
that other internal cancers, particularly of the gastrointestinal
tract and haematological malignancies, are linked aetiologically to
arsenic exposure (Chen and Lin, 1994).


Animal studies suggest arsenic is embryotoxic and teratogenic but
reliable human data are scarce (Council on Scientific Affairs, 1985).

Daya et al (1989) reported a 22 year-old female who ingested 340 mg
sodium arsenate while 20 weeks pregnant. Treatment with dimercaprol
150 mg four hourly was commenced two hours post ingestion, the maximum
24 hour urine arsenic excretion was 3030 µg/L and a healthy infant was
delivered at 36 weeks.

A woman in the third trimester of pregnancy developed acute renal
failure after ingesting a large quantity of an arsenical rat poison.
Her baby was delivered on the fourth day post ingestion but died
within a few hours from hyaline membrane disease. At autopsy the
infant showed significant arsenic accumulation in the liver, brain and
kidneys (liver arsenic concentration 0.74 mg/100 g tissue) (Lugo et
al, 1969).

Genotoxicity (sodium arsenate)

Cultured human peripheral lymphocytes: Induced chromosomal aberrations
and sister chromatid exchanges.

Syrian hamster cells and human lymphocytes: Induced sister chromatid
exchanges and chromosomal aberrations.

Chinese hamster ovary cells: Induced chromosomal aberrations.

Drosophilia melanogaster: Wing spot test negative (sodium arsenate
is highly toxic to Drosophilia and hence could only be tested at
very low concentrations) (DOSE, 1994).

Fish toxicity (arsenic)

EC50 (96 hr) fathead minnow 141-144 mg/L.

LC50 96 hr) knifefish 31 mg/L.

Oral administration (0.52 mg/kg/day for 24 weeks) to rainbow trout
caused chronic inflammatory changes in subepithelial tissues of the
gall bladder wall in 71 per cent of the group.

LC50 (96 hr) striped base 30 mg (DOSE, 1992).

EC Directive on Drinking Water Quality 80/778/EEC

Maximum admissible concentration 50 µg/L, as arsenic (DOSE, 1992).

WHO Guidelines for Drinking Water Quality

Guideline value 10 µg/L, as arsenic (WHO, 1993).


SM Bradberry BSc MB MRCP
WN Harrison PhD CChem MRSC
ST Beer BSc

National Poisons Information Service (Birmingham Centre),
West Midlands Poisons Unit,
City Hospital NHS Trust,
Dudley Road,
B18 7QH

This monograph was produced by the staff of the Birmingham Centre of
the National Poisons Information Service in the United Kingdom. The
work was commissioned and funded by the UK Departments of Health, and
was designed as a source of detailed information for use by poisons
information centres.

Date of last revision


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[Pulmonary cancer caused by arsenic in an industrial worker.]
Zentralbl Arbeitsmed Arbeitsschutz 1970; 20: 75-8.

Reynolds JEF, ed.
Martindale: The Extra Pharmacopoeia.
London: The Pharmaceutical Press, 1993.

Rezuke WN, Anderson C, Pastuszak WT, Conway SR, Firshein SI.
Arsenic intoxication presenting as a myelodysplastic syndrome: a case
Am J Hematol 1991; 36: 291-3.

Roses OE, García Fernández JC, Villaamil EC, Camussa N, Minetti SA,
Martínez de Marco M et al.
Mass poisoning by sodium arsenite.
Clin Toxicol 1991; 29: 209-13.

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In: Tomes plus. Environmental Health and Safety Series I. Vol 26.
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Arsenic poisoning.
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Arsenic-induced skin toxicity.
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INTOX Home Page
This report contains the collective views of international groups of experts and does not necessarily represent the decisions or the stated policy of the United Nations Environment Programme, the International Labour Organization, or the World Health Organization.

Environmental Health Criteria 224

Second edition

The first and second drafts of this monograph were prepared, under the coordination of Dr J. Ng, by the authors A. Gomez-Caminero, P. Howe, M. Hughes, E. Kenyon, D.R. Lewis, M. Moore, J. Ng, and by A. Aitio and G. Becking.

Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

World Health Organization

Geneva, 2001

The International Programme on Chemical Safety (IPCS), established in 1980, is a joint venture of the United Nations Environment Programme (UNEP), the International Labour Organization (ILO), and the World Health Organization (WHO). The overall objectives of the IPCS are to establish the scientific basis for assessment of the risk to human health and the environment from exposure to chemicals, through international peer-review processes, as a prerequisite for the promotion of chemical safety, and to provide technical assistance in strengthening national capacities for the sound management of chemicals.

The Inter-Organization Programme for the Sound Management of Chemicals (IOMC) was established in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO, the United Nations Industrial Development Organization, the United Nations Institute for Training and Research, and the Organisation for Economic Co-operation and Development (Participating Organizations), following recommendations made by the 1992 UN Conference on Environment and Development to strengthen cooperation and increase coordination in the field of chemical safety. The purpose of the IOMC is to promote coordination of the policies and activities pursued by the Participating Organizations, jointly or separately, to achieve the sound management of chemicals in relation to human health and the environment.

WHO Library Cataloguing-in-Publication Data

Arsenic and arsenic compounds.

(Environmental health criteria ; 224)

1.Arsenic – toxicity

2.Arsenicals – toxicity

3.Environmental exposure

I. International Programme on Chemical Safety

II. WHO Task Group on Environmental Health Criteria for Arsenic and Arsenic Compounds


ISBN 92 4 157224 8

(NLM Classification: QV 294)

ISSN 0250-863X

The World Health Organization welcomes requests for permission to reproduce or translate its publications, in part or in full. Applications and enquiries should be addressed to the Office of Publications, World Health Organization, Geneva, Switzerland, which will be glad to provide the latest information on any changes made to the text, plans for new editions, and reprints and translations already available.

©World Health Organization 2001

Publications of the World Health Organization enjoy copyright protection in accordance with the provisions of Protocol 2 of the Universal Copyright Convention. All rights reserved.

The designations employed and the presentation of the material in this publication do not imply the expression of any opinion whatsoever on the part of the Secretariat of the World Health Organization concerning the legal status of any country, territory, city or area or of its authorities, or concerning the delimitation of its frontiers or boundaries.

The mention of specific companies or of certain manufacturers’ products does not imply that they are endorsed or recommended by the World Health Organization in preference to others of a similar nature that are not mentioned. Errors and omissions excepted, the names of proprietary products are distinguished by initial capital letters.





1.1 Properties and analytical procedures

1.2 Sources and occurrence of arsenic in the environment

1.3 Environmental transport and distribution

1.4 Environmental levels and human exposure

1.5 Kinetics and metabolism

1.6 Effects on laboratory animals and in vitro systems

1.7 Effects on human health

1.8 Effects on other organisms in the environment


2.1 Identity

2.2 Chemical and physical properties of arsenic compounds

2.3 Analytical procedures

2.4 Sample preparation and treatment

2.4.1 Sampling and collection

2.4.2 Oxidative digestion

2.4.3 Extraction

2.4.4 Supercritical fluid extraction

2.5 Macro-measurement

2.6 Colorimetric methods

2.7 Methods for total inorganic arsenic

2.8 Atomic spectrometry

2.9 ICP methodologies

2.10 Voltammetry

2.11 Radiochemical methods

2.12 X-ray spectroscopy

2.13 Hyphenated techniques


3.1 Natural sources

3.2 Sources of environmental pollution

3.2.1 Industry

3.2.2 Past agricultural use

3.2.3 Sewage sludge

3.3 Uses


4.1 Transport and distribution between media

4.1.1 Air

4.1.2 Freshwater and sediment

4.1.3 Estuarine and marine water and sediment

4.1.4 Soil

4.2 Biotransformation

4.2.1 Oxidation and reduction

4.2.2 Methylation

4.2.3 Degradation Abiotic degradation Biodegradation

4.2.4 Bioaccumulation Microorganisms Macroalgae Aquatic invertebrates Fish Terrestrial plants Terrestrial invertebrates Birds


5.1 Environmental levels

5.1.1 Air

5.1.2 Precipitation

5.1.3 Surface water

5.1.4 Groundwater

5.1.5 Sediment

5.1.6 Sewage sludge

5.1.7 Soil

5.1.8 Biota Freshwater Marine Terrestrial

5.2 General population exposure

5.2.1 Air

5.2.2 Food and beverages

5.2.3 Drinking-water

5.2.4 Soil

5.2.5 Miscellaneous exposures

5.3 Occupational exposures

5.4 Total human intake of arsenic from all environmental pathways


6.1 Inorganic arsenic

6.1.1 Absorption Respiratory deposition and absorption Gastrointestinal absorption Dermal absorption Placental transfer

6.1.2 Distribution Fate of inorganic arsenic in blood Tissue distribution

6.1.3 Metabolic transformation Animal studies Human studies

6.1.4 Elimination and excretion Animal studies Human studies

6.1.5 Retention and turnover Animal studies Human studies

6.1.6 Reaction with body components

6.2 Organic arsenic compounds

6.2.1 Absorption Respiratory deposition and absorption Gastrointestinal absorption Dermal absorption Placental transfer

6.2.2 Distribution Fate of organic arsenic in blood Tissue distribution

6.2.3 Metabolic transformation Animal studies Human studies

6.2.4 Elimination and excretion Animal studies Human studies

6.2.5 Retention and turnover

6.3 Biomarkers of arsenic exposure

6.3.1 Arsenic in hair and nails

6.3.2 Blood arsenic

6.3.3 Arsenic and metabolites in urine


7.1 Inorganic arsenic

7.1.1 Single exposure Acute toxicity data

7.1.2 Short-term exposure Oral Inhalation Dermal Parenteral

7.1.3 Long-term exposure Oral Inhalation Dermal

7.1.4 Skin and eye irritation; sensitization Contact sensitivity

7.1.5 Reproductive toxicity, embryotoxicity, and teratogenicity In vivo embryo and fetal toxicity In vitro embryo and fetal toxicity Teratogenicity Gene expression Induction of heat shock proteins Male reproductive toxicity

7.1.6 Genotoxicity and related end-points Bacteria Mammalian cells Human cells In vivo genotoxicity Mechanism of genotoxicity Resistance/hypersensitivity to arsenic cytotoxicity

7.1.7 Carcinogenicity Pulmonary carcinogenicity Skin tumorigenicity Long-term study in monkeys Long-term study in mice

7.1.8 Other special studies Cardiovascular system Nervous system Skin Immune system Haem biosynthesis and urinary excretion of porphyrins Apoptosis

7.1.9 Factors modifying toxicity; toxicity of metabolites Interactions with other compounds Biological role of arsenic Induction of proteins

7.1.10 Potential mechanisms of toxicity – mode of action Toxicity of trivalent inorganic arsenic Toxicity of pentavalent inorganic arsenic Carcinogenicity

7.2 Organic arsenic compounds

7.2.1 Single exposure Acute toxicity data

7.2.2 Short-term exposure Oral

7.2.3 Long-term exposure Oral Inhalation Dermal

7.2.4 Skin and eye irritation; sensitization

7.2.5 Reproductive toxicity, embryotoxicity, and teratogenicity In vivo embryo and fetal toxicity Teratogenicity

7.2.6 Genotoxicity and related end-points Bacteria Mammalian cells Human cells In vivo genotoxicity Apoptosis

7.2.7 Carcinogenicity Bladder Promotion

7.2.8 Factors modifying toxicity; toxicity of metabolites Interaction with thiols Inhibition of GSH reductase Induction of proteins

7.2.9 Potential mechanisms of toxicity: mode of action Acute toxicity Carcinogenicity


8.1 Short-term effects

8.2 Long-term effects: historical introduction

8.3 Levels of arsenic in drinking-water in epidemiological studies

8.4 Vascular diseases

8.4.1 Peripheral vascular disease

8.4.2 Cardio- and cerebrovascular disease

8.4.3 Hypertension

8.5 Diabetes mellitus

8.6 Neurotoxicity

8.7 Cancer

8.7.1 Exposure via inhalation Lung cancer Cancer at other sites

8.7.2 Exposure via drinking-water

8.7.3 Dermal effects, including skin cancer

8.8 Reproductive toxicity

8.9 Genotoxicity and related end-points


9.1 Laboratory experiments

9.1.1 Microorganisms Water Soil Bacterial resistance to arsenic

9.1.2 Aquatic organisms Macroalgae Aquatic plants Invertebrates Vertebrates

9.1.3 Terrestrial organisms Plants Invertebrates Vertebrates

9.2 Field observations

9.2.1 Microorganisms

9.2.2 Aquatic organisms

9.2.3 Terrestrial organisms Plants Vertebrates


10.1 Effects on human health

10.1.1 Acute effects

10.1.2 Vascular effects

10.1.3 Diabetes mellitus

10.1.4 Neurological effects

10.1.5 Cancer of the lung, bladder, and kidney

10.1.6 Cancer and precancerous lesions of the skin

10.1.7 Cancer at other sites

10.1.8 Reproductive toxicity

10.1.9 Genotoxicity

10.1.10 Supporting data from experimental studies

10.1.11 Conclusions

10.2 Evaluation of effects on the environment

10.2.1 Exposure

10.2.2 Effects

10.2.3 Environmental modification of toxicity

10.2.4 Risk evaluation


11.1 Human health

11.2 Environmental







Every effort has been made to present information in the criteria monographs as accurately as possible without unduly delaying their publication. In the interest of all users of the Environmental Health Criteria monographs, readers are requested to communicate any errors that may have occurred to the Director of the International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland, in order that they may be included in corrigenda.

* * *

A detailed data profile and a legal file can be obtained from the International Register of Potentially Toxic Chemicals, Case postale 356, 1219 Châtelaine, Geneva, Switzerland (telephone no. + 41 22 – 9799111, fax no. + 41 22 – 7973460, E-mail [email protected]).

* * *

This publication was made possible by grant number 5 U01 ES02617-15 from the National Institute of Environmental Health Sciences, National Institutes of Health, USA, and by financial support from the European Commission.

The Commonwealth Department of Health and Aged Care, Australia, contributed financially to the preparation of this Environmental Health Criteria monograph. The Task Group meeting was arranged by the National Research Centre for Environmental Toxicology, Australia.

Environmental Health Criteria



In 1973 the WHO Environmental Health Criteria Programme was initiated with the following objectives:


to assess information on the relationship between exposure to environmental pollutants and human health, and to provide guidelines for setting exposure limits;


to identify new or potential pollutants;


to identify gaps in knowledge concerning the health effects of pollutants;


to promote the harmonization of toxicological and epidemiological methods in order to have internationally comparable results.

The first Environmental Health Criteria (EHC) monograph, on mercury, was published in 1976 and since that time an ever-increasing number of assessments of chemicals and of physical effects have been produced. In addition, many EHC monographs have been devoted to evaluating toxicological methodology, e.g. for genetic, neurotoxic, teratogenic and nephrotoxic effects. Other publications have been concerned with epidemiological guidelines, evaluation of short-term tests for carcinogens, biomarkers, effects on the elderly and so forth.

Since its inauguration the EHC Programme has widened its scope, and the importance of environmental effects, in addition to health effects, has been increasingly emphasized in the total evaluation of chemicals.

The original impetus for the Programme came from World Health Assembly resolutions and the recommendations of the 1972 UN Conference on the Human Environment. Subsequently the work became an integral part of the International Programme on Chemical Safety (IPCS), a cooperative programme of UNEP, ILO and WHO. In this manner, with the strong support of the new partners, the importance of occupational health and environmental effects was fully recognized. The EHC monographs have become widely established, used and recognized throughout the world.

The recommendations of the 1992 UN Conference on Environment and Development and the subsequent establishment of the Intergovernmental Forum on Chemical Safety with the priorities for action in the six programme areas of Chapter 19, Agenda 21, all lend further weight to the need for EHC assessments of the risks of chemicals.


The criteria monographs are intended to provide critical reviews on the effect on human health and the environment of chemicals and of combinations of chemicals and physical and biological agents. As such, they include and review studies that are of direct relevance for the evaluation. However, they do not describe every study carried out. Worldwide data are used and are quoted from original studies, not from abstracts or reviews. Both published and unpublished reports are considered and it is incumbent on the authors to assess all the articles cited in the references. Preference is always given to published data. Unpublished data are used only when relevant published data are absent or when they are pivotal to the risk assessment. A detailed policy statement is available that describes the procedures used for unpublished proprietary data so that this information can be used in the evaluation without compromising its confidential nature (WHO (1990) Revised Guidelines for the Preparation of Environmental Health Criteria Monographs. PCS/90.69, Geneva, World Health Organization).

In the evaluation of human health risks, sound human data, whenever available, are preferred to animal data. Animal and in vitro studies provide support and are used mainly to supply evidence missing from human studies. It is mandatory that research on human subjects is conducted in full accord with ethical principles, including the provisions of the Helsinki Declaration.

The EHC monographs are intended to assist national and international authorities in making risk assessments and subsequent risk management decisions. They represent a thorough evaluation of risks and are not, in any sense, recommendations for regulation or standard setting. These latter are the exclusive purview of national and regional governments.


The layout of EHC monographs for chemicals is outlined below.

• Summary – a review of the salient facts and the risk evaluation of the chemical

• Identity – physical and chemical properties, analytical methods

• Sources of exposure

• Environmental transport, distribution and transformation

• Environmental levels and human exposure

• Kinetics and metabolism in laboratory animals and humans

• Effects on laboratory mammals and in vitro test systems

• Effects on humans

• Effects on other organisms in the laboratory and field

• Evaluation of human health risks and effects on the environment

• Conclusions and recommendations for protection of human health and the environment

• Further research

• Previous evaluations by international bodies, e.g. IARC, JECFA, JMPR

Selection of chemicals

Since the inception of the EHC Programme, the IPCS has organized meetings of scientists to establish lists of priority chemicals for subsequent evaluation. Such meetings have been held in Ispra, Italy, 1980; Oxford, United Kingdom, 1984; Berlin, Germany, 1987; and North Carolina, USA, 1995. The selection of chemicals has been based on the following criteria: the existence of scientific evidence that the substance presents a hazard to human health and/or the environment; the possible use, persistence, accumulation or degradation of the substance shows that there may be significant human or environmental exposure; the size and nature of populations at risk (both human and other species) and risks for environment; international concern, i.e. the substance is of major interest to several countries; adequate data on the hazards are available.

If an EHC monograph is proposed for a chemical not on the priority list, the IPCS Secretariat consults with the Cooperating Organizations and all the Participating Institutions before embarking on the preparation of the monograph.


The order of procedures that result in the publication of an EHC monograph is shown in the flow chart on p. xvii. A designated staff member of IPCS, responsible for the scientific quality of the document, serves as Responsible Officer (RO). The IPCS Editor is responsible for layout and language. The first draft, prepared by consultants or, more usually, staff from an IPCS Participating Institution, is based initially on data provided from the International Register of Potentially Toxic Chemicals, and reference data bases such as Medline and Toxline.

The draft document, when received by the RO, may require an initial review by a small panel of experts to determine its scientific quality and objectivity. Once the RO finds the document acceptable as a first draft, it is distributed, in its unedited form, to well over 150 EHC contact points throughout the world who are asked to comment on its completeness and accuracy and, where necessary, provide additional material. The contact points, usually designated by governments, may be Participating Institutions, IPCS Focal Points, or individual scientists known for their particular expertise. Generally some four months are allowed before the comments are considered by the RO and author(s). A second draft incorporating comments received and approved by the Director, IPCS, is then distributed to Task Group members, who carry out the peer review, at least six weeks before their meeting.

The Task Group members serve as individual scientists, not as representatives of any organization, government or industry. Their function is to evaluate the accuracy, significance and relevance of the information in the document and to assess the health and environmental risks from exposure to the chemical. A summary and recommendations for further research and improved safety aspects are also required. The composition of the Task Group is dictated by the range of expertise required for the subject of the meeting and by the need for a balanced geographical distribution.

EHC Preparation Flow Chart

The three cooperating organizations of the IPCS recognize the important role played by nongovernmental organizations. Representatives from relevant national and international associations may be invited to join the Task Group as observers. Although observers may provide a valuable contribution to the process, they can only speak at the invitation of the Chairperson. Observers do not participate in the final evaluation of the chemical; this is the sole responsibility of the Task Group members. When the Task Group considers it to be appropriate, it may meet in camera.

All individuals who as authors, consultants or advisers participate in the preparation of the EHC monograph must, in addition to serving in their personal capacity as scientists, inform the RO if at any time a conflict of interest, whether actual or potential, could be perceived in their work. They are required to sign a conflict of interest statement. Such a procedure ensures the transparency and probity of the process.

When the Task Group has completed its review and the RO is satisfied as to the scientific correctness and completeness of the document, it then goes for language editing, reference checking and preparation of camera-ready copy. After approval by the Director, IPCS, the monograph is submitted to the WHO Office of Publications for printing. At this time a copy of the final draft is sent to the Chairperson and Rapporteur of the Task Group to check for any errors.

It is accepted that the following criteria should initiate the updating of an EHC monograph: new data are available that would substantially change the evaluation; there is public concern for health or environmental effects of the agent because of greater exposure; an appreciable time period has elapsed since the last evaluation.

All Participating Institutions are informed, through the EHC progress report, of the authors and institutions proposed for the drafting of the documents. A comprehensive file of all comments received on drafts of each EHC monograph is maintained and is available on request. The Chairpersons of Task Groups are briefed before each meeting on their role and responsibility in ensuring that these rules are followed.



Dr C. Abernathy, Office of Water/Office of Science and Technology, Health and Ecological Criteria Division, US Environmental Protection Agency, Washington, D.C., USA (Chairperson)

Dr D. Chakraborti, School of Environmental Studies, Jadavpur University, Calcutta, India

Professor J.S. Edmonds, Department of Chemistry, De Montfort University, Leicester, United Kingdom

Dr H. Gibb, US Environmental Protection Agency, National Center for Environmental Assessment, Washington DC, USA

Dr P. Hoet, Industrial and Occupational Medicine Unit, Catholic University of Louvain, Brussels, Belgium

Dr C. Hopenhayn-Rich, Department of Preventive Medicine and Environmental Health, University of Kentucky, Lexington, KY, USA

Mr P.D. Howe, Centre for Ecology and Hydrology, Monks Wood Experimental Station, Abbots Ripton, Huntingdon, Cambridgeshire, United Kingdom

Dr L. Järup, Department of Epidemiology and Public Health, Imperial College School of Medicine, London, United Kingdom

Dr A.A. Meharg, Department of Plant and Soil Science, Aberdeen, United Kingdom

Professor M.R. Moore, Director, Queensland Health Scientific Services and National Research Centre for Environmental Toxicology, Queensland, Australia (Vice-Chairperson)

Dr J. C. Ng, National Research Centre for Environmental Toxicology, Brisbane, Australia

Dr A. Nishikawa, Division of Pathology, National Institute of Health Sciences, Tokyo, Japan

Dr L. Pyy, Director of the Deptartment, Oulu Regional Institute of Occupational Health, Oulu, Finland

Dr M. Sim, Unit of Occupational and Environmental Health, Department of Epidemiology and Preventive Medicine, Monash University, Victoria, Australia

Dr J. Stauber, CSIRO Energy Technology, Lucas Heights Science and Technology Centre, Bangor, NSW, Australia

Professor M. Vahter, Institute of Environmental Medicine, Karolinska Institute, Stockholm, Sweden


Dr P. Imray, Scientific Adviser, Environmental Health Branch, Queensland Health, Brisbane, Australia

Dr L. Tomaska, Canberra, Australia (representing the Australia New Zealand Food Authority)

Mr D. Hughes, MIM Holdings Limited, Brisbane, Australia (representing the Mining Industry)


Dr A. Aitio, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland

Dr G. Becking, Kingston, Ontario, Canada (Adviser to the Secretariat)

Dr K. Buckett, Director DHAC, Public Health Division, Canberra, Australia

Mr P. Callan, Assistant Director, National Health and Medical Research Council, Canberra, Australia

Dr M.F. Hughes, NHEERL/ET/PKB, US Environmental Protection Agency, Research Triangle Park, NC, USA

Dr E.M. Kenyon, NHEERL/ET/PKB, US Environmental Protection Agency, Research Triangle Park, NC, USA

Dr D.R. Lewis, Human Studies Division, NHEERL, US Environmental Protection Agency, Research Triangle Park, NC, USA

Dr M. Younes, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland


The first and second drafts of this monograph were prepared, under the coordination of Dr J. Ng, by the authors A. Gomez-Caminero, P. Howe, M. Hughes, E. Kenyon, D.R. Lewis, M. Moore, J. Ng, and by A. Aitio and G. Becking. The group of authors met at National Health and Environmental Effects Research Laboratory, US. EPA, North Carolina, on 20–22 July 1998.

A WHO Task Group on Environmental Health Criteria for Arsenic and Arsenic Compounds met at the National Research Centre for Environmental Toxicology, Brisbane, Australia, on 15–19 November 1999. The group reviewed the draft and the peer review comments, revised the draft and made an evaluation of the risks for human health and environment from exposure to arsenic and arsenic compounds.

After the meeting, and based on the peer reviewer comments and Task Group advice, Drs Gibb, Hopenhayn-Rich, Järup, Sim, and Aitio revised and updated the section on Effects on Human Health. This section was then sent for review to a selected group of experts.

The document was revised on the basis of the peer review comments received, these revisions were verified, and the document was finalized by a Review Board, consisting of Drs D. Anderson, H. Gibb, L. Järup, M. Sim and A. Aitio, in TNO BIBRA, Carshalton, UK. The document was finally approved by the Task Group in a mail ballot.

The cut-off date for the literature searches for the document was the Task Group meeting, i.e. November 1999, with the exception of the section on effects on human health, for which the last literature searches were performed in November 2000.

Peer review comments at the first stage international review were received from:

Dr J. Ahlers, Umwelt Bundes Amt, Germany

Dr R. Benson, Region VIII, Environmental Protection Agency, USA

Professor GB Bliss, N.N. Petrov’s Research Institute of Oncology, Russian Federation

Dr M. Bolger, Food and Drug Administration, USA

Professor M. Cíkrt, Centre of Industrial Hygiene and Occupational Diseases, Czech Republique

Professor I. Dési, Albert Szent-Györgyi University, Hungary

Professor J Duffus, The Edinburgh Centre for Toxicology, UK

Dr P Edwards, Department of Health, UK

Dr H Falk, Agency for Toxic Substances and Disease Registry, USA

Dr H. Gibb, Environmental Protection Agency, USA

Dr N. Kurzeja European Environmental Bureau, Germany

Dr I. Mangelsdorf, Fraunhofer Institute, Germany

Dr TG Rossman, NYU School of Medicine

Professor H Taskinen, Finnish Institute of Occuational Health

Mr S Tsuda, Ministry of Halth and Welfare, Japan

Dr G. Ungváry, József Fodor National Center for Public Health, Hungary

Professor M. Vahter, Karolinska Institute, Sweden,

Bureau of Chemical Safety, Canada

Elf Atochem North America, USA

Environmental Protection Agency Office of Research and Development, USA


Finnish Institute of Occupational Health, Finland

Comments on the revised section on effects on human health were received from members of the Task Group, and from:

Dr D Anderson, TNO BIBRA International, UK

Dr Michael Bates, Kenepuru Science Centre, New Zealand

Dr R. Calderon, National Health and Environmental Effects Research Laboratory, US EPA

Professor PE Enterline, University of Pittsburgh, USA

Dr A. Gomez-Caminero, National Health and Environmental Effects Research Laboratory, US. EPA

Dr J Lubin, National Cancer Institute, USA

Professor AH Smith, University of California, USA

Dr A. Aitio of the IPCS central unit was responsible for the scientific aspects of the monograph, and Kathleen Lyle for the technical editing.

The efforts of all, especially Queensland Health and the Natinal Research Centre for Environmental Toxicology, Australia, who helped in the preparation and finalization of the monograph are gratefully acknowledged.



atomic absorption spectrometry


ankle–brachial index


atomic fluorescence spectrometry


silver diethyldithiocarbamate


aminolaevulinic acid


anodic stripping voltammetry


adenosine triphosphatase


area under the curve




bioconcentration factor


blackfoot disease

BFD-endemic area

Geographic area in south-western Taiwan, where arsenic-contaminated artesian well water has been used as drinking water, and where BFD is endemic; the area has been also called the “arseniasis” area, or “hyperendemic” area. In this document it is called BFD-endemic area, to differentiate it from other areas e.g. in Taiwan, where high arsenic concentrations in drinking water have been reported


body mass index




chromosome aberrations


Chemical Abstract Service


copper chrome arsenate




complementary DNA


capillary electrophoresis


confidence interval; unless otherwise stated, the 95% CI is given. Accordingly, the term statistically significant in this documents denotes significance at 95% level


cardiovascular disease


sodium dibenzyldithiocarbamate


duplicate diet study


dimethylarsinic acid


dimethylarsinous acid


dimethylarsinic acid thioglycolic acid methyl ester


dimercaptosuccinic acid


differential pulse cathodic stripping voltammetry


disodium arsenate heptahydrate


redox potential




electrothermal atomic absorption spectrometry


flame atomic absorption spectrometry


flame atomic fluorescence spectrometry


frequency ratio


gas chromatography


granulocyte macrophage-colony stimulating factor




oxidized glutathione


guanosine triphosphate


high frequency cell


hydride generation atomic absorption spectrometry


hanging mercury drop electrode


high pressure liquid chromatography


hypoxanthine phosphoribosyltransferase


Hazardous Substances Data Bank


inductively coupled plasma atomic emission spectrometry


inductively coupled plasma mass spectrometry




ischaemic heart disease


liquid chromatography


median lethal concentration


market basket survey


micellar liquid chromatography


monomethylarsonic acid


monomethylarsonous acid


monomethylarsonic acid thioglycolic acid methyl ester




messenger RNA


monosodium methanearsonate


5,10-methylene-tetradrofolate reductase


neutron activation analysis


sodium (bistrifluoroethyl) dithiocarbamate


nucleotide excision repair


4-nitroquinoline oxide


odds ratio


periodate-oxidized adenosine


particle-induced X-ray emission spectrometry


prevalence odds ratio


peripheral vascular disease


replication index


reversed phase liquid chromatography


Registry of Toxic Effects of Chemicals




S-adenosyl methionine


sister chromatid exchange


standard deviation


standard error of mean


scanning electron microscopy


supercritical fluid chromatography


supercritical fluid extraction


standardized incidence ratio


standardized mortality ratio


sheep red blood cell




transforming growth factor


thioglycolic acid methylester




trimethylarsine oxide


time-weighted average


United Nations




X-ray absorption fine structure spectroscopy


X-ray fluorescence

1.1 Properties and analytical procedures
Arsenic is a metalloid widely distributed in the earth’s crust and present at an average concentration of 2 mg/kg. It occurs in trace quantities in all rock, soil, water and air. Arsenic can exist in four valency states: –3, 0, +3 and +5. Under reducing conditions, arsenite (As(III)) is the dominant form; arsenate (As(V)) is generally the stable form in oxygenated environments. Elemental arsenic is not soluble in water. Arsenic salts exhibit a wide range of solubilities depending on pH and the ionic environment.

There is a variety of instrumental techniques for the determination of arsenic. These include AAS, AFS, ICP-AES, ICP-MS and voltammetry. Some of these (e.g. ICP-MS) can serve as element-specific detectors when coupled to chromatographic separation techniques (e.g. HPLC and GC). These so-called “hyphenated” methods are used for determining individual arsenic species. Additional sensitivity for a limited range of arsenic compounds can often be achieved by the use of hydride generation techniques. A test kit based on the colour reaction of arsine with mercuric bromide is currently used for groundwater testing in Bangladesh and has a detection limit of 50–100 µg/litre under field conditions.

1.2 Sources and occurrence of arsenic in the environment
Arsenic is present in more than 200 mineral species, the most common of which is arsenopyrite.

It has been estimated that about one-third of the atmospheric flux of arsenic is of natural origin. Volcanic action is the most important natural source of arsenic, followed by low-temperature volatilization.

Inorganic arsenic of geological origin is found in groundwater used as drinking-water in several parts of the world, for example Bangladesh.

Organic arsenic compounds such as arsenobetaine, arsenocholine, tetramethylarsonium salts, arsenosugars and arsenic-containing lipids are mainly found in marine organisms although some of these compounds have also been found in terrestrial species.

Elemental arsenic is produced by reduction of arsenic trioxide (As2O3) with charcoal. As2O3 is produced as a by-product of metal smelting operations. It has been estimated that 70% of the world arsenic production is used in timber treatment as copper chrome arsenate (CCA), 22% in agricultural chemicals, and the remainder in glass, pharmaceuticals and non-ferrous alloys.

Mining, smelting of non-ferrous metals and burning of fossil fuels are the major industrial processes that contribute to anthropogenic arsenic contamination of air, water and soil. Historically, use of arsenic-containing pesticides has left large tracts of agricultural land contaminated. The use of arsenic in the preservation of timber has also led to contamination of the environment.

1.3 Environmental transport and distribution
Arsenic is emitted into the atmosphere by high-temperature processes such as coal-fired power generation plants, burning vegetation and volcanism. Natural low-temperature biomethylation and reduction to arsines also releases arsenic into the atmosphere. Arsenic is released into the atmosphere primarily as As2O3 and exists mainly adsorbed on particulate matter. These particles are dispersed by the wind and are returned to the earth by wet or dry deposition. Arsines released from microbial sources in soils or sediments undergo oxidation in the air, reconverting the arsenic to non-volatile forms, which settle back to the ground. Dissolved forms of arsenic in the water column include arsenate, arsenite, methylarsonic acid (MMA) and dimethylarsinic acid (DMA). In well-oxygenated water and sediments, nearly all arsenic is present in the thermodynamically more stable pentavalent state (arsenate). Some arsenite and arsenate species can interchange oxidation state depending on redox potential (Eh), pH and biological processes. Some arsenic species have an affinity for clay mineral surfaces and organic matter and this can affect their environmental behaviour. There is potential for arsenic release when there is fluctuation in Eh, pH, soluble arsenic concentration and sediment organic content. Weathered rock and soil may be transported by wind or water erosion. Many arsenic compounds tend to adsorb to soils, and leaching usually results in transportation over only short distances in soil.

Three major modes of arsenic biotransformation have been found to occur in the environment: redox transformation between arsenite and arsenate, the reduction and methylation of arsenic, and the biosynthesis of organoarsenic compounds. There is biogeochemical cycling of compounds formed from these processes.

1.4 Environmental levels and human exposure
Mean total arsenic concentrations in air from remote and rural areas range from 0.02 to 4 ng/m3. Mean total arsenic concentrations in urban areas range from 3 to about 200 ng/m3; much higher concentrations (> 1000 ng/m3) have been measured in the vicinity of industrial sources, although in some areas this is decreasing because of pollution abatement measures. Concentrations of arsenic in open ocean seawater are typically 1–2 µg/litre. Arsenic is widely distributed in surface freshwaters, and concentrations in rivers and lakes are generally below 10 µg/litre, although individual samples may range up to 5 mg/litre near anthropogenic sources. Arsenic levels in groundwater average about 1–2 µg/litre except in areas with volcanic rock and sulfide mineral deposits where arsenic levels can range up to 3 mg/litre. Mean sediment arsenic concentrations range from 5 to 3000 mg/kg, with the higher levels occurring in areas of contamination. Background concentrations in soil range from 1 to 40 mg/kg, with mean values often around 5 mg/kg. Naturally elevated levels of arsenic in soils may be associated with geological substrata such as sulfide ores. Anthropogenically contaminated soils can have concentrations of arsenic up to several grams per 100 ml.

Marine organisms normally contain arsenic residues ranging from < 1 to more than 100 mg/kg, predominantly as organic arsenic species such as arsenosugars (macroalgae) and arsenobetaine (invertebrates and fish). Bioaccumulation of organic arsenic compounds, after their biogenesis from inorganic forms, occurs in aquatic organisms. Bioconcentration factors (BCFs) in freshwater invertebrates and fish for arsenic compounds are lower than for marine organisms. Biomagnification in aquatic food chains has not been observed. Background arsenic concentrations in freshwater and terrestrial biota are usually less than 1 mg/kg (fresh weight). Terrestrial plants may accumulate arsenic by root uptake from the soil or by adsorption of airborne arsenic deposited on the leaves. Arsenic levels are higher in biota collected near anthropogenic sources or in areas with geothermal activity. Some species accumulate substantial levels, with mean concentrations of up to 3000 mg/kg at arsenical mine sites.

Non-occupational human exposure to arsenic in the environment is primarily through the ingestion of food and water. Of these, food is generally the principal contributor to the daily intake of total arsenic. In some areas arsenic in drinking-water is a significant source of exposure to inorganic arsenic. In these cases, arsenic in drinking-water often constitutes the principal contributor to the daily arsenic intake. Contaminated soils such as mine tailings are also a potential source of arsenic exposure. The daily intake of total arsenic from food and beverages is generally between 20 and 300 µg/day. Limited data indicate that approximately 25% of the arsenic present in food is inorganic, but this depends highly on the type of food ingested. Inorganic arsenic levels in fish and shellfish are low (< 1%). Foodstuffs such as meat, poultry, dairy products and cereals have higher levels of inorganic arsenic. Pulmonary exposure may contribute up to approximately 10 µg/day in a smoker and about 1 µg/day in a non-smoker, and more in polluted areas. The concentration of metabolites of inorganic arsenic in urine (inorganic arsenic, MMA and DMA) reflects the absorbed dose of inorganic arsenic on an individual level. Generally, it ranges from 5 to 20 µg As/litre, but may even exceed 1000 µg/litre.

In workplaces with up-to-date occupational hygiene practices, exposure generally does not exceed 10 µg/m3 (8-h time-weighted average [TWA]). However, in some places workroom atmospheric arsenic concentrations as high as several milligrams per cubic metre have been reported.

1.5 Kinetics and metabolism
Absorption of arsenic in inhaled airborne particles is highly dependent on the solubility and the size of particles. Both pentavalent and trivalent soluble arsenic compounds are rapidly and extensively absorbed from the gastrointestinal tract. In many species arsenic metabolism is characterized by two main types of reactions: (1) reduction reactions of pentavalent to trivalent arsenic, and (2) oxidative methylation reactions in which trivalent forms of arsenic are sequentially methylated to form mono-, di- and trimethylated products using S-adenosyl methionine (SAM) as the methyl donor and glutathione (GSH) as an essential co-factor. Methylation of inorganic arsenic facilitates the excretion of inorganic arsenic from the body, as the end-products MMA and DMA are readily excreted in urine. There are major qualitative and quantitative interspecies differences in methylation, to the extent that some species exhibit minimal or no arsenic methylation (e.g. marmoset monkey, guinea-pig, chimpanzee). However, in humans and most common laboratory animals, inorganic arsenic is extensively methylated and the metabolites are excreted primarily in the urine. Factors such as dose, age, gender and smoking contribute only minimally to the large inter-individual variation in arsenic methylation observed in humans. However, lower methylation efficiency in children has been observed in only one study out of three. Studies in humans suggest the existence of a wide difference in the activity of methyltransferases, and the existence of polymorphism has been hypothesized. Animal and human studies suggest that arsenic methylation may be inhibited at high acute exposures. The metabolism and disposition of inorganic arsenic may be influenced by its valence state, particularly at high dose levels. Studies in laboratory animals indicate that administration of trivalent inorganic arsenic such as As2O3 and arsenite initially results in higher levels in most tissues than does the administration of pentavalent arsenic. However, the trivalent form is more extensively methylated, leading to similar long-term excretion. Ingested organoarsenicals such as MMA, DMA and arsenobetaine are much less extensively metabolized and more rapidly eliminated in urine than inorganic arsenic in both laboratory animals and humans.

Levels of arsenic or its metabolites in blood, hair, nails and urine are used as biomarkers of arsenic exposure. Blood arsenic is a useful biomarker only in the case of acute arsenic poisoning or stable chronic high-level exposure. Arsenic is rapidly cleared from blood, and speciation of its chemical forms in blood is difficult. Arsenic in hair and nails can be indicators of past arsenic exposure, provided care is taken to prevent external arsenic contamination of the samples. Arsenic in hair may also be used to estimate relative length of time since an acute exposure. Speciated metabolites in urine expressed either as inorganic arsenic or as the sum of metabolites (inorganic arsenic + MMA + DMA) provide the best quantitative estimate of recently absorbed dose of arsenic. However, consumption of certain seafood, mainly seaweed and some bivalves, may confound estimation of inorganic arsenic exposure because of metabolism of arsenosugars to DMA in the body or the presence of DMA in the seafood. Such food should be avoided for 2–3 days before urine sampling for monitoring of exposure to inorganic arsenic.

1.6 Effects on laboratory animals and in vitro systems
Both inorganic and organic forms of arsenic may cause adverse effects in laboratory animals. The effects induced by arsenic range from acute lethality to chronic effects such as cancer. The degree of toxicity of arsenic is basically dependent on the form (e.g. inorganic or organic) and the oxidation state of the arsenical. It is generally considered that inorganic arsenicals are more toxic than organic arsenicals, and within these two classes, the trivalent forms are more toxic than the pentavalent forms, at least at high doses. Several different organ systems are affected by arsenic, including skin, respiratory, cardiovascular, immune, genitourinary, reproductive, gastrointestinal and nervous systems.

Several animal carcinogenicity studies on arsenic have been carried out, but limitations such as high dose levels, limited time of exposure and limited number of animals make these inconclusive. However, a recently reported animal model may be a useful tool for future carcinogenicity studies. In that study, female C57B1/6J mice exposed to arsenic in drinking-water containing 500 µg As(V)/litre over 2 years was associated with increased incidence in tumours involving mainly lung, liver, gastrointestinal tract and skin. Inorganic arsenic does not induce point mutations. However, arsenic can produce chromosomal aberrations in vitro, affect methylation and repair of DNA, induce cell proliferation, transform cells and promote tumours. One study has indicated that DMA may cause cancer of the urinary bladder in male rats at high doses.

1.7 Effects on human health
Soluble inorganic arsenic is acutely toxic, and ingestion of large doses leads to gastrointestinal symptoms, disturbances of cardiovascular and nervous system functions, and eventually death. In survivors, bone marrow depression, haemolysis, hepatomegaly, melanosis, polyneuropathy and encephalopathy may be observed.

Long-term exposure to arsenic in drinking-water is causally related to increased risks of cancer in the skin, lungs, bladder and kidney, as well as other skin changes such as hyperkeratosis and pigmentation changes. These effects have been demonstrated in many studies using different study designs. Exposure–response relationships and high risks have been observed for each of these end-points. The effects have been most thoroughly studied in Taiwan but there is considerable evidence from studies on populations in other countries as well. Increased risks of lung and bladder cancer and of arsenic-associated skin lesions have been reported to be associated with ingestion of drinking-water at concentrations £ 50 µg arsenic/litre.

Occupational exposure to arsenic, primarily by inhalation, is causally associated with lung cancer. Exposure–response relationships and high risks have been observed. Increased risks have been observed at cumulative exposure levels ³ 0.75 (mg/m3) × year (e.g. 15 years of exposure to a workroom air concentration of 50 µg/m3). Tobacco smoking has been investigated in two of the three main smelter cohorts and was not found to be the cause of the increased lung cancer risk attributed to arsenic; however, it was found to be interactive with arsenic in increasing the lung cancer risk.

Even with some negative findings, the overall weight of evidence indicates that arsenic can cause clastogenic damage in different cell types with different end-points in exposed individuals and in cancer patients. For point mutations, the results are largely negative.

Chronic arsenic exposure in Taiwan has been shown to cause blackfoot disease (BFD), a severe form of peripheral vascular disease (PVD) which leads to gangrenous changes. This disease has not been documented in other parts of the world, and the findings in Taiwan may depend upon other contributing factors. However, there is good evidence from studies in several countries that arsenic exposure causes other forms of PVD.

Conclusions on the causality of the relationship between arsenic exposure and other health effects are less clear-cut. The evidence is strongest for hypertension and cardiovascular disease, suggestive for diabetes and reproductive effects and weak for cerebrovascular disease, long-term neurological effects, and cancer at sites other than lung, bladder, kidney and skin.

1.8 Effects on other organisms in the environment
Aquatic and terrestrial biota show a wide range of sensitivities to different arsenic species. Their sensitivity is modified by biological and abiotic factors. In general, inorganic arsenicals are more toxic than organoarsenicals and arsenite is more toxic than arsenate. The mode of toxicity and mechanism of uptake of arsenate by organisms differ considerably. This may explain why there are interspecies differences in organism response to arsenate and arsenite. The primary mechanism of arsenite toxicity is considered to result from its binding to protein sulfhydryl groups. Arsenate is known to affect oxidative phosphorylation by competition with phosphate. In environments where phosphate concentrations are high, arsenate toxicity to biota is generally reduced. As arsenate is a phosphate analogue, organisms living in elevated arsenate environments must acquire the nutrient phosphorous yet avoid arsenic toxicity.

Arsenic compounds cause acute and chronic effects in individuals, populations and communities at concentrations ranging from a few micrograms to milligrams per litre, depending on species, time of exposure and end-points measured. These effects include lethality, inhibition of growth, photosynthesis and reproduction, and behavioural effects. Arsenic-contaminated environments are characterized by limited species abundance and diversity. If levels of arsenate are high enough, only species which exhibit resistance may be present.

2.1 Identity
Elemental arsenic (As) is a member of Group 15 of the periodic table, with nitrogen, phosphorus, antimony and bismuth. It has an atomic number of 33 and an atomic mass of 74.91. The Chemical Abstract Service (CAS), National Institute for Occupational Safety and Health Registry of Toxic Effects of Chemicals (RTECS), Hazardous Substances Data Bank (HSDB), European Commission, and UN transport class numbers are 7440-38-2, HSB 509, CG 05235 000, 033-001-00-X and UN 1558, respectively.

This monograph deals with arsenic and inorganic and organic arsenic compounds, except arsine (AsH3), for which a Concise International Chemical Assessment Document (CICAD) is being prepared.

2.2 Chemical and physical properties of arsenic compounds
Arsenic is a metalloid widely distributed in the earth’s crust. It can exist in four valency states; –3, 0, +3, and +5. In strongly reducing environments, elemental arsenic and arsine (–3) can exist. Under moderately reducing conditions, arsenite (+3) may be the dominant form, but arsenate (+5) is generally the stable oxidation state in oxygenated environments.

Arsenic and its compounds occur in crystalline, powder, amorphous or vitreous forms. They usually occur in trace quantities in all rock, soil, water and air. However, concentrations may be higher in certain areas as a result of weathering and anthropogenic activities including metal mining and smelting, fossil fuel combustion and pesticide use.

Arsenical salts exhibit a range of aqueous solubilities depending on the pH and the ionic environment.

There are many arsenic compounds of environmental importance. Representative marine arsenic-containing compounds, of which some are found in terrestrial systems, are shown in Table 1; their molecular structures are shown Fig. 1. Other arsenic compounds discussed in the text are listed in Table 2.

Table 1. Naturally occurring inorganic and organic As species
(see Fig. 1 for structures [1]–[22])










methylarsonic acid

monomethylarsonic acid, MMA



dimethylarsinic acid

cacodylic acid, DMA



trimethylarsine oxide



tetramethylarsonium ion











[20], [21]

dimethylarsinoylribitol sulfate


Speciation determines how arsenic compounds interact with their environment. For example, the behaviour of arsenate and arsenite in soil differs considerably. Movement in environmental matrices is a strong function of speciation and soil type. In a non-absorbing sandy loam, arsenite is 5–8 times more mobile than arsenate (Gulens et al., 1979). Soil pH also influences arsenic mobility. At a pH of 5.8 arsenate is slightly more mobile than arsenite, but when pH changes from acidic to neutral to basic, arsenite increasingly tends to become the more mobile species, though mobility of both arsenite and arsenate increases with increasing pH (Gulens et al., 1979). In strongly adsorbing soils, transport rate and speciation are influenced by organic carbon content and microbial population. Both arsenite and arsenate are transported at a slower rate in strongly adsorbing soils than in sandy soils.

Figure 1

Table 2. Other As compounds of environmental significance referred to in the text





Inorganic As, trivalent


As(III) oxide

As trioxide, arsenous oxide, white As

As2O3 (or As4O6)


arsenenous acid

arsenious acid



As(III) chloride

As trichloride, arsenous trichloride



As(III) sulfide

As trisulfide orpiment, auripigment


Inorganic As, pentavalent


As(V) oxide

As pentoxide



arsenic acid

ortho-arsenic acid



arsenenic acid

meta-arsenic acid


arsenates, salts of ortho-arsenic acid

H2AsO4–, HAsO42–, AsO43–

Organic As











(4-aminophenyl)-arsonic acid

arsanilic acid, p-aminobenzene-arsonic acid

Chemical structure


4,4-arsenobis(2-aminophenol) dihydrochloride

arsphenamine, salvarsan

Chemical structure


[4-[aminocarbonyl-amino]phenyl] arsonic acid

carbarsone, N-carbamoylarsanilic acid

Chemical structure


[4-[2-amino-2-oxoethyl)amino]-phenyl] arsonic acid


Chemical structure


3-nitro-4-hydroxy-phenylarsonic acid

Chemical structure


4-nitrophenylarsonic acid

p-nitrophenylarsonic acid

Chemical structure





Under oxidizing and aerated conditions, the predominant form of arsenic in water and soil is arsenate. Under reducing and waterlogged conditions (< 200 mV), arsenites should be the predominant arsenic compounds. The rate of conversion is dependent on the Eh and pH of the soil as well as on other physical, chemical and biological factors.

In brief, at moderate or high Eh, arsenic can be stabilized as a series of pentavalent (arsenate) oxyanions, H3AsO4, H2AsO4–, HAsO42– and AsO43–. However, under most reducing (acid and mildly alkaline) conditions, arsenite predominates. A pH and Eh diagram is shown in Fig. 2.

Figure 2

2.3 Analytical procedures
Historically, colorimetric and gravimetric methods have been used for the determination of arsenic. However, these methods are either semi-quantitative or lack sensitivity. In recent years, atomic absorption spectrometry (AAS) has become the method of choice, as it offers the possibility of selectivity and sensitivity in the detection of a wide range of metals and non-metals including arsenic. Popular methods for generating atoms for AAS are flame and electrothermally heated graphite furnaces. However, a commonly used technique for the measurement of arsenic is the highly sensitive hydride generation atomic absorption spectrometric method (HGAAS). However, although it is suitable for total arsenic determination after appropriate digestion the technique is only routinely used to speciate a limited number of compounds – arsenite, arsenate, MMA, DMA, trimethylarsine oxide (TMAO).

Hydride generation followed by cryogenic trapping and AAS detection is a relatively inexpensive technique for the speciation of inorganic arsenic and its methylated metabolites (Ng et al., 1998a), although more expensive hyphenated techniques may also be used.

A number of other approaches have been reported for speciation of arsenic. Inductively coupled plasma-mass spectrometry (ICP-MS) offers very high sensitivity for the determination of arsenic, and coupled with HPLC enables equally sensitive estimation of a wide variety of arsenic species.

2.4 Sample preparation and treatment
2.4.1 Sampling and collection
Care must be taken to avoid contamination and prevent speciation changes during sample collection and storage. Plastic containers should be acid washed and traces of oxidizing and reducing agents avoided to preserve the oxidation state of arsenic compounds. Freezing samples to –80 °C has also been recommended (Crecelius, 1986). Concentrated hydrochloric acid (1 ml to 100 ml urine) has been added to urine to prevent bacterial growth (Concha et al., 1998a).

For particulates in air and aerosols sampling, various types of filter have been employed including polytetrafluoroethylene (Rabano et al., 1989), cellulose ester (Yager et al., 1997), glass microfibre (Beceiro-Gonzalez et al.,1997) and filter paper (Tripathi et al., 1997).

2.4.2 Oxidative digestion
Acid digestion (George & Roscoe, 1951) and dry ashing (George et al., 1973) are the two basic methods which have been widely employed for oxidative digestion of samples before analysis. In more recent years, microwave-assisted digestion has been used (Le et al., 1994b; Thomas et al., 1997). For analysis of biological soft tissues by ICP techniques, a simple partial digestion in a closed vessel at low temperature and pressure is often sufficient for the sample preparation and pretreatment step.

2.4.3 Extraction
For speciation of arsenic, solvent extraction is often required before analysis. For example, arsenite and arsenate in soil can be speciated after a hydrochloric acid and chloroform extraction procedure (Chappell et al., 1995; Ng et al., 1998b). Water has been used for the extraction of soluble arsenic compounds from soil with the aid of ultrasonic treatment (Hansen et al., 1992). Forms of arsenic compounds can also be separated by sequential extractions based on procedures described by Tessier et al. (1979). Aqueous methanol has been widely used for the extraction of organic arsenic species (Edmonds & Francesconi, 1981a; Shiomi et al., 1988a; Shibata et al., 1996; Kuehnelt et al., 1997). Yu & Wai (1991) and Laintz et al. (1992) described the use of sodium bis(trifluoroethyl) dithiocarbamate (NaFDDC) as a selective chelation reagent of arsenic followed by either a gas chromatograph (GC) detection or supercritical fluid chromatography (SFC) detection. The former gave a limit of detection of 10 µg As/litre in water and the latter gave similar sensitivity after 100–1000-fold preconcentration of the chelate complex in organic solvent.

2.4.4 Supercritical fluid extraction
There are very few publications on the use of supercritical fluid extraction (SFE) for the determination of arsenic. Wenclawiak & Krah (1995) reported a procedure for the measurement of arsenic species using SFE followed by GC or SFC detection. The authors described a rapid extraction of organic and inorganic arsenic species from spiked sand and soil samples by SFE with on-line derivatization using thioglycollic acid methylester (TGM) under supercritical conditions. The TGM derivatives are thermally stable, which makes them amenable to GC–SFC determination. The extracts were chromatographed without further clean-up steps. The limits of detection were 1 ng As/µl and 3 ng As/µl injection for DMA-TGM and MMA-TGM respectively.

2.5 Macro-measurement
Most procedures for the separation and determination of arsenic are based on distillation and hydrogen sulfide precipitation methods. Beard & Lyerly (1961) reported a gravimetric method for the measurement of arsenic following extraction of arsenic as AsCl3 by benzene in strong hydrochloric acid. The recovery was close to 100% when 20 mg was spiked into an aqueous solution.
Vogel (1954) described the historic Marsh test, a qualitative method based on the generation of arsine (AsH3) by the addition of Zn granules to sulfuric acid. If the gas is mixed with hydrogen, and conducted through a heated glass tube, it decomposes into hydrogen and metallic arsenic which is deposited as a brownish-black “mirror” just beyond the heated part of the tube.

2.6 Colorimetric methods
George & Roscoe (1951) reported a spectroscopic emission measurement of the blue complex formed by the reaction of ammonium molybdate and hydrazine sulfate with arsenic in various biological materials. The sensitivity was about 0.01 µg.

George et al. (1973) carried out a collaborative study for a colorimetric measurement of arsenic in poultry and swine tissues using silver diethyldithiocarbamate (AgDDTC) as the complexing agent. The sensitivity was 0.1 mg/kg in tissues. Dhar et al. (1997) reported a detection limit of 0.04 mg/litre with 95% confidence limit using AgDDTC in chloroform with hexamethylenetetramine.

Gutzeit’s test (Vogel, 1954) is based on the generation of arsine from arsenic compounds by the addition of zinc granules to concentrated sulfuric acid. The arsine can be detected by means of a strip of filter paper moistened with silver nitrate or mercuric chloride. The arsine reacts with silver nitrate to give a grey spot, and with mercuric chloride to give a yellow to reddish-brown spot. The sensitivity is about 1 µg. A modification of this method, using mercuric bromide, is found in a test kit currently being used in Bangladesh for groundwater testing which has a limit of detection of 50–100 µg/litre under field conditions.

2.7 Methods for total inorganic arsenic
Methods for the analysis of inorganic arsenic based on its conversion to arsenic trichloride or arsenic tribromide by treatment with 6 mol/litre hydrochloric acid or hydrobromic acid have been described. The arsenic trihalide is separated from the remaining organic arsenic either by distillation (Maher, 1983) or by solvent extraction (Brooke & Evans, 1981). The methods have been applied routinely to the measurement of inorganic arsenic in a variety of foodstuffs, including those of marine origin where any inorganic arsenic is a small percentage of the total arsenic present (Flanjak, 1982; Shinagawa et al., 1983).

2.8 Atomic spectrometry
Common flame atomic absorption spectrometric methods are flame AAS (FAAS), electrothermal AAS (ETAAS) and hydride generation AAS (HGAAS). FAAS is relatively less sensitive for the determination of arsenic than ETAAS and HGAAS. Its detection limit is usually in the range of sub-milligram quantities per litre, and therefore it has limited application, especially for biological samples.

ETAAS, referred to also as graphite furnace-AAS (GFAAS), is generally one of the most sensitive atomic spectroscopic methods. Julshamn et al. (1996) reported factors that are known to interfere with the GFAAS determination of arsenic. The study was carried out by four participating laboratories using five marine standard reference materials. A mixture of palladium and magnesium salts has been recommended as a chemical modifier to avoid nickel contamination of the graphite furnace. The use of a pyrolytically coated graphite furnace tube with the L’vov platform improves sensitivity. Larsen (1991) achieved characteristic masses of about 16 pg of arsenic for arsenate, monomethylarsonate, DMA, arsenobetaine, arsenocholine and tetramethylarsonium ion calculated from aqueous standard solutions.

HGAAS is probably the most widely used method for the determination of arsenic in various matrices. Most of the reported errors in the determination of arsenic by HGAAS with NaBH4 can be attributed to variation in the production of the hydride and its transport into the atomizer. The reaction and atomization of arsine have been reviewed and discussed by Welz et al. (1990). The addition of a solution of l-cysteine to a sample before hydride generation eliminates interference by a number of transition metals in the generation of arsine from arsenite and arsenate (Boampong et al., 1988), and improves responses of arsine generated from MMA and DMA in the presence of arsenite and arsenate (Le et al., 1994a).

Holak & Specchio (1991) described the determination of total arsenic, arsenite and arsenate in foods by HGAAS after a chloroform extraction procedure. The recovery was > 80%. Similar methods (Chappell et al., 1995; Ng et al., 1998a) have been developed for arsenic speciation in soils. Ybanez et al. (1992) described a HGAAS determination of arsenic in dry ashed mussel products and reported a detection limit of 0.017 µg As/g with a precision of 3%.

HGAAS has been used for arsenic speciation of inorganic arsenic and its urinary metabolites, MMA and DMA, since 1973, when Braman & Foreback (1973) introduced a cold-trapping step into a basic hydride generation system. Since then a number of improvements have been made to this method (Crecelius, 1978; Buchet & Lauwerys, 1981; Van Cleuvenbergen et al., 1988). Ng et al. (1998b) described an optimized procedure for the speciation of arsenic metabolites in the urine of occupationally exposed workers and experimental animals with detection limits of 1, 1.3 and 3 ng per reaction of inorganic arsenic, MMA and DMA (equivalent to 0.25 µg/litre, 0.325 µg/litre, and 0.75 µg/litre respectively), using 4 ml of urine per reaction.

HGAAS has also been widely employed for analysis of arsenic in water (Chen et al., 1994; Chatterjee et al., 1995; Mandal et al., 1996; Dhar et al., 1997; Biswas et al., 1998). Hasegawa et al. (1994) published the first report of trivalent methyl arsenicals, namely monomethylarsonous acid [MMA(III)] and dimethylarsinous acid [DMA(III)], being found and measured in natural waters. Arsenious acid, MMA(III) and DMA(III) were separated from the pentavalent species by solvent extraction using diethylammonium diethyldithiocarbamate (DDDC) and determined by HGAAS after cold trapping and chromatographic separations. The detection limits were 13–17 pmol/litre and 110–180 pmol/litre for the trivalent and pentavalent species respectively.

Atomic fluorescence spectrometry (AFS) has recently been used for the detection of arsenic hydrides in the ultraviolet spectral region because of the small background emission produced by the relatively cool hydrogen diffusion flame (Gomez-Ariza et al., 1998). The use of cold vapour or hydride generation, together with intense light sources, allows very low detection limits to be achieved. For example, arsenic species in seawater have been measured using hydride generation and cold trapping, coupled with AFS detection at 193.7 nm (Featherstone et al., 1998). They found detection limits of 2.3, 0.9, 2.4 and 3.7 ng/litre for arsenite, arsenate, MMA and DMA respectively (in a 5 ml sample), with a precision of 3.5%.

2.9 ICP methodologies
The main advantages of ICP-MS over ICP-AES are lower detection limits (sub-nanogram to sub-picogram) with wide linear range and isotope analysis capability of high precision. The detection limits of ICP-AES are typically in the range of sub-micrograms to sub-nanograms.

ICP-MS is more susceptible to isobaric interferences arising from the plasma. For example, hydrochloric acid and perchloric acid are not desirable for sample preparation, because the chloride ions generated in the plasma combine with the argon gas to form argon chloride (ArCl). This has the same mass as arsenic (75) which could lead to error if not corrected. Therefore, whenever possible, only nitric acid should be used in sample preparation. Careful sample preparation is as important as the final measurement, and special care should be taken to avoid contamination and losses by volatilization, adsorption and precipitation.

2.10 Voltammetry
Voltammetric stripping methods are mostly based on the chemical reduction of As(V) to As(III) before the deposition step, because it has been generally assumed that As(V) is electrochemically inactive. Mercury and gold (or gold-plated) electrodes are most commonly used for the determination of arsenic.

Sadana (1983) used differential pulse cathodic stripping voltammetry (DPCSV) coupled to a hanging mercury drop electrode (HMDE) to determine arsenic in drinking-water in the presence of Cu2+ and reported a detection limit of 1 ng/ml and a relative standard deviation of 6.4%. Zima & van den Berg (1994) reported a detection limit of 3 nmol/litre in seawater. DPCSV was employed by Higham & Tomkins (1993) to determine arsenic in canned tuna fish. They evaluated a number of digestion procedures and found the best procedure gave 93–96% recovery. No detection limit was reported.

A gold electrode affords better sensitivity than a mercury electrode. Hua et al. (1987) reported an automated determination of total arsenic in seawater by flow constant-current stripping analysis with a gold film fibre electrode, in which As(V) in the sample was reduced to As(III) with potassium iodide; the detection limit was 0.15 µg/litre. The reduction of As(V) to As(III) can also be achieved by reaction with sulfur dioxide or hydrazinium chloride for use with a gold electrode or HMDE respectively (Esteban et al., 1994).

Huiliang et al. (1988) have shown that As(V) can be reduced to elemental arsenic provided that extremely low reduction potentials are used. They used this method to measure As(V) and total arsenic in seawater and urine. The detection limit was 0.1 µg/litre using constant-current stripping voltammetry on a gold-coated platinum-fibre electrode. Greulach & Henze (1995) developed a cathodic stripping voltammetric method for the determination of As(V) in water and stream sediment on the basis that As(V) can be reduced in perchloric acid solution containing d-mannitol, combined with the accumulation of arsenic by co-precipitation with copper on an HMDE. The detection limit was 4.4 µg/litre.

Pretty et al. (1993) developed an on-line anodic stripping voltammetry (ASV) flow cell coupled to ICP-MS for the determination of arsenic in spiked urine. The detection limit was 130 pg/ml and the recovery was 94–113%.

2.11 Radiochemical methods
Orvini et al. (1974) reported a combustion technique for sample preparation and determination of arsenic, selenium, zinc, cadmium and mercury by neutron activation analysis (NAA) in environmental matrices including a range of standard reference materials. The recoveries were 98–100%. Sharif et al. (1993) described a NAA technique for the determination of arsenic in eight species of marine fishes caught in the bay of Bengal, Bangladesh.

Haddad & Zikovsky (1985) measured several elements including arsenic in air from workroom welding fumes by NAA and reported a detection limit of 0.17 ± 0.07 µg/m3. Landsberger & Wu (1995) reported the use of NAA to measure arsenic from environmental tobacco smoke in indoor air with a detection limit of 0.2 ng.

Chutke et al. (1994) described a radiochemical solvent extraction procedure for the determination of arsenite using an arsenic-76 tracer. The procedure is based on the complexation of arsenite with toluene-3,4-dithiol (TDT) at pH 2 and subsequent extraction in benzene. This isotopic dilution technique was employed to measure arsenic in a range of standard and certified reference materials. The detection limit was 250 ng with an accuracy of about 4% error and 170 ng with about 12% error.

2.12 X-ray spectroscopy
Particle-induced X-ray emission spectrometry (PIXES) is an analytical technique that entails the bombardment of a sample (target) with charged particles, resulting in the emission of characteristic X-rays of the elements present. PIXES is a multi-elemental technique with a detection limit of approximately 0.1 µg As/g. It has the advantage of using small samples (1 mg or less) and being a non-destructive technique. Applications of PIXES in the environmental field have mostly focused on atmospheric particulate material (aerosol samples) (Maenhaut, 1987).

Castilla et al. (1993) described the determination of arsenite and arsenate by X-ray fluorescence (XRF) spectroscopy in water with a detection limit of 3.1 ng/g. The recovery was 97 ± 2.1% and 103 ± 1.4% for arsenite and arsenate respectively. In this method, the water sample was acidified to pH 2 and arsenite co-precipitated with sodium dibenzyldithiocarbamate (DBDTC). Arsenate in the filtrate was then reduced to arsenite with potassium iodide before the co-precipitation step for the XRF measurement.

Although there are a variety of methods to determine the concentration and oxidation states of arsenic in coal and ash, there have been few attempts to determine the mineral forms of arsenic. Huffman et al. (1994) described the use of X-ray absorption fine structure (XAFS) spectroscopy and its capability of providing speciation information at realistic concentrations of 10–100 mg/kg. They identified arsenic present as arsenopyrite in one coal sample and as aluminosilicate slag and calcium orthoarsenate in combustion ashes.

2.13 Hyphenated techniques
Hyphenated techniques is a term referring to the coupling of more than two instrumental systems to form a single technique.

The combination of chromatographic separation with element-specific spectrometric detection has been proved to be particularly useful for the speciation of arsenic compounds at trace levels in environmental samples. Woller et al. (1995) used AFS detection in combination with ultrasonically nebulized liquid chromatography (LC) for on-line speciation of arsenic, but found that the technique had limited sensitivity owing to matrix interferences. More recently, Slejkovec et al. (1998) used LC and purge-and-trap GC interfaced with AFS to separate and quantify six arsenic species with detection limits of 0.5 ng/ml As (100 µl). Gomez-Ariza et al. (1998) coupled anion-exchange HPLC, hydride generation and AFS to achieve detection limits of 0.17, 0.45, 0.30 and 0.38 µg/litre for arsenite, DMA, MMA and arsenate respectively (using a 20 µl loop). Arsenobetaine was also determined by introducing an on-line photo-oxidation step after the chromatographic separation.

Ebdon et al. (1988) described a number of coupled chromatograph–atomic spectrometry methods for arsenic speciation including GC or HPLC with detection by atomic spectrometry, namely FAAS, flame atomic fluorescence spectrometry (FAFS) and ICP-AES. The FAAS system is capable of detection at less than 1 µg/kg (0.22–0.55 ng absolute for different species) when levels permit; HPLC–hydride generation–FAAS is probably the simplest routine method and HPLC–hydride generation–ICP-AES is preferred for multi-elemental analysis. HPLC–ICP-AES has been employed for the speciation of organic arsenic of aquatic origin (Francesconi et al., 1985; Gailer & Irgolic, 1996). Gjerde et al. (1993) described the coupling of microbore columns with direct-injection nebulization to ICP-AES and reported a detection limit of 10 µg/litre (100 pg). Microbore HPLC has the advantage of analysing small sample size using low flow rates (80–100 µl/min) of mobile phases.

Numerous methods (Shum et al., 1992; Larsen et al., 1993; Magnuson et al., 1996; Thomas et al., 1997; Le & Ma, 1997) have been developed for the speciation of arsenic using the separation power of chromatography coupled to the sensitivity of ICP-MS detection. Heitkemper et al. (1989) described an anion-exchange HPLC–ICP-MS method for the speciation of arsenite, arsenate, MMA and DMA in urine with absolute detection limits ranging from 36 to 96 pg (corresponding to 0.7–1.9 µg/litre in a 50 µl injection). Beauchemin et al. (1989) reported detection limits for arsenic species in DORM-1 (a dogfish muscle certified reference material) ranging between 50 and 300 pg using ion pairing and ion exchange HPLC-ICP-MS. Anion exchange is more tolerant because of the higher buffering capacity of the mobile phase. Cation pairing is more suitable for the determination of DMAA and arsenobetaine in biological samples containing high concentrations of salts. Pergantis et al. (1997) analysed and speciated animal feed additives using microbore HPLC–ICP-MS with detection limits ranging from 0.1 to 0.26 pg. Hakala and Pyy (1992) described an ion-pairing HPLC-HGAAS method for speciation of arsenite, arsenate, MMA and DMA in urine with detection limits of 1.0, 1.6, 1.2 and 4.7 µg/litre respectively.

Ding et al. (1995) described the coupling of micellar liquid chromatography (MLC) and ICP-MS for the speciation of arsenite, arsenate, MMA and DMA with detection limits of 90 pg for DMA and 300 pg for the other species. MLC is a type of chromatography that uses surfactants in aqueous solutions, well above their critical micelle concentration, as alternative mobile phases for reversed-phase liquid chromatography (RPLC). MLC extends the analyte candidates to almost all hydrophobic and many hydrophilic compounds providing they can partition to the micelles. Other advantages of MLC over RPLC include simultaneous separation of both ionic and non-ionic compounds, faster analysis times and improved detection sensitivity and selectivity.

Capillary electrophoresis (CE) is a versatile technique for the separation of a variety of analytes ranging from small inorganic ions to large biomolecules such as proteins and nucleic acids. CE-ICP-MS has been described for the speciation of arsenic by Liu et al. (1995) with detection limits of 100 pg arsenite/ml and 20 pg arsenate/ml and Olesik et al. (1995) with a detection limit of 8 µg/litre (1 pg injection).

Although techniques such as HPLC–ICP-MS and MLC–ICP-MS offer the advantages of high sensitivity and selectivity as well as low detection limits, species identification is based on the comparison of chromatographic retention times to those of available standards. When structure information is required, as well as quantification, electrospray HPLC–MS (Siu et al., 1991) and ionspray MS (Corr, 1997) should be considered. Corr & Larsen (1996) reported the use of LC–MS–MS for speciation of arsenic with a detection limit of 2 pg for the tetramethylarsonium cation.

3.1 Natural sources
Arsenic is the main constituent of more than 200 mineral species, of which about 60% are arsenate, 20% sulfide and sulfosalts and the remaining 20% include arsenides, arsenites, oxides and elemental arsenic (Onishi, 1969). The most common of the arsenic minerals is arsenopyrite, FeAsS, and arsenic is found associated with many types of mineral deposits, especially those including sulfide mineralization (Boyle & Jonasson, 1973). The ability of arsenic to bind to sulfur ligands means that it tends to be found associated with sulfide-bearing mineral deposits, either as separate As minerals or as a trace of a minor constituent of the other sulfide minerals. This leads to elevated levels in soils in many mineralized areas where the concentrations of associated arsenic can range from a few milligrams to > 100 mg/kg.

Concentrations of various types of igneous rocks range from < 1 to 15 mg As/kg, with a mean value of 2 mg As/kg. Similar concentrations (< 1–20 mg As/kg) are found in sandstone and limestone. Significantly higher concentrations of up to 900 mg As/kg are found in argillaceous sedimentary rocks including shales, mudstone and slates. Up to 200 mg As/kg can be present in phosphate rocks (O’Neill, 1990).

Concentrations of arsenic in open ocean water are typically 1–2 µg/litre. The concentrations of arsenic in unpolluted surface water and groundwater are typically in the range of 1–10 µg/litre. Elevated concentrations in surface water and groundwater of up to 100–5000 µg/litre can be found in areas of sulfide mineralization (Welch et al., 1988; Fordyce et al., 1995). Elevated concentrations (> 1 mg As/litre) in groundwater of geochemical origins have also been found in Taiwan (Chen et al., 1994), West Bengal, India (Chatterjee et al., 1995; Das et al., 1995, 1996; Mandal et al., 1996) and more recently in most districts of Bangladesh (Dhar et al., 1997; Biswas et al., 1998). Elevated arsenic concentrations were also found in the drinking-water in Chile (Borgono et al., 1977); North Mexico (Cebrian et al., 1983); and several areas of Argentina (Astolfi et al., 1981; Nicolli et al., 1989; De Sastre et al., 1992). Arsenic-contaminated groundwater was also found in parts of PR China (Xinjiang and Inner Mongolia) and the USA (California, Utah, Nevada, Washington and Alaska) (Valentine, 1994). More recently, arsenic concentrations of < 0.98 mg/litre have been found in wells in south-western Finland (Kurttio et al., 1998). Levels as high as 35 mg As/litre and 25.7 mg As/litre have been reported in areas associated with hydrothermal activity (Kipling, 1977; Tanaka, 1990).

In nature, arsenic-bearing minerals undergo oxidation and release arsenic to water. This could be one explanation for the problems of arsenic in the groundwater of West Bengal and Bangladesh. In these areas the groundwater usage is very high. It has been estimated that there are about 4–10 million tube wells in Bangladesh alone. The excessive withdrawal and lowering of the water table for rice irrigation and other requirements lead to the exposure and subsequent oxidation of arsenic-containing pyrite in the sediment. As the water table recharges after rainfall, arsenic leaches out of the sediment into the aquifer.

However, recent studies seem to favour the reduction of Fe/As oxyhydroxides as the source for arsenic contamination in groundwater (Nickson et al., 1998; BGS, 2000; BGS & DPHE, 2001). Arsenic forms co-precipitates with ferric oxyhydroxide. Burial of the sediment, rich in ferric oxyhydroxide and organic matter, has led to the strongly reducing groundwater conditions. The process has been aided by the high water table and fine-grained surface layers which impede the penetration of air to the aquifer. Microbial oxidation of organic carbon has depleted the dissolved oxygen in the groundwater. The highly reducing nature of the groundwater explains the presence of arsenite (< 50%) in the water. The “pyrite oxidation” hypothesis is therefore unlikely to be a major process, and the “oxyhydroxide reduction” hypothesis (Nickson et al., 1998; Acharyya et al., 1999) is probably the main cause of arsenic contamination in groundwater. Although the oxyhydroxide reduction hypothesis requires further validation, there is no doubt that the source of arsenic in West Bengal and Bangladesh is geological, as none of the explanations for anthropogenic contamination can account for the regional extent of groundwater contamination. During the past 30 years the use of phosphate fertilizers has increased threefold in this region. The widespread withdrawal of groundwater may have mobilized phosphate derived from fertilizers and from the decay of natural organic materials in shallow aquifers. The increase in phosphate concentration could have promoted the growth of sediment biota and the desorption of arsenic from sediments, and the combined microbiological and chemical process might have increased the mobility of arsenic (Acharyya et al., 1999).

Marine organisms naturally accumulate considerable quantities of organic arsenic compounds. In marine animals the bulk of this arsenic is present as arsenobetaine, whereas marine algae contain most of the arsenic as dimethylarsinoylribosides. Humans are therefore exposed to these arsenic compounds through any diet that includes seafoods. This subject is fully discussed in Chapter 4.

Some arsenic compounds are relatively volatile and consequently contribute significant fluxes in the atmosphere. It has been estimated that the atmospheric flux of As is about 73 540 tonnes/year of which 60% is of natural origin and the rest is derived from anthropogenic sources (Chilvers & Peterson, 1987). Volcanic action is the next most important natural source of arsenic after low-temperature volatilization, and on a local scale it will be the dominant atmospheric source.

3.2 Sources of environmental pollution
3.2.1 Industry
It has long been recognized that the smelting of non-ferrous metals and the production of energy from fossil fuel are the two major industrial processes that lead to anthropogenic arsenic contamination of air, water and soil. Other sources of contamination are the manufacture and use of arsenical pesticides and wood preservatives.

Smelting activities generate the largest single anthropogenic input into the atmosphere (Chilvers & Peterson, 1987).

Tailings from metal-mining operations are a significant source of contamination, and can lead to contamination of the surrounding topsoils, and, because of leaching, sometimes the groundwater too. It has been estimated that several billion tons of tailings waste exist in the USA alone (Wewerka et al., 1978). As sulfur is often present in these tailings, exposure to the atmosphere in the presence of water leads to the production of an acid solution that can leach many elements including arsenic.

Elevated concentrations of arsenic in acid sulfate soils in Canada and New Zealand are associated with pyrite (Dudas, 1987). Concentrations of arsenic < 0.5% through lattice substitution of sulfur in this pyrite iron-rich bauxite have been recorded.

In the United Kingdom, the estimated arsenic releases (Hutton & Symon, 1986) were 650 tonnes/year from the non-ferrous metal industry, 9 tonnes/year emission into the atmosphere and 179 tonnes/year to landfill from iron and steel production, and 297 tonnes/year into the atmosphere and 838 tonnes/year to landfill from fossil fuel combustion. In 1996, the estimated total releases of arsenic to the air in the UK were 50 tonnes (DG Environment, 2000).

The working group of the European Union DGV (the directorate with responsibility for the environment) concluded that there were large reductions in the emissions of arsenic to air in several member countries of the European Union in the 1980s and early 1990s. In 1990, the total emissions of arsenic to the air in the member states were estimated to be 575 tonnes, of which 492 tonnes came from stationary combustion (mainly coal and oil combustion) and 77 tonnes from production processes, mainly from the iron and steel industry (35 tonnes) and the non-ferrous metal industry (31 tonnes) (DG Environment, 2000).

Arsenic is present in the rock phosphate used to manufacture fertilizers and detergents. In 1982, the United Kingdom imported 1324 × 103 tonnes of rock phosphate with an estimated arsenic burden of 10.2 tonnes (Hutton & Symon, 1986).

3.2.2 Past agricultural use
In 1983, arsenical pesticides were one of the largest classes of biocontrol agent in the USA (Woolson, 1983). From the 1960s there was a shift, in herbicide use, from inorganic compounds (including lead and calcium arsenate and copper acetoarsenite) to inorganic and organic compounds (arsenic acid, sodium arsenate, MMA and DMA). Use of total arsenical pesticides, excluding wood preservatives, at the time of publication (1983) was estimated at 7–11 × 103 tonnes As/year. Annual historical applications of lead arsenate to orchards in the USA ranged from 32 to 700 kg As/ha. Residues in orchard soils as high as 2500 mg/kg have been reported, but they are more commonly in the range of 100–200 mg/kg. In Australia between 1900 and 1950 As2O3 was widely used for controlling cattle ticks (Boophilus microplus), resulting in widespread arsenic contamination (Seddon, 1951).

3.2.3 Sewage sludge
The levels of arsenic in sewage sludge reflect the extent of industrialization of the area served by the local sewage system. Significant quantities may be added by arsenic-contaminated wastewater runoff derived from sources including atmospherically deposited arsenic, residues from pesticide usage, phosphate detergents and industrial effluent, particularly from the metal-processing industry. Levels of 0–188 mg As/kg dry weight have been reported in the United Kingdom (Woolson, 1983). Zhu & Tabatabai (1995) reported levels of 2.4–39.6 mg As/kg with a mean of 9.8 for sewage sludges from waste treatment plants in Iowa, USA.

O’Neil (1990) estimated that in the UK as a whole about 2.5 tonnes As/year is added to the agricultural land by use of sludge, compared to 6.1 tonnes As/year when phosphate fertilizer is used.

3.3 Uses
Arsenic is produced commercially by reduction of As2O3 with charcoal. As2O3 is produced as a by-product of metal-smelting operations. It is present in flue dust from the roasting of ores, especially those produced in copper smelting. In the 1960s, the pattern of use for As2O3 in the USA is believed to have been 77% as pesticides, 18% as glass, 4% as industrial chemicals and 1% as medicine. However, the pattern has changed over the years as the use of arsenic compounds for timber treatment has been increasingly popular since the late 1980s. Worldwide usage in the early 1980s was estimated to be 16 000 tonnes As/year as a herbicide, 12 000 tonnes As/year as a cotton desiccant/defoliant and 16 000 tonnes As/year in wood preservative (Chilvers & Peterson, 1987). By 1990, the estimated end-use of arsenic in the USA was 70% in wood preservatives, 22% in agricultural chemicals, 4% in glass, 2% in non-ferrous alloys and 2% in other uses including semiconductors (US DOI, 1991). Arsenic pentoxide and As2O3 are used as additives in alloys, particularly with lead and copper; arsenic and As2O3 are used in the manufacturing of low-melting glasses. High-purity arsenic metal and gallium arsenide are used in semiconductor products. Fowler’s solution (1% potassium arsenite solution) was used as a medication (Cuzick et al., 1992). As2O3 has been used for the treatment of acute promyelocytic leukaemia (Soignet et al., 1998).

Hutton & Symon (1986) reported that about 5000 tonnes/year As2O3 is imported to the United Kingdom for conversion to other arsenic compounds. These processes result in an estimated discharge of 87 tonnes As/year in manufacturing sludges on landfilled sites. Currently about 500 tonnes As/year is utilized in copper chrome arsenate (CCA) timber treatment, of which at most 5 tonnes/year is retained in sludges. Small amounts of arsenic are used in the production of glass, and most of the remainder is re-exported.

4.1 Transport and distribution between media
4.1.1 Air
Arsenic is primarily emitted into the atmosphere by high-temperature processes such as coal-fired power generation, smelting, burning vegetation and vulcanism. Natural low-temperature biomethylation and microbial reduction also release arsenic into the atmosphere; microorganisms can form volatile methylated derivatives of arsenic under both aerobic and anaerobic conditions, and can reduce arsenic compounds to release arsine gas (Cheng & Focht, 1979; Tamaki & Frankenberger, 1992) (see section 4.2.2). Arsenic is released into the atmosphere primarily as As2O3 or, less frequently, as one of several volatile organic compounds. Arsenic released to air exists mainly in the form of particulate matter (Coles et al., 1979). These particles are dispersed by the wind to a varying extent, depending on their size, and the particles are returned to the earth by wet or dry deposition. Arsines that are released from microbial sources in soils or sediments undergo oxidation in the air, reconverting the arsenic to less volatile forms that settle back to the ground (Wood, 1974; Parris & Brinckman, 1976).

Pacyna et al. (1989) studied atmospheric transport of arsenic from various sources in Europe to selected receptor sites in Norway. By modelling long-range transport they were able to calculate a dry deposition velocity for arsenic of 0.4 cm/second. Scudlark & Church (1988) measured arsenic in acid precipitation on the mid-Atlantic coast of the USA during 1985 and 1986. They calculated the total annual arsenic deposition rate to range from 38 to 266 µg/m2, with dry deposition estimated to comprise 29–55% of the total. Davidson et al. (1985) calculated the annual dry deposition flux of arsenic to the Olympic National Park, Washington (USA) to range from 76.7 to 208 µg/m2. The average annual wet deposition of arsenic at Chesapeake bay (Maryland, USA) was found to be 49 µg/m2 (Scudlark et al., 1994).

Total atmospheric arsenic emissions from both natural and anthropogenic sources have been estimated to be 31× 109 g/year, and total atmospheric arsenic removal was estimated to be 30–50 × 109 g/year. The global tropospheric residence time of arsenic appears to be about 9 days (Walsh et al., 1979). Nakamura et al. (1990) estimated global atmospheric emissions into the atmosphere and deposition of arsenic. Total emissions were estimated at 36 × 109 g/year, with the major source of atmospheric arsenic being anthropogenic emissions; the major natural source of arsenic was volcanic activity. Emissions from anthropogenic sources were estimated at 24 × 109 g/year, representing 64% of total arsenic influxes. Depositions from the atmosphere to the land and the oceans were estimated at 24 × 109 g/year and 9 × 109 g/year respectively. Akeredolu et al. (1994) calculated the total annual transport of arsenic into the Arctic atmosphere at 285 t (285 × 106 g) on the basis of a chemical transport modelling approach previously used for sulfur.

Arsenic in the atmosphere exists primarily adsorbed to particulate matter and mostly to particles < 2 µm in diameter (Coles et al., 1979). Waslenchuk (1978) found that atmospheric arsenate at the continental shelf of the south-eastern USA is associated exclusively with the particulate fraction. Rabano et al. (1989) collected size-fractionated aerosol samples at an urban site during 1987. A greater proportion (75%) of the arsenic was observed in the fine particles (< 2.5 µm). The As(III)/As(V) ratio for both fine and coarse (> 2.5 µm) particles was approximately 1. Similarly, Waldman et al. (1991) reported that 65% of the arsenic in aerosol samples collected at an urban site (China) was associated with fine particles (< 2.5 µm). Kelley et al. (1995) monitored arsenic in aerosol collected from the Kola Peninsula (Russia). They found 68% of arsenic associated with fine particles (< 1 µm), 26% with coarse particles (1–10 µm) and 7% with large particles (> 10 µm). The atmospheric residence time of particulate-bound arsenic depends on particle size and meteorological conditions, but a typical value is about 9 days (US EPA, 1982).

4.1.2 Freshwater and sediment
The dissolved forms of arsenic in the water column include arsenate, arsenite, monomethylarsonic acid (MMA) and dimethylarsinic acid (DMA) (Braman & Foreback, 1973). Some As(III) and As(V) species can interchange oxidation states depending on Eh, pH and biological processes (Ferguson & Gavis, 1972). Some arsenic species have an affinity for clay mineral surfaces and organic matter, and this can affect their environmental behaviour. Methylation and demethylation reactions are also important transformations controlling the mobilization and subsequent distribution of arsenicals (Mok & Wai, 1994). Transport and partitioning of arsenic in water depends on the chemical form of the arsenic and on interactions with other materials present. Arsenic may be adsorbed from water on to clays, iron oxides, aluminium hydroxides, manganese compounds and organic material (Callahan et al., 1979; Welch et al., 1988). The distribution and transport of arsenic in sediment is a complex process that depends on water quality, native biota and sediment type. There is a potential for arsenic release when there is fluctuation in Eh, pH, soluble arsenic concentration and sediment organic content (Abdelghani et al., 1981).

Ferguson & Gavis (1972) proposed an arsenic cycle for a stratified lake. In the aerobic epilimnetic water, reduced forms of arsenic tend to be oxidized to arsenate, which co-precipitates with ferric oxyhydroxide. Turbulent dispersion and convection transports some of the arsenate across the thermocline to the oxygen-depleted hypolimnion, where reduction to HAsO2 and AsS2– takes place, depending on the sulfur concentration and the Eh. Co-precipitation, adsorption and epitaxial crystal growth cause arsenic to be removed to the sediments, where reduction of ferric iron, arsenate and arsenite result in either solubilization or stabilization as an insoluble sulfide or arsenic metal. Microbial reduction and methylation to arsine solubilize the arsenic (see section 4.2), and diffusion through the sediments or mixing by currents or burrowing organisms (see section 4.1.4) cause arsenic to re-enter the water column.

Aurilio et al. (1994) studied the speciation and fate of arsenic in three lakes of the Aberjona watershed (Massachusetts, USA). Speciation appeared to be controlled by reduction, methylation, and oxidation processes, and by adsorption to and desorption from particles. Biologically mediated reduction, at rates of 0.2–0.5% total arsenic/day, and methylation, at rates of 0.4–0.6% total arsenic/day, occurred in the mixed layers of these lakes. These processes are slow or even absent in the hypolimnion, however, allowing arsenate to accumulate in seasonally anoxic hypolymnetic waters. High micromolar concentrations of arsenic, predominantly arsenite, persisted in the saline, sulfidic monimolimnion of one lake.

Clement & Faust (1981) studied the release of arsenic from contaminated sediments. Anaerobic conditions led to aqueous levels of arsenic, principally as arsenite, about 10 times higher than concentrations reached with aerobic conditions. Under aerobic conditions arsenic in the overlying water comprised 70% arsenate and 20% organic arsenic. The authors found that adsorption–desorption equilibria and the amount of ‘available’ arsenic present in the sediment greatly influenced the soluble arsenic concentration found in the aqueous phase. In sediment under oxidized conditions arsenic solubility was low and 87% of the arsenic in solution was present as arsenate. On reduction, arsenite became the major arsenic species in solution and solubility increased (Masscheleyn et al., 1991b). Ahmann et al. (1997) identified rapid arsenic mobilization from aquatic sediments in upper Aberjona (Massachusetts, USA) sediment microcosms. The findings suggest that arsenic reduction by microorganisms may contribute to arsenic flux from anoxic sediments in this arsenic-contaminated watershed.

The predominant arsenic species in the water column of lakes is arsenate, as expected in oxidizing environments (Seyler & Martin, 1989). Arsenite is usually present and sometimes dominates in bottom water which contains high concentrations of Fe(II) and low oxygen. Peterson & Carpenter (1983) reported that the arsenate : arsenite concentration ratio was 15 : 1 in the oxic region of the water column and 1 : 12 in the anoxic zone. Seasonal trends reveal higher concentrations of arsenic in summer than in winter. The source of arsenic in the summer is most likely surface sediments that have become anoxic causing a release into the water column of arsenic adsorbed on iron and manganese oxides (Singh et al., 1988; Crecelius et al., 1994).

Pettine et al. (1992) found that arsenate was the predominant arsenic species in the river Po (Italy). The main factors affecting dissolved concentrations included flow and suspended matter concentration and biological activity. The ratio between oxidized and reduced species appears to be significantly influenced by iron and manganese oxides. Abdel-Moati (1990) monitored arsenic in the Nile delta lakes and found arsenate to be the dominant arsenic species (85–95%). Increased arsenite (14–33%) was found near local sewage discharge points. Dimethylarsenic was the dominant organic species, reaching 22% of the total dissolved arsenic.

A temporal study of arsenic speciation in Davis Creek Reservoir, a seasonally anoxic lake in northern California (USA), demonstrated that dimethylarsinic acid increased sufficiently to become the dominant form of dissolved arsenic within the surface photic zone during late summer and early autumn. Methylated forms decreased and arsenate increased when the lake ‘turned over’ in early December, suggesting a degradation of dimethylarsinic acid (Anderson & Bruland, 1991).

Aggett & O’Brien (1985) report that Lake Ohakuri (New Zealand) becomes stratified during the summer. During this period arsenic released from the sediment accumulates in the hypolimnion until turnover when it is mixed with epilimnetic water. It is estimated that this turnover effect causes a temporary increase in arsenic concentrations of 10–20%. Aggett & Roberts (1986) conclude that arsenate and phosphate are incorporated into Lake Ohakuri sediments by co-precipitation at the time of formation of the hydrous oxides rather than by adsorption on existing surfaces. Aggett & Kriegman (1988) show that in sediment cores from Lake Ohakuri over 90% of arsenic in interstitial waters was present as arsenite, an indication that reduction from arsenate, the predominant form adsorbed from the lake water, was taking place. When conditions at the sediment–water interface became anoxic, arsenite diffused across the interface into the hypolimnion.

Johnson & Thornton (1987) studied the seasonal variation of arsenic in the Carnon river, south-west England (UK). Approximately 85% of the arsenic was found to originate from mine waters. Arsenic is found to a large extent (~80%) in the particulate phase; the authors suggest that sorptive or co-precipitation processes are responsible for the regulation of dissolved concentrations of arsenic in these waters. These processes are largely independent of pH. Adsorption appears to be important in the removal of arsenic from solution, with 80% being removed on entering estuarine waters.

Both adsorption of arsenic on iron-rich oxides on the surface of the sediments and incorporation of arsenic into the sediments by co-precipitation with hydrous iron oxides are factors controlling mobilization of sediment arsenic. The major arsenic species leached was arsenate; release of arsenate was found to be pH dependent and related to the total iron and free iron oxides in the sediments (Mok & Wai, 1989). Arsenate and arsenite differ in adsorption characteristics, and this influences their mobilization and subsequent distribution during water–sediment interactions. The extent of adsorption and remobilization varies with the oxidation state of arsenic, the Eh and the pH of the water. The increase in mobility of arsenate under more reducing conditions is generally attributed to the reduction of Fe3+ to Fe(II), with subsequent release of arsenate, and reduction of arsenate to arsenite (Mok & Wai, 1994).

Brannon & Patrick (1987) found that arsenate added to sediment became associated with relatively immobile iron and aluminium compounds. Addition of arsenate to sediments before anaerobic incubation also resulted in accumulation of arsenite and organic arsenic in the interstitial water and exchangeable phases of anaerobic sediments. Seyler & Martin (1989) report that the presence of arsenic in the anoxic zone of a permanently stratified lake was due to adsorption on to iron and manganese.

Sorption of arsenate, MMA and DMA on anaerobic bottom sediments from the Menominee river, Wisconsin (USA) is described by Langmuir isotherms (Holm et al., 1979). Singh et al. (1988) found that the adsorption of arsenite from aqueous solution followed first-order adsorption expression obeying Langmuir’s model of adsorption. Similar findings were reported by Yadava et al. (1988) for adsorption of arsenite by china clay, with maximum adsorption at pH 8. Sediment adsorption of arsenate, monosodium methanearsonate (MSMA) and DMA was positively correlated with clay content (Wauchope & McDowell, 1984). Organic sediments adsorbed arsenic more strongly than sandy sediments (Faust et al., 1987b).

The extent of uptake and the rate of adsorption of arsenate decrease with an increase in temperature from 20 °C to 40 °C. The amount of arsenate adsorbed increases as the pH of the system increases and reaches its maximum at pH 4.2 for haematite and pH 6.2 for feldspar. The removal of arsenate from aqueous solution by adsorption on to geological materials such as haematite and feldspar follows first-order kinetics, and intraparticle diffusion seems to control the mass transfer (Prasad, 1994). The adsorption of arsenate on alumina, haematite, kaolin and quartz was influenced by the charge of the solid surface and the arsenic speciation in solution as determined by pH (Xu et al., 1988).

Adsorption of arsenate by fly ash was significantly greater at pH 4 than at pH 7 or 10 and was found to be almost irreversible. Adsorption fitted both the Freundlich and Langmuir adsorption models (Sen & De, 1987; Diamadopoulos et al., 1993). The partitioning of arsenic between acidic fly ash and leachate is controlled by sorption on iron oxyhydroxide. The leaching of arsenic is mainly controlled by sorption on hydroxylamine-extractable (“amorphous”) iron oxyhydroxide; crystalline iron oxides appear to have little influence on the process (Van der Hoek & Comans, 1996). Thanabalasingam & Pickering (1986) found that arsenic sorption by humic acid varies with pH, adsorbate concentration and ash content of the substrate. At fixed pH, the amount of arsenic sorbed conformed to a Langmuir relationship, with calculated capacities in the region of maximum uptake (~pH 5.5) being of the order of 5250–6750 mg/kg for arsenite and 6750–8250 mg/kg for arsenate.

Gupta & Chen (1978) report that arsenic acid and arsenious acid species are effectively adsorbed in the pH range 4–7. Laboratory adsorption experiments indicated that arsenite is less effectively removed than arsenate. Adsorption of MMA and DMA on ferric oxyhydroxide and activated alumina decreased with increasing pH (4–11) (Cox & Ghosh, 1994).

Arsenic in porewater is controlled by the solubility of iron and manganese oxyhydroxides in the oxidized zone and metal sulfides in the reduced zone. Diagenetic sulfides are important sinks for arsenic in reduced, sulfidic sediments. During reduction, oxyhydroxides of iron and manganese dissolve, arsenic sulfides precipitate and arsenic is released to groundwater dominantly as arsenite (Moore et al., 1988). Therefore, mobilization of arsenic is more likely to occur in sediments low in iron and manganese oxyhydroxides, and calcium carbonate (Brannon & Patrick, 1987; Mok & Wai, 1990).

Bright et al. (1994) found that arsenite was the predominant arsenical in sediment porewater throughout a watershed receiving gold-mine effluent; dissolved arsenic in water column samples was mostly arsenate. Arsenic distribution in surficial sediments was controlled partially by the bulk movement of sediments, followed by burial with less-contaminated sediments in the upper reaches of the watershed. Particulate concentrations of arsenic contributed significantly (< 70%) to the total arsenic concentrations in the water column downstream of the gold-mine discharge. Azcue et al. (1994b) found that 66–83% of the arsenic in sediment porewater from a mine-polluted lake was arsenite. The concentration gradient of total dissolved arsenic indicated an upward diffusion of arsenic towards the water column, with the estimated annual fluxes being 0.8–3.8 µg/cm2.

Mok & Wai (1990) reported that acid precipitation caused increased release of both arsenate and arsenite from contaminated sediments (pH 2.7). Arsenic release decreased with increasing pH; lowest levels of release were found at pH 8.3 for arsenite and pH 6.3 for arsenate. Release of arsenic increased at more alkaline pH values. Xu et al. (1991) concluded that environmental acidification would increase the leaching of arsenic from sediments to surface waters under reducing conditions as a result of the release of arsenite from iron oxyhydroxide phases, but could also reduce the mobility because of enhanced adsorption under oxidizing conditions. However, a large reduction in pH (to £ 4) would enhance the mobility of arsenic even under oxidizing conditions.

4.1.3 Estuarine and marine water and sediment
An arsenic cycle has also been outlined for the estuarine environment. Sanders (1980) found that the major inputs to the marine environment were river runoff and atmospheric deposition. Biological uptake caused changes in arsenic speciation resulting in measurable concentrations of reduced and methylated arsenic species. The overall cycle is similar to the phosphate cycle, but the regeneration time for arsenic is much slower. Arsenic flows into the estuary as arsenate and arsenite from river water and mine adits. There is oxidation of arsenite to arsenate, microbiological reduction of arsenate to arsenite and removal of arsenic by dilution with seawater and subsequent transport out of the estuary. Inorganic arsenic can be adsorbed on to charged particles of iron oxyhydroxides and manganese oxides and deposited as flocculated particles to sediment. There is subsequent release of dissolved arsenite and arsenate following the reduction and dissolution of the iron and manganese carrier phases in the anoxic sediments. Arsenate can be reduced, either microbially or chemically, to arsenite within the anoxic sediment, and arsenic (as arsenate or arsenite) can enter by sediment resuspension (Sanders, 1980; Knox et al., 1984). Studies on the pH dependence of arsenate and arsenite adsorption to soils and sediments and to minerals are not consistent. For example, greater adsorption of arsenate to fly ash occurred at pH4 than at pHs 7 and 10 (Sen & De, 1987; Diamadopoulos et al., 1993), whereas Mok & Wai (1990) reported that absorption increased as pH increased for sediments.

Arsenic entering unpolluted estuaries associated with particulates remains adsorbed, and accumulates in sediment. Remobilization has only a small effect (< 7%) on the dissolved arsenic concentration in the water column. Dissolved arsenic species form complexes with low-molecular-weight dissolved organic matter, and these tend to prevent adsorption and co-precipitation interactions between arsenic and flocculating iron oxyhydroxides and humics (Waslenchuk & Windom, 1978). Langston (1983) reports that more than 80% of arsenic entering Restronguet creek in southwest England (United Kingdom) was retained by the sediment, which consequently acts as a sink for riverine inputs and limits transport of dissolved species to coastal waters. Iron oxyhydroxide scavenging seems to be a predominant factor in the removal of arsenic from the Scheldt estuary (The Netherlands) (van der Sloot et al., 1985). Millward et al. (1997) estimated an annual arsenic budget for the Thames plume (United Kingdom) and found that cycling of arsenic by phytoplankton was the dominant process. Inorganic arsenic was removed from the water column by phytoplankton and recycled during phytoplankton degradation and consumption.

An arsenic budget for Puget sound (Washington, USA) revealed that sediments accumulate less than 30% of the arsenic entering the sound (Crecelius et al., 1975). Carpenter et al. (1978) found that sedimentation processes including adsorption–desorption reactions with natural Puget sound suspended matter remove less than 15% of the dissolved arsenic input, with iron oxyhydroxides dominating what removal does occur. Most of the arsenic entering the sound is removed by advection of surface waters out into the strait of Juan de Fuca . A similar budget for Lake Washington (USA) showed equal inputs of arsenic from the atmosphere and from rivers, and subsequent removal by outflowing water (45%) and by accumulation in the sediments (55%) (Crecelius, 1975).

Riedel (1993) studied the distribution of dissolved and solid arsenic species in contaminated estuarine sediment. Arsenite was the dominant dissolved and solid species in the deeper reduced sediment, and arsenate was dominant in the oxidized surface layer. Arsenite in the interstitial water diffused toward the surface layer, where it was mostly oxidized to arsenate.

Howard et al. (1988) found that the distribution of dissolved inorganic arsenic in an estuary appears to be determined by a combination of secondary inputs arising from old mine drainage and advective transport of arsenic-enriched sediment interstitial waters into the water column. Bioutilization of the element during the warmer months results in the release of dissolved monomethylarsenic and dimethylarsenic. Inorganic arsenite and methylated arsenic species can account for up to 41% and 70% of the dissolved arsenic respectively, but only when the water temperature exceeds 12 °C (Howard et al., 1984). Arsenate was the dominant form in a temperate estuary throughout the year except late winter when a dimethylarsenic species was dominant (Riedel, 1993).

Andreae (1978, 1979) monitored seawater samples from the northeast Pacific and southern Californian coast (USA). Methylation and reduction of arsenate to arsenite and methylarsenic acids occur in the photic zone. Arsenic is taken up by planktonic organisms in the surface waters and transported to deeper waters with biogenic debris. At intermediate levels regeneration of arsenate occurs. There was a good correlation between photosynthetic activity and concentration of methylated arsenicals. Andreae & Froelich (1984) and Sadiq (1990) found that arsenate is more abundant in oxic seawaters whereas arsenite is more abundant in anoxic seawaters.

Waslenchuk (1978) found that concentrations of arsenic species in continental shelf waters of the south-eastern USA are controlled mainly by simple mixing of shelf waters and Gulf Stream intrusions. Riverine and atmospheric arsenic inputs to the shelf waters were relatively insignificant, and uptake of arsenic by biota had only a minor effect on arsenic distribution.

Byrd (1988) studied the seasonal cycle of arsenic on the continental shelf of the South Atlantic. During periods of high winds in the winter and early spring, inorganic arsenic concentrations are reduced to as little as 20% of typical open-ocean concentrations by sorption on to suspended sediments or incorporation into phytoplankton. In the late summer and early autumn arsenic is remobilized and returned to the water column, elevating arsenic concentrations to 50% more than open-ocean concentrations. Belzile (1988) analysed vertical profiles of arsenic in cores from the Laurentian trough in the gulf of St Lawrence. The surface enrichment of solid arsenic and the increase of dissolved arsenic with depth suggested that the mobile portion of arsenic is associated with iron oxyhydroxides. It follows a redox pattern of dissolution in the suboxic zone, upwards diffusion, and precipitation near the sediment–water interface under non-steady-state conditions.

Nereis succinea, a burrowing polychaete, affected distribution and flux of arsenic from sediments by its production of irrigated burrows. These burrows increased both the effective surface area of the sediment and the diffusion of arsenic by a factor of five. Although physical suspension can produce large pulses of materials from contaminated sediments, it is the continuous biological activity that is likely to be more important in the mobilization of arsenic from sediments (Riedel et al., 1987).

Riedel et al. (1989) reported that three species of burrowing invertebrates (N. succinea, Macoma balthica and Micura leidyi) cause a measurable flux of arsenic out of contaminated sediments which was not measurable in the absence of fauna. Arsenic release from sediment was primarily arsenate and arsenite, with trace amounts of methylated arsenic compounds.

4.1.4 Soil
Arsenic from weathered rock and soil may be transported by wind or water erosion. However, because many arsenic compounds tend to adsorb to soils, leaching usually results in transportation over only short distances in soil (Moore et al., 1988; Welch et al., 1988). However, rainwater or snowmelt may leach soluble forms into surface water or groundwater, and soil microorganisms may reduce a small amount to volatile forms (arsines) (Woolson, 1977a; Richardson et al., 1978; Cheng & Focht, 1979; Turpeinen et al., 1999).

Under reducing conditions, arsenite dominates in soil (Deuel & Swoboda, 1972a; Haswell et al., 1985) but elemental arsenic and arsine can also be present (Walsh & Keeney, 1975). Arsenic would be present in well-drained soils as H2AsO4– if the soil was acidic or as HAsO42– if the soil was alkaline. Oxidation, reduction, adsorption, dissolution, precipitation and volatilization of arsenic reactions commonly occur in soil (Bhumbla & Keefer, 1994). In the porewater of aerobic soils arsenate is the dominant arsenic species, with small quantities of arsenite and MMA in mineralized areas.

The amount of arsenic sorbed from solution increases as the free iron oxide, magnesium oxide, aluminium oxide or clay content of the soil increases; removal of amorphous iron or aluminium components by treatment with oxalate eliminates or appreciably reduces the arsenic sorption capacity of the soil (Dickens & Hiltbold, 1967; Jacobs et al., 1970a; Galba, 1972; Wauchope, 1975; Livesey & Huang, 1981). Barry et al. (1995) examined the adsorption characteristics of a forest soil profile. The greatest sorption capacity for arsenic occurred at a depth of 30 cm in the profile, in the B2 horizon where there was a predominance of clay and oxyhydroxides of iron and aluminium. Adsorption of arsenic on soil colloids depends on the adsorption capacity and behaviour of these colloids (clay, oxides or hydroxides of aluminium, iron and manganese, calcium carbonates or organic matter). In general, iron oxides/hydroxides are the most commonly involved in adsorption of arsenic in both acidic and alkaline soils (Sadiq, 1997). Manning & Goldberg (1997) studied the adsorption of arsenic in three arid-zone soils. They found that the soil with the highest citrate–dithionite extractable iron and percentage of clay had the highest affinity for arsenite and arsenate and displayed adsorption behaviour similar to that of pure ferric oxide. Adsorption isotherms indicated that arsenate species adsorbed more strongly than arsenite.

The surfaces of aluminium oxides/hydroxides and clay may play a role in arsenic adsorption, but only in acidic soils. Carbonate minerals are expected to adsorb in calcareous soils (Sadiq, 1997), and Goldberg & Glaubig (1988) concluded that carbonates play a major role in arsenate adsorption at pH > 9. Phosphate substantially suppresses arsenate adsorption by soil, with the extent of the suppression varying from soil to soil (Livesey & Huang, 1981). Roy et al. (1986) found that the adsorption of arsenate was significantly reduced by competitive interactions with phosphate in three different soil types (clay, silt loam and ultisol). Darland & Inskeep (1997) found that phosphate effectively competed with arsenate for adsorption sites on sand in batch isotherms as well as in saturated transport studies. The phosphate competition was not, however, sufficient to desorb all of the applied arsenate either in simultaneously applied pulses, or in a column where arsenate was applied before a concentrated pulse of phosphate. Approximately 40% of the applied arsenate remained sorbed to the sand even after the total phosphate loading exceeded the column capacity by more than two orders of magnitude. The authors concluded that rates of arsenate desorption play an important role in transport of arsenate through porous media. Elkhatib et al. (1984) found that arsenite adsorption was not reversible, with only small amounts of sorbed arsenite released during subsequent desorption procedures. No significant correlation was found between arsenic adsorption and soil organic carbon or cation exchange capacity (CEC) (Hayakawa & Watanabe, 1982).

Jones et al. (1997) found that increased mobility of arsenic after liming appears to be consistent with the pH dependence of sorption reactions of arsenic on iron oxide minerals rather than dissolution–precipitation reactions of solid metal arsenates.

Sakata (1987) reports distribution coefficients (Kd) for arsenite for 15 subsurface soils from different sites in Japan with Kd values ranging from 75 to 1200. The distribution coefficient was significantly correlated with the extractable iron content of the soils.

Precipitation is another mechanism of arsenic removal from soil. Thermodynamic calculations showed that in acidic oxic and suboxic soils, iron arsenate may control arsenic solubility, whereas in anoxic soils, sulfides of arsenite may control the concentrations of the dissolved arsenic in soil solutions. In alkaline, acidic, oxic and suboxic soils, precipitation of both iron arsenate and calcium arsenate may limit arsenic concentrations in soil solutions (Sadiq et al., 1983; Sadiq, 1997). Carey et al. (1996) studied the sorption of arsenic in two free-draining sandy soils in New Zealand. They concluded that arsenate sorption occurred primarily through adsorption rather than a precipitation mechanism.

Many soil organisms are capable of converting arsenate and arsenite to several reduced forms, largely methylated arsines which are volatile (see section 4.2). Woolson (1977b) proposed that about 12% of the arsenic applied and present in a soil is lost through volatilization of alkylarsines each year. Woolson & Isensee (1981) report total losses of 14–15% per year from soil treated with sodium arsenite, DMA or MMA. Most of the loss was through volatilisation, although some apparent loss was caused by movement to or mixing with subsoil. Sandberg & Allen (1975) estimated an arsenic loss of 17–35% per year through volatilization. Sanford & Klein (1988) report that arsenic volatilization showed a direct relationship with nutrient levels and microbial growth in soil.

Leaching does not appear to be a significant route of arsenic loss from soil. Arsenic as MMA was applied to three soil types over a 6-year period. Percentage recovery of applied arsenic averaged 67%, 57% and 39% in a fine sandy loam, a silt loam and a sandy loam soil respectively. All of the arsenic recovered in the soils was detected in the ploughed layer (< 30 cm) with no evidence of leaching into deeper zones (Hiltbold et al., 1974). Elfving et al. (1994) monitored the movement of arsenic following the application of lead arsenate to fruit orchards for insect control. The rate of decrease in concentration of arsenic with depth was significantly greater in a sandy soil than in clay, suggesting that downward movement occurred less readily in the former. Peryea & Creger (1994) studied the vertical distribution of arsenic in six contaminated orchard soils. Most of the arsenic was restricted to the upper 40 cm, with maximum arsenic concentrations ranging from 57.8 to 363.8 mg/kg. Absolute soil enrichment with arsenic occurred to depths between 45 and > 120 cm, with arsenic concentrations of 5.3–47.3 mg/kg at 120 cm. The authors state that the deeper movement found in this study compared with many others is due to high loading rates of lead arsenate, coarse soil texture, low organic matter content and use of irrigation. The use of phosphate fertilizers significantly increases the amount of arsenic leached from soil contaminated with lead arsenate pesticide residues (Davenport & Peryea, 1991).

Masscheleyn et al. (1991a) found that at soil Eh levels of 200 and 500 mV arsenic solubility was low and the major part (65–98%) of the arsenic in solution was arsenate. Under moderately reduced soil conditions (at 0 and –100 mV) arsenic solubility was controlled by the dissolution of iron oxyhydroxides. Arsenic was co-precipitated as arsenate with iron oxyhydroxides and released on solubilization. On reduction to –200 mV the soluble arsenic content increased to 13 times what it was at 500 mV.

Richardson et al. (1978) monitored surface runoff of arsenic from a fine montmorillonitic clay after application of arsenic acid for desiccation of cotton (Gossypium hirsutum). They calculated that approximately 7% of the amount applied would be transported from the watershed by runoff and erosion, 38% in solution and 62% attached to sediment.

Tammes & de Lint (1969) calculated an average half-life of 6.5 ± 0.4 years for arsenic persistence on two Netherlands soils after application of arsenite.

4.2 Biotransformation
Most environmental transformations of arsenic appear to occur in the soil, in sediments, in plants and animals, and in zones of biological activity in the oceans. Biomethylation and bioreduction are probably the most important environmental transformations of the element, since they can produce organometallic species that are sufficiently stable to be mobile in air and water. However, the biomethylated forms of arsenic are subject to oxidation and bacterial demethylation back to inorganic forms (IPCS, 1981, section 4).

Three major modes of biotransformation of arsenic species have been found to occur in the environment: redox transformation between arsenite and arsenate, the reduction and methylation of arsenic, and the biosynthesis of organoarsenic compounds. There is biogeochemical cycling of compounds formed by these processes (Andreae, 1983).

Arsenic is released into the atmosphere primarily as As2O3 or, less frequently, in one of several volatile organic compounds, mainly arsines (US EPA, 1982). Trivalent arsenic and methyl arsines in the atmosphere undergo oxidation to the pentavalent state, and arsenic in the atmosphere is usually a mixture of the trivalent and pentavalent forms (Scudlark & Church, 1988). Photolysis is not considered an important breakdown process for arsenic compounds (Callahan et al., 1979).

Arsenic can undergo a complex series of transformations, including redox reactions, ligand exchange and biotransformation (Callahan et al., 1979; Welch et al., 1988). Factors affecting fate processes in water include the Eh, pH, metal sulfide and sulfide ion concentrations, iron concentrations, temperature, salinity, and distribution and composition of the biota (Callahan et al., 1979; Wakao et al., 1988).

4.2.1 Oxidation and reduction
Oscarson et al. (1980) observed oxidation of arsenite (10 mg/litre) to arsenate in sediments from lakes in Saskatchewan (Canada). The oxidation process was unaffected by flushing nitrogen or air through the system or by the addition of mercuric chloride. The authors therefore concluded that the oxidation was an abiotic process, with microorganisms playing a very minor role in the system. However, Scudlark & Johnson (1982) examined the oxidation of arsenite in seawater at low levels. They found that abiotic oxidation proceeded at a slow and constant rate with rapid oxidation occurring only in the presence of certain aquatic bacteria. The rate of abiotic oxidation, after spiking water with an initial arsenite concentration of 4 µg/litre (53 nmol/litre), was 0.2 µg/litre per day in distilled water and 0.3 µg/litre per day in artificial seawater. Baker et al. (1983a) found no methylated arsenic compounds in sterile lake sediments incubated in the presence of arsenate or arsenite.

Scudlark & Johnson (1982) studied the biological oxidation of arsenite in seawater in Narragansett bay (Rhode Island, USA). They found that oxidation was primarily due to microbial activity. Oxidation obeyed first-order kinetics with a rate constant of 0.06 h–1 and half-lives ranging from 8.9 to 12.8 h for initial arsenite concentrations ranging from 7.5 µg/litre to 6.9 mg/litre (0.1–91.8 µmol/litre). Under aerobic conditions the mixed microbial cultures of lake sediments were able to reduce arsenate to arsenite and also to oxidize arsenite to arsenate. However, under anaerobic conditions only reduction was observed (Freeman et al., 1986).

In seawater containing free dissolved oxygen, arsenate is the thermodynamically stable form of the element. Arsenite is present in amounts exceeding those of arsenate only in reduced, oxygen-free porewaters of sediments and in anoxic basins such as the Baltic sea. However, significant amounts of arsenite (up to 10% of total arsenic) are found in the surface and deep waters of the oceans and, conversely, some arsenate is still present in anoxic water (Andreae, 1983). The presence of arsenite in seawater suggests that some reduction of arsenate occurs, and indeed Johnson (1972) demonstrated that bacterial arsenate reduction can take place under laboratory conditions. Matsuto et al. (1984) isolated a cyanobacterium (Phormidium sp.) from the coastal marine waters of Suruga bay (Japan) that was capable of readily reducing adsorbed arsenate to arsenite.

Freeman (1985) isolated an Anabaena oscillaroides–bacteria assemblage from the arsenic-rich Waikato river (New Zealand) capable of reducing arsenate to arsenite. In continuous culture the cyanophyte–bacteria assemblage could reduce arsenate to arsenite at a rate of 12 ng As/106 cells per day. Wakao et al. (1988) detected microbial arsenite oxidation occurring in acid mine waters (pH 2.0–2.4) containing 2–13 mg As/litre. Ahmann et al. (1994) isolated a microorganism from arsenic-contaminated sediment in eastern Massachusetts (USA) which used the reduction of arsenate to arsenite to gain energy for growth. Similarly, Macy et al. (1996) found that an anaerobic bacterium Chrysiogenes arsenatis from gold-mine wastewater grew by reducing arsenate to arsenite using acetate as the electron donor and carbon source.

On the bassis of both aqueous and solid-phase observations, McGeehan (1996) found that arsenate was reduced to arsenite in flooded soil under batch conditions. Reduction of arsenate to arsenite has also been reported for both freshwater and marine macroalgae (Blasco, 1975; Johnson & Burke, 1978; Andreae & Klumpp, 1979; Wrench & Addison, 1981). Calculations based on the measured rates of reduction indicate that 15–20% of the total arsenic is reduced by phytoplankton during spring and autumn blooms on the continental shelf (Sanders & Windom, 1980).

4.2.2 Methylation
The biomethylation of arsenic was first recognized when arsines were produced from cultures of a fungus, Scopulariopsis brevicaulis (Challenger, 1945). Subsequently, the methylation of arsenic by methanogenic bacteria (McBride & Wolfe, 1971) and by reaction with methyl cobalamine (Schrauzer et al., 1972) or l-methionine-methyl-d3 (Cullen et al., 1977) has been demonstrated in laboratory work. Cox & Alexander (1973) showed that cultures of the fungus Candida humicola methylate arsenite, arsenate, methylarsonate and DMA to trimethylarsine. Further experiments have shown that growing cells of C. humicola can be induced to produce trimethylarsine from arsenate and DMA by preconditioning with DMA (Cullen et al., 1979b). Cullen et al. (1979a) incubated C. humicola in the presence of 74As-arsenate, 14C-methylarsonate or 14C-DMA. They identified arsenite, methylarsonate, DMA and trimethylarsine oxide as intermediates in a biological synthesis of trimethylarsine. However, they tentatively conclude that methylarsonate does not occur as a free intermediate in the arsenate to trimethylarsine pathway.

McBride et al. (1978) reported that dimethylarsine was mainly produced by anaerobic organisms, whereas trimethylarsine resulted from aerobic methylation.

Methylated arsenic compounds were detected in aerobic sediments from various locations in Ontario (Canada) incubated with or without the addition of extraneous arsenic. Two pure bacterial cultures, Aeromonas sp. and Flavobacterium sp., isolated from lake water, were also found to methylate arsenic compounds in a synthetic medium (Wong et al., 1977).

Baker et al. (1983a) incubated lake sediment in the presence of arsenate or arsenite (7.5 mg As/litre). Methylation occurred over the pH range 3.5–7.5, with analysis revealing the presence of both methyl arsonic acid and dimethylarsinic acid. The amount of arsenic recovered in the methylated species ranged from 0 to 0.4% of the total inorganic arsenic added. Maeda et al. (1988) exposed the cyanobacterium Phormidium sp. (isolated from an arsenic-polluted environment) to arsenate (128 mg/kg) and found that 3.2% of the accumulated arsenic had been methylated.

Huysmans & Frankenberger (1991) isolated a Penicillium sp. from evaporation pond water capable of methylating and subsequently volatilizing organic arsenic. The conditions optimum for trimethylarsine production were a minimal medium containing 100 mg/litre methylarsonic acid, pH 5–6, a temperature of 20 °C and a phosphate concentration of 0.1–50 mmol/litre.

Reimer & Thompson (1988) found a strong positive correlation between the sum of the methylarsenic compounds and the total dissolved arsenic in marine interstitial waters influenced by mine tailings discharges indicating in situ microbial methylation. Laboratory studies have shown that microorganisms present in both natural marine sediments and sediments contaminated with mine tailings are capable of methylating arsenic under aerobic and anaerobic conditions (Reimer, 1989).

Biomethylation is primarily restricted to the high-salinity regions of estuaries with the presence of methylated arsenic at lower salinities predominantly as a result of the mixing of saline water (containing bioarsenicals) with river water (Howard & Apte, 1989).

Several authors have reported arsenic methylation in macroalgae, particularly in marine organisms (Edmonds & Francesconi, 1977; Andreae & Klumpp, 1979; Wrench & Addison, 1981; Maeda et al., 1987b; Cullen et al., 1994). In fact, most diatoms, dinoflagellates and macroalgae as well as freshwater higher plants, release protein-bound arsenic as a result of sequential methylation and adenosylation (Benson et al., 1988). Baker et al. (1983b) reported that freshwater green algae were capable of methylating sodium arsenite in lake water. Analysis revealed the presence of MMA, DMA and trimethylarsine oxide; however, volatile arsine and methylarsines were not detected. Similarly, Wrench & Addison (1981) identified MMA and DMA as polar arsenic metabolites synthesized by the marine phytoplankton Dunaliella tertiolecta. Maeda et al. (1987b) exposed five arsenic-resistant freshwater algae from an arsenic-polluted environment to arsenate. Small amounts of methylated arsenic compounds were detected and these were strongly bound with proteins or polysaccharides. Methylated arsenic compounds were found mainly in the lipid-soluble fractions and the major form was a dimethyl arsenic compound. No methylation occurred in algal cells (Chlorella vulgaris) exposed to arsenate under in vitro conditions; however, in vivo a small fraction of the arsenic accumulated was first transformed to methyl and dimethyl arsenic compounds during the early exponential phase and finally transformed to trimethylarsenic species (Maeda et al., 1992b). The marine algae Ecklonia radiata and Polyphysa peniculus methylated arsenate to produce a dimethylarsenic derivative. It was concluded that methionine or S-adenosylmethionine was the source of the methyl groups in this biological alkylation (Edmonds & Francesconi, 1988a; Cullen et al., 1994). S-adenosylmethionine is also likely to be the source of adenosyl and ribosyl groups in the arsenosugars.

The organic arsenical arsenobetaine was first identified in the late 1970s (Edmonds & Francesconi, 1981b) and has now been isolated in a variety of marine organisms (Edmonds & Francesconi, 1981b; Norin & Christakopoulos, 1982; Shiomi et al., 1984; Edmonds et al., 1992). Edmonds & Francesconi (1981a) identified arsenosugars isolated from brown kelp (Ecklonia radiata) as intermediates in the cycling of arsenic and stated that these compounds could be subsequently metabolized to arsenobetaine. Edmonds et al. (1982) have shown that the simpler arsenosugars in the brown alga are degraded under anaerobic conditions to dimethyloxarsylethanol. The transformation of dimethyloxarsylethanol to arsenobetaine would require both a reduction-methylation step and an oxidation step; these are probably bacterially mediated (Edmonds & Francesconi, 1987a, 1988b). Edmonds & Francesconi (1988b) concluded that arsenobetaine is probably formed by the conversion of arsenate to dimethyl(ribosyl)arsine oxides by algae, and that the microbially mediated transformation to arsenobetaine or its immediate precursors occurs in sediments. Phillips & Depledge (1985, 1986) proposed that phospholipids containing arsenoethanolamine or arsenocholine moieties may be formed as intermediates in the formation of arsenosugars and arsenobetaine. Edmonds et al. (1992) identified arsenocholine-containing lipids as natural products in the digestive gland of the rock lobster (Panulirus cygnus). Phillips & Depledge (1985) concluded that arsenic replaces nitrogen in phospholipid synthesis leading to a large number of arsenic-containing intermediates, which would be either water-soluble or lipid-soluble. Arsenic-containing compounds are catabolized as they pass through the food web, yielding arsenobetaine as a stable end-product.

Inorganic arsenic administered orally to brown trout (Salmo trutta) was detected in tissues as organoarsenical species, whereas arsenic administered by injection was taken up as inorganic arsenic and slowly converted to the organic form. It was concluded that biosynthesis of arsenic was occurring in the gastrointestinal tract (Penrose, 1975). Oladimeji et al. (1979) reported that arsenic given as an oral dose to rainbow trout (Oncorhynchus mykiss) was rapidly converted to organic forms. The ratio of total organic to inorganic increased with time in all tissues, with the organic arsenic fraction accounting for about 50% after 6 h and over 80% within 24 h. The major organic arsenical appeared to be an arsenobetaine-related compound. Similarly, Penrose et al. (1977) found that sea urchins (Strongylocentrotus droebachiensis) were also able to convert inorganic arsenic to an organic form, but to a more limited degree than trout. However, Wrench et al. (1981) concluded that organic arsenic synthesized in the brine shrimp (Artemia salina) is methylated by intestinal microflora and not by the filter feeder itself.

Maeda et al. (1990c) found that 85% of arsenic accumulated by the guppy (Poecilia sp.) was in the di- and tri-methylated forms. The percentage of organic species was much higher than that found in phytoplankton and zooplankton in the same model ecosystem. Similarly, Maeda et al. (1990a) found that biomethylation of arsenic increased successively with trophic level in another model ecosystem: goldfish (Carassius sp.) > zooplankton (Moina sp.) > alga (Chlorella sp.).

4.2.3 Degradation Abiotic degradation
The rates of photochemical decomposition of arsenite, DMA, MMA and arsenobetaine have been studied in both distilled water and seawater. All species were found to degrade rapidly in aerated distilled water. In deaerated solutions the rate of oxidation of arsenite was almost two orders of magnitude slower. Half-lives for the degradation of DMA, MMA and arsenite were 9.2, 11.5 and 0.9 min respectively for aerated distilled water and 25, 19 and 8 min for deaerated distilled water. In seawater, the rates of photochemical decomposition were slower. For example, in seawater only 20% of DMA was converted to MMA after 300 min with no other products detected, whereas in distilled water DMA was completely degraded within 100 min (Brockbank et al., 1988). This study suggests that UV irradiation is of limited use for the pretreatment of saline samples to convert organoarsenic species to As(V) before analysis. The implications for photochemical decomposition of arsenic species in natural waters is not clear, because sunlight is deficient in the lower-wavelength bands generated by the mercury lamp used in this study. In addition, colloids and suspended particulates in the photic zone may play a significant role in arsenic decomposition in natural waters.

Von Endt et al. (1968) concluded that degradation of MSMA in soil was primarily due to soil microorganisms rather than abiotic factors. In 60-day tests in non-sterile soil 1.7–10% of the 14C-MSMA was degraded, whereas under steam-sterilized conditions only 0.7% was degraded. Biodegradation
The predominant form of arsenic in water is usually arsenate (Callahan et al., 1979; Wakao et al., 1988), but aquatic microorganisms may reduce the arsenate to arsenite and a variety of methylated arsenicals.

Marine organisms tend to contain much higher levels of arsenic than terrestrial organisms; this is because of the high arsenate/phosphate ratio in oceans, which is a consequence of the very low phosphate concentration. Most of the arsenic accumulated in marine organisms is in a water-soluble form of arsenic, namely arsenobetaine. Hanaoka et al. (1987) incubated marine sediments in the presence of arsenobetaine and demonstrated microbial degradation, with arsenate, arsenite, MMA, DMA and arsenobetaine being identified. Further experiments revealed the formation of trimethylarsine oxide during aerobic incubation of bottom sediments with arsenobetaine as the carbon source (Kaise et al., 1987). Under aerobic conditions, arsenobetaine is converted to its metabolites to a much greater extent than other methylarsenicals. Under anaerobic conditions little or no degradation of arsenobetaine occurred, whereas trimethylarsine oxide and DMA were converted to less methylated compounds (Hanaoka et al., 1990). Degradation of arsenobetaine has also been demonstrated in the water column in the presence of suspended substances (Hanaoka et al., 1992).

Organoarsenical pesticides (e.g. MMA and DMA) applied to soil are metabolized by soil bacteria to alkylarsines, MMA, and arsenate (ATSDR, 1993). The half-time of DMA in soil is about 20 days (ATSDR, 1993).

Cheng & Focht (1979) added arsenate, arsenite, methylarsonate and DMA to three different soil types. Arsine was produced in all three soils from all substrates but methylarsine and dimethylarsine were only produced from methylarsonate and DMA respectively. Both Pseudomonas sp. and Alicaligenes sp. produced arsine as the sole product when incubated anaerobically in the presence of arsenate or arsenite. The authors concluded that reduction to arsine, not methylation to trimethylarsine, was the primary mechanism for gaseous loss of arsenicals from soil.

Degradation of MSMA by soil microorganisms was studied by Von Endt et al. (1968). In 60-day tests they found that 1.7–10% of the 14C-MSMA was degraded; four soil microorganisms isolated in pure cultures degraded 3–20% of 14C-MSMA to 14CO2 when grown in liquid culture at 10 mg MSMA/litre. Woolson & Kearney (1973) showed that sodium DMA was degraded to arsenate in soil under aerobic conditions but not under anaerobic conditions. Degradation of MSMA has been shown to be associated with soil organic matter oxidation. In a loamy soil, degradation increased with increasing organic matter content (Dickens & Hiltbold, 1967). Akkari et al. (1986) studied the degradation of MSMA in soils at concentrations up to 5 mg As/kg. It was found that degradation followed first-order kinetics. The rate constant was temperature dependent only at soil water contents less than field capacity, and the temperature effect was less under flooded conditions. The differences in degradation rate under aerobic conditions and 20% water content were related to differences in the texture of the three soils. Half-lives for the clay and silty loam soils were 144 and 88 days respectively. Under anaerobic (flooded) soil conditions MSMA degradation occurs by reductive methylation to form arsenite and alkylarsine gases. The half-life values for the two soils indicate significantly faster degradation at 25 and 41 days respectively. The third soil, a sandy loam, produced the slowest degradation rate (t½ = 178 days) probably because of its low organic matter content which may have supported fewer microorganisms.

The overall percentage of DMA (sodium salt) and MMA mineralized in a silty clay soil after 70 days ranged from 3% to 87% – values much higher than arsenic loss as volatile arsines (0.001–0.4%). Arsenate was the main metabolite from the degradation of both sodium DMA and MMA. The amount of sodium DMA mineralized was linearly related to the concentration of sodium DMA in the soil, indicating that the rate is first order. Mineralization of sodium DMA increased with increasing soil moisture and temperature. It was concluded that the loss of arsenic from some soils to the atmosphere may not be a major pathway and that inorganic arsenic may accumulate in soil from arsenical usage (Gao & Burau, 1997).

4.2.4 Bioaccumulation
Bioconcentration of arsenic under laboratory conditions occurs in aquatic organisms, primarily in algae and lower invertebrates. Bioconcentration factors (BCFs) measured in freshwater invertebrates for several arsenic compounds generally ranged up to 20; bioconcentration factors in fish were < 5; higher concentration factors have been observed in algae. Biomagnification in aquatic food chains does not appear to be significant (Callahan et al., 1979). Terrestrial plants may accumulate arsenic by root uptake from the soil or by adsorption of airborne arsenic deposited on the leaves, some species accumulating substantial levels. Microorganisms
Maeda et al. (1987a) exposed cyanobacteria (Nostoc sp.) to arsenate concentrations of 1 and 10 mg As(V)/litre for 32 days with no effect on cell growth. Nostoc sp. accumulated 32 and 77 mg As/kg (dry cell) respectively at the two exposure concentrations.

Lindsay & Sanders (1990) report BCFs ranging from 1132 to 3688 for estuarine phytoplankton (Thalassiosira pseudomonas, Skeletonema costatum and Dunaliella tertiolecta) exposed to 25 µg As(V)/litre as arsenate for up to 48 h.

Phytoplankton take up arsenate readily and incorporate a small proportion into the cell. Most of the arsenate is reduced, methylated and released to the surrounding media. Phytoplankton batch cultures exposed to elevated levels of arsenate take up additional arsenic during the log phase of growth. Studies using 74As indicate that the uptake rate varies from 0.15 ng As(V)/106 cells per hour in unenriched cultures to 2.3 ng As(V)/106 cells per hour in cultures containing 25 µg As(V)/litre. Cultured Skeletonema costatum increase their arsenic concentrations approximately 40% from 22 to 29 mg/kg (dry weight) in response to arsenate concentrations of 6–25 µg As(V)/litre (Sanders & Windom, 1980).

Phytoplankton readily incorporated dissolved arsenic, with average arsenic residues increasing from 5.7 to 17.7 mg/kg (dry weight) when cultured for 48–96 h at 25 µg As(V)/litre as arsenate (Sanders et al., 1989). Arsenate added to a freshwater model ecosystem was readily accumulated in plankton with arsenic residues of 37–47 mg/kg (dry weight) at 5 µg As(V)/litre and > 200 mg/kg at 50 µg As(V)/litre after 65-day exposures. Accumulation in other biota was much lower than for phytoplankton (Reuther, 1992).

Giddings & Eddlemon (1977) studied the uptake of radioactively labelled arsenic (added as sodium arsenate at 50 µg As(V)/litre) in model ecosystems (7 and 70 litres) for 5 weeks. Mean BCFs for algae ranged from 370 for sand microcosms to 4300 for lake mud microcosms. Algal arsenic concentrations were significantly greater in the 70-litre microcosms and in the sand microcosms than in the 7-litre and sediment microcosms.

Green algae (Chlorella vulgaris) exposed to arsenate concentrations of 7 to 9 mg As(V)/litre accumulated maximum residues of 3.75 g total As/kg (dry mass) within 10 days (Maeda et al., 1992c).

Maeda et al. (1985) found that arsenate uptake increased with an increase in the arsenic exposure concentration with C. vulgaris isolated from an arsenic-polluted environment. Maximum BCFs of 200–300 were observed during the log phase. At the highest exposure concentration (10 g As(V)/litre) algae were able to accumulate 50 g As/kg (dry weight). Approximately half of the arsenic taken up was estimated to be adherent to the extraneous coat of the cell with the remainder accumulated by the cell. Arsenate accumulation was affected by the growth phase; arsenic was most actively accumulated when the cell was exposed to arsenic during the early exponential phase (Maeda et al., 1992a).

Accumulation of arsenic (1 mg As(V)/litre as arsenate) by Dunaliella sp. was rapid, with equilibrium established within 8 h. Arsenic accumulation was studied at temperatures ranging from 10 °C to 33 °C, pH 4–10, light intensity from 0 to 10 000 lux and sodium chloride concentrations from 1 to 100 g/litre. Maximum arsenic residues under optimum conditions (22 °C; pH 8.2; 5000–10 000 lux and 20 g NaCl/litre) were 2000 mg As/kg. Increased phosphate significantly decreased the uptake of arsenic in the culture (Yamaoka et al., 1988). Yamaoka et al. (1992) found that D. salina accumulated more arsenic at nitrogen concentrations of 72 mg/litre than at 4.5 mg/litre. Macroalgae
Fucus vesiculosus accumulated approximately 120 mg As/kg during an 85-day exposure to 7.5 µg As(V)/litre as arsenate. Filamentous algae and planaria accumulated less than 40 mg As/kg (dry weight), and cyanobacteria and various zooplankton accumulated less than 20 mg As/kg (Rosemarin et al., 1985).

Klumpp (1980) studied the effect of a variety of factors on the uptake of labelled arsenic by the seaweed Fucus spiralis. Neither pH (pH 7–9) nor salinity (9–36 g/litre) affected the uptake of arsenic; however, uptake at 30 °C was twice that at 16 °C. Arsenate uptake was reduced with increasing phosphate concentration (40–400 µmol/litre).

Lee et al. (1991) grew the aquatic plant Hydrilla verticillata in both mine-waste pool water and deionized distilled water contaminated with arsenate (0.4 and 0.8 mg As(V)/litre) for up to 16 days. Accumulation of arsenic reached steady state at 2–6 days in pool water at BCFs of 110–190. In deionized water maximum arsenic accumulation occurred after 8 days at a BCF of around 300. Phosphate (³ 12 mg/litre) inhibited the uptake of arsenic by H. verticillata. Aquatic invertebrates
Sanders et al. (1989) studied the uptake of arsenic from water and from phytoplankton by the copepod Eurytemora affinis and the barnacle Balanus improvisus. In 24-h tests, E. affinis exhibited no uptake of dissolved arsenic; the arsenic content of copepods fed phytoplankton increased to 11.2 mg/kg (dry weight) compared with 8.9 mg/kg in controls. In 22-day tests, B. improvisus exposed to dissolved arsenate (55 µg As(V)/litre) in water did not accumulate arsenic, with levels remaining around 0.88 mg/kg; however, levels in shell material increased from 0.3 mg/kg to 2 mg/kg. Barnacles fed arsenic-contaminated phytoplankton (~18 mg/kg) exhibited an increase in total arsenic concentrations from 0.3 mg/kg to 1.7 mg/kg. In further experiments with oysters (Crassostrea virginica) no accumulation of arsenic from water was observed in 28-day tests, but tissue concentrations increased significantly from 5.3 mg/kg to 8.2 mg/kg in oysters fed arsenic-contaminated phytoplankton. Zaroogian & Hoffman (1982) reported maximum total arsenic residues in soft tissues of oysters (Crassostrea virginica) of 12.6, 12.7 and 14.1 mg/kg (dry weight) at arsenite exposure concentrations of 1.2 (control), 3 and 5 mg As(III)/litre during 112-day exposures. Generally, arsenic body burdens increased with increases in phytoplankton concentration and it appears that food contributes more to arsenic uptake than do seawater arsenic concentrations. No relationship between arsenic uptake and seawater arsenic concentrations was found.

Ünlü & Fowler (1979) exposed mussels (Mytilus galloprovincialis) to arsenate (74As) concentrations ranging from 20 to 100 µg As(V)/litre at 12 °C and 21 °C. Mean concentration factors after 20 days were low, at respectively 8.8 and 12.1 for the two temperatures; however, mussels did accumulate significantly more arsenic at 21 °C than at 12 °C. Arsenic uptake was inversely related to salinity over the range 31–19 g/litre. Arsenic loss was essentially biphasic, with biological half-times of approximately 3 and 32 days for the fast and slow compartments respectively. The active secretion of arsenic in the byssal threads contributed to the total elimination of the element from the mussels. Similarly, Ünlü (1979) found a biphasal loss of arsenic from crabs (Carcinus maenas) during a 43-day depuration period. The elimination of 74As by the crabs after ingestion of arsenic-contaminated mussels was dependent on the chemical form of the arsenic. After ingestion of mussel containing mostly lipid- and water-soluble arsenic species (undetermined), biological half-times were 3.4 and 19.6 days for the first and second phase of loss. After ingestion of mussel containing mostly arsenite and residual arsenic, half-times were 1.6 and 9.3 days respectively.

Naqvi et al. (1990) exposed red crayfish (Procambarus clarkii) to MSMA at concentrations of 0.5, 5 and 50 mg As/litre. Uptake of arsenic was dose-dependent but not time-dependent. Maximum whole-body residues were 1.36, 4.29 and 9 mg As/kg respectively for each of the exposure concentrations during the 8-week uptake period.

Gibbs et al. (1983) reported equilibrium BCFs, based on 74As, for the cirratulid polychaete Tharyx marioni ranging from 4.5 at an exposure concentration of 10 mg As(V)/litre (as arsenate) to 111.6 at 0.01 mg/litre after 7 days. A lower BCF of only 15.9 at 0.01 mg/litre was reported for the polychaete Caulleriella caputesocis.

Shrimps exposed to water concentrations ranging from 0.1 to 1.5 mg As(V)/litre (as arsenate) or food (Chlorella sp.) containing, 1940 mg total As/kg contained arsenic residues ranging from 18.9 to 31.8 mg/kg (dry weight) (Maeda et al., 1992c).

Fowler & Ünlü (1978) reported BCFs of less than 10 for shrimps exposed to arsenate (74As) concentrations of 20–100 µg As(V)/litre for 14 days. Arsenic loss was biphasic with half-lives of 3 and 26 days for the fast and slow compartments respectively. Moults shed during loss contained 2–5% of the shrimp’s 74As body burden.

Lindsay & Sanders (1990) found no bioaccumulation of arsenate directly from the water (25 µg As(V)/litre) or from food for the grass shrimp (Palaemonetes pugio). Brine shimp (Artemia sp.) grown in elevated arsenic concentrations exhibited small, but significant, increases in arsenic content from an average of 16.8 mg/kg (dry weight) in controls to 17.8 mg/kg at 25 µg As(V)/litre; no accumulation was observed when brine shrimps were fed arsenic-contaminated food. Fish
Barrows et al. (1980) exposed bluegill sunfish (Lepomis macrochirus) to 130 µg As(III)/litre of As2O3 for 28 days. The maximum BCF was found to be 4, with a half-life in tissues of 1 day. Nichols et al. (1984) found no accumulation of arsenic in a 6-month study on coho salmon (Oncorhynchus kisutch) exposed to As2O3 concentrations of < 300 µg As(III)/litre. Whole-body residues were below 0.4 mg As/kg (wet weight) and were not dose dependent.

Sorensen (1976) found that green sunfish (Lepomis cyanellus) exposed to higher arsenic concentrations of 100, 500 and 1000 mg As(V)/litre (as arsenate) accumulated whole-body arsenic concentrations of 33.4, 541.2 and 581.6 mg/kg (BCFs ranging from 0.3 to 1.1). Green sunfish exposed to 60 mg As(V)/litre for 6 days accumulated mean arsenic residues of 158.7, 47.7, 18.9 and 14.2 mg/kg in the gallbladder (plus bile), liver, spleen and kidney respectively (BCFs ranging from 0.2 to 2.6) (Sorensen et al., 1979).

Cockell & Hilton (1988) fed rainbow trout (O. mykiss) on diets containing As2O3 (180–1477 mg As/kg diet), disodium arsenate heptahydrate (DSA) (137–1053 mg As/kg diet), DMA (163–1497 mg As/kg diet) or arsanilic acid (193–1503 mg As/kg) for 8 weeks. For each of the arsenicals investigated, carcass arsenic concentration showed a dose–response relationship to dietary arsenic concentration and exposure rate. At lower levels of exposure (137 mg As/kg diet), dietary DSA yielded the highest mean carcass arsenic concentrations (6.9 mg As/kg), but at higher levels, dietary As2O3 (1477 mg As/kg diet) yielded the highest mean residues (21.6 mg As/kg). Inorganic arsenicals were accumulated from the diet to a greater degree than the organic forms. In a 16-week study, dietary DSA (8–174 mg As/kg diet) accumulated in the carcass (0.25–5.7 mg As/kg), liver (0.7–34.4 mg As/kg) and kidney (1.1–31.9 mg As/kg) in a dose-related manner (Cockell et al., 1991).

Oral administration of sodium arsenate to estuary catfish (Cnidoglanis macrocephalus) and school whiting (Sillago bassensis) resulted in an accumulation of trimethylarsine oxide in their tissues (Edmonds & Francesconi, 1987b). Yelloweye mullet (Aldrichetta forsteri) fed the organic arsenicals 2-dimethylarsinylethanol, 2-dimethylarsinylacetic acid or 2-dimethyllarsinothioylethanol showed no arsenic accumulation in their tissues; fish fed arsenate-contaminated food showed a small but significant increase in arsenic concentration (muscle tissue = 1 mg As/kg wet weight). However, administering arsenobetaine or arsenocholine in the diet led to muscle concentrations of around 24 mg As/kg (wet weight) (Francesconi et al., 1989).

Oladimeji et al. (1984) fed rainbow trout (O. mykiss) on a diet containing 10, 20 or 30 mg As(III)/kg (as sodium arsenite) (equivalent to 0.2, 0.4 and 0.6 mg/kg fish wet weight per day) for up to 8 weeks. Arsenic accumulation was dose related, with residues ranging from 1.28 to 1.52 mg/kg (dry weight) for muscle, 1.55 to 5.21 mg/kg for liver, 0.84 to 1.88 mg/kg for gills and 1.21 to 1.98 mg/kg for skin tissue. Terrestrial plants
Arsenic species can enter into edible tissues of food crops through absorption (i.e. not just surface contamination) (Woolson, 1973; Helgesen & Larsen, 1998). Helgesen & Larsen (1998) demonstrated that bioavailability of arsenic pentoxide to carrots in soil from a wood preservative treatment plant (soil was contaminated with CCA) was 0.47 ± 0.06% of total soil arsenic burden. This study showed that arsenite, arsenate, MMA and DMA were present in carrot tissue, where only arsenite and arsenate were present in soil. In soils dosed with arsenate (0–500 µg/g) at the concentrations which inhibited growth of vegetable crops (green bean, lima bean, spinach, cabbage, tomato and radish), high levels of accumulation when found in the edible parts of radish (76 µg/g) spinach (10 µg/g) and green bean (4.2 µg/g). Arsenic accumulation in Lima bean, cabbage and tomato ranged from 0.7–1.5 µg/g. The studies of Woolson (1973) and Helgesen & Larsen (1998) highlight the potential of movement of arsenic species from soil into agronomic crops.

Uptake of arsenate (10 mg/litre [133 µmol/litre]) by moss (Hylocomium splendens) from nutrient solution displayed saturation kinetics at pH 5 that could be described in terms of Michaelis–Menten parameters with a mean Km value of 31.4 mg/litre (418 µmol/litre). Phosphate was a competitive inhibitor of arsenate uptake with an inhibition constant (Km phosphate) of 82 µmol/litre (Wells & Richardson, 1985).

Asher & Reay (1979) studied the uptake of arsenate (1 mg/litre [15 µmol/litre]) from nutrient solution by barley (Hordeum vulgare) seedlings. They found that uptake consisted of a rapid initial phase followed by a less rapid ‘steady-state’ phase, both of which were strongly inhibited by phosphate and positively correlated with temperature.

The marsh plant species Spartina alterniflora was grown in sediment treated with ~50 µg As(V)/litre (as arsenate) and accumulated significantly elevated total concentrations of arsenic after 9 days; new and old leaf blades contained mean arsenic concentrations of 6.3 and 5 mg/kg (dry weight) respectively, relative to 1 and 0.4 mg/kg in control plants (Sanders & Osman, 1985).

Meharg & Macnair (1991b) found that non-tolerant genotypes of Holcus lanatus accumulated significantly more arsenate than tolerant plants during a 6-h period of growth in 3.75 mg As(V)/litre (0.05 mol/m3) arsenate. They found that tolerant plants transported a much greater proportion of arsenic to their shoots than non-tolerant plants. Phosphate (0.05 or 0.5 mol/m3) decreased arsenate uptake in both tolerant and non-tolerant genotypes. Arsenate tolerance involves reduced accumulation of arsenate through suppression of the high-affinity phosphate–arsenate uptake system (Meharg et al., 1994).

Anastasia & Kender (1973) grew lowbush blueberry (Vaccinium angustifolium) plants in greenhouse soil at As2O3 concentrations ranging from 7.7 (controls) to 84.5 mg As(III)/kg for 17 weeks. Arsenic was accumulated in a dose-dependent manner with arsenic residues of 0.78–15 mg/kg for leaves, 0.27–13.3 mg/kg for stems and 2.4–164.2 mg/kg for roots.

Otte et al. (1990) grew Urtica dioica and Phragmites australis in soil containing up to 30 mg As/kg added as lead arsenate or sodium DMA. Concentrations of arsenic in shoots and roots of P. australis increased significantly only at the highest arsenic concentration in soil with mean values of up to 1 mg/kg (dry weight) in shoots and 44.3 mg/kg in roots, whereas the arsenic content of U. dioica increased by a factor of 4 at 5 mg As/kg with plants accumulating mean arsenic concentrations of up to 150 mg/kg in roots at the highest exposure.

Onken & Hossner (1995) grew rice (Oryza sativa) in two soil types treated with up to 45 mg As(III) or As(V)/kg (as arsenite or arsenate) for 60 days. The arsenic concentration of rice plants correlated with the mean soil solution arsenate concentration in the clay soil and to the mean soil solution arsenite for the silt loam. The rate of arsenic uptake by plants increased as the rate of plant growth increased. Terrestrial invertebrates
Meharg et al. (1998) exposed earthworms (Lumbricus terrestris) to arsenate (40 mg/kg dry weight) for 23 days. There was a steady-state increase in residues for depurated and undepurated worms and by 12 days earthworm residues were equivalent to those of the soil. Arsenic residues were accumulated to three times soil levels by the end of the 23-day exposure in depurated worms; however, undepurated worms did not appear to bioconcentrate arsenic beyond the level of the surrounding soil. Birds
Daghir & Hariri (1977) administered arsanilic acid (used as a feed medication for poultry) to White Leghorn laying hens at 50 and 100 mg/kg for 15 weeks. Maximum concentrations in eggs were reached after 4–5 weeks at 0.13 and 0.24 mg As/kg (dry weight) for the two dose levels respectively. Residual arsenic was negligible 2 weeks after the withdrawal of the drug from the feed.

Proudfoot et al. (1991) found a higher concentration of arsenic in liver and muscle of broilers that were fed arsanilic acid (99 mg/kg diet) compared with controls. Mean arsenic residues of up to 1.5 mg/kg and 0.4 mg/kg were measured for the two tissues respectively. Broilers fed a diet containing 100 or 500 mg/kg arsanilic acid accumulated up to 2.3 and 8 mg As/kg in liver tissue at the two exposure concentrations respectively. Lower levels were accumulated in muscle tissue, with arsenic concentrations of up to 0.15 and 0.67 mg/kg (VanderKop & MacNeil, 1989).

Holcman & Stibilj (1997) fed Rhode Island Red hens on diets containing 7.5, 15 or 30 mg As(III)/kg (as As2O3) for 19 days. Eggs were collected on days 8–19 day of the experiment, and arsenic residues were consistent throughout this period. Mean concentrations were respectively 0.2, 0.42 and 0.96 mg As/kg (dry weight) in egg yolk and in 0.06, 0.14 and 0.3 mg As/kg egg white for the three exposure concentrations.

Hoffman et al. (1992) fed mallard on a diet containing 200 mg As(V)/kg (as sodium arsenate) for 4 weeks. Arsenic accumulated in the liver at a concentration of 2.3 mg As/kg (wet weight). Birds maintained on a restricted protein and exposed to the same arsenic-contaminated diet accumulated 5.1 mg As/kg. Stanley et al. (1994) maintained mallards on diets containing 25, 100 or 400 mg As(V)/kg (as sodium arsenate) for 16–18 weeks. Arsenic was accumulated in a dose dependent manner; mean concentrations in adult livers were 0.49–6.6 mg As/kg (dry weight), in duckling livers from 0.65–33 mg/kg and in whole eggs from 0.46–3.6 mg/kg.

5.1 Environmental levels
Arsenic is a natural component of the earth’s crust, and found in all environmental media. Concentrations in air in remote locations range from < 1 to 3 ng/m3, but concentrations in cities may range up to 100 ng/m3. Concentrations in water are usually < 10 µg/litre, although higher concentrations can occur near natural mineral deposits or anthropogenic sources. Natural levels in soils usually range from 1 to 40 mg/kg, but pesticide application or waste disposal can produce much higher values.

5.1.1 Air
Levels of arsenic in ambient air are summarized in Table 3. Examples are given of mean total arsenic concentrations in remote and rural areas ranging from 0.02 to 4 ng/m3. Levels of arsenic in outdoor air near to urban and industrial sources are summarized in Table 4. Examples are given of mean total arsenic concentrations in urban areas ranging from 3 to 200 ng/m3; much higher concentrations (> 1000 ng/m3) have been measured in the vicinity of industrial sources. Arsenic in ambient air is usually a mixture of arsenite and arsenate, with organic species being of negligible importance except in areas of substantial methylated arsenic pesticide application or biotic activity. Schroeder et al. (1987) reviewed worldwide arsenic concentrations associated with particulate matter. They identified arsenic levels ranging from 0.007 to 1.9 ng/m3 for remote areas, 1 to 28 ng/m3 for rural areas and 2 to 2320 ng/m3 in urban areas. The highest arsenic levels detected in the atmosphere were near non-ferrous-metal smelters.

Typical background levels for arsenic are now 0.2–1.5 ng/m3 for rural areas, 0.5–3 ng/m3 for urban areas and < 50 ng/m3 for industrial sites (DG Environment, 2000).

Table 3. Concentrations of As in ambient aira



Particle size and/or species






Brimblecombe (1979)

Antarctic Ocean



0.05 (0.01–0.2)

Nakamura et al. (1990)



0.002 (single sample)

Nakamura et al. (1990)

North Pacific Ocean



0.1 (0.01–0.95)

Nakamura et al. (1990)



0.008 (0.001–0.03)

Nakamura et al. (1990)

North Atlantic Ocean



0.1 (0.01–0.45)

Nakamura et al. (1990)



0.007 (0.001–0.3)

Nakamura et al. (1990)

Baltic Sea


1.1 (0.3–3.7)

Häsänen et al. (1990)

Mid-Atlantic coast, USA



Scudlark & Church (1988)

Continental shelf waters, south-eastern USA



1.7 (0.2–4.3)

Waslenchuk (1978)

Northern Chesapeake Bay, USA


< 10 µm

0.66 (0.11–1.96)

Wu et al. (1994)

Rural US sites (National Parks)


0.45 µm

< 1.6–2.3 (range of means)

Davidson et al. (1985)

Midwestern USA


1.6 (0.7–2.5)

Burkhard et al. (1994)

Natural geysers, northern California, USA



0.22 & 0.54 (0.06–3.08)

Solomon et al. (1993)


0.46 (0.08–1.3) & 2.29 (0.7–6.54)

Solomon et al. (1993)

Bagauda, Nigeria



Beavington & Cawse (1978)

Pelindaba, South Africa



Beavington & Cawse (1978)

Chilton, United Kingdom



Beavington & Cawse (1978)

Rural sites, United Kingdom


1.5–2.5 (range of means)

Peirson et al. (1974)

Rural area near Thessaloniki, Greece



Misaelides et al. (1993)

Birkenes, Norway



1.2 (0.02–12)

Amundsen et al. (1992)



0.63 (< 0.04–4.6)

Amundsen et al. (1992)

a As = inorganic As; As = organic As; NS = not stated

b Mean and ranges of total As unless stated otherwise

Table 4. Concentrations of As in outdoor air near urban and industrial sources


Distance from source (km)

Sampling period

Particle size and/or species

Concentration (ng/m3)a


Industrial sites, UK




1.2–24 (ng/kg, range of means)

Peirson et al. (1974)

Urban area, Thessaloniki, Greece



0.45 µm


Misaelides et al. (1993)

Urban area, Yokohama, Japan



0.45 µm; inorganic As

2.5 (1–5.1)

Nakamura et al. (1990)

0.45 µm; organic As

0.01 (0.001–0.64)

Nakamura et al. (1990)

Los Angeles, USA



< 2.5 µm; As(III)

7.4 (< 1.2–44)

Rabano et al. (1989)



> 2.5 µm; As(III)

1.8 (< 0.9–4.8)

Rabano et al. (1989)



< 2.5 µm; As(V)

5.2 (< 0.9–18.7)

Rabano et al. (1989)



> 2.5 µm; As(V)

2.2 (< 0.8–6.6)

Rabano et al. (1989)

Wuhan City, China



< 2.5 µm


Waldman et al. (1991)



³ 2.5 µm < 10 µm


Waldman et al. (1991)

Calcutta, India



0.45 µm

180 (91–512)

Chakraborti et al. (1992)

Kola peninsula, Russia near (Cu–Ni smelter)





Kelley et al. (1995)

Caletones, Chile (near Cu smelter)

< 10


0.4 µm


Romo-Kröger & Llona (1993)

< 20


0.4 µm


Romo-Kröger & Llona (1993)

< 30


0.4 µm


< 10


0.8 µm


Romo-Kröger & Llona (1993)

< 20


0.8 µm


< 30


0.8 µm


Romo-Kröger & Llona (1993)



< 2.5 µm


Romo-Kröger et al. (1994)



2.5–10 µm


a Mean and ranges of total As unless stated otherwise

b Fine particle concentration was 23 ng/m3 during a strike period at the smelter

NS = not stated

5.1.2 Precipitation
Arsenic has been detected in rainwater at mean concentrations of 0.2–0.5 µg/litre (Welch et al., 1988). Peirson et al. (1974) report mean arsenic concentrations in rainfall ranging from < 6 µg/litre for a rural site to 45 µg/litre at a North Sea gas platform. Arsenic concentrations in precipitation at the mid-Atlantic coast of the USA ranged from < 0.005 to 1.1 µg/litre with an average of 0.1 µg/litre (Scudlark & Church, 1988). Andreae (1980) collected rainwater samples from non-urban sites in California (USA) and state parks in Hawaii and found mean arsenic concentrations ranging from 0.013 to 0.032 µg/litre. Samples from a rural site in Washington state (USA) contained a mean concentration of 1.1 µg As/litre; the author states that the site is 154 km north of a large copper smelter. Vermette et al. (1995) monitored arsenic levels in wet deposition at three sites (Colorado, Illinois and Tennessee, USA) and found mean concentrations ranging from 0.09 to 0.16 µg/litre. Reimann et al. (1997) monitored rainwater samples during the summer of, 1994 in eight Arctic catchments. Median arsenic concentrations (0.45 µm) ranged from 0.07 µg/litre at the most remote site to 12.3 µg/litre near a smelter.

Barbaris & Betterton (1996) analysed fresh snowpack samples from high-elevation forests of north-central Arizona (USA) during late winter and early spring 1992–1994. Arsenic concentrations ranged from 0.02 to 0.4 µg/litre with a mean value of 0.14 µg/litre.

5.1.3 Surface water
Levels of arsenic in seawater are summarized in Table 5. Concentrations of arsenic in open ocean seawater are typically 1–2 µg/litre. The dissolved forms of arsenic in seawater include arsenate, arsenite, MMA and DMA, with adsorption on to particulate matter being the physical process most likely to limit dissolved arsenic concentrations (Maher & Butler, 1988). Levels of arsenic in estuarine water are summarized in Table 6. Tremblay & Gobeil (1990) noted that arsenic concentrations increased with increasing salinity (0–31 g/litre) from 0.5 to 1.4 µg/litre (6.6 to 18.9 nmol/litre) in the St Lawrence estuary (Canada) and from 0.1 to 1.4 µg/litre (1.1 to 18.7 nmol/litre) in its tributary Saguenay fjord. Penrose et al. (1975) monitored seawater in Moreton’s Harbour, Newfoundland near a long-term stibnite mine. Total inorganic arsenic concentrations were 5.3 µg/litre near the mine but declined to normal levels (1–2 µg/litre) within 200 m.

Howard et al. (1988) report that concentrations of methylated arsenic increased with salinity in the Tamar estuary (United Kingdom). Concentrations of monomethylarsenic ranged from 0.02 to 0.46 µg As/litre and dimethylarsenic from 0.02 to 1.27 µg As/litre; these two methylated forms of arsenic were typically 4% and 10% of the total soluble arsenic levels respectively.

Levels of arsenic in surface freshwaters are summarized in Table 7. Surveys of arsenic concentrations in rivers and lakes indicate that most values are below 10 µg/litre, although individual samples may range up to 1 mg/litre (Page, 1981; Smith et al., 1987; Welch et al., 1988). Mean total arsenic concentrations of 2000 µg/litre have been recorded near a pesticide plant, with MMA being the predominant arsenic species (Faust et al., 1983; 1987a). Crearley (1973) measured arsenic in two lakes near a manufacturing plant which had been producing arsenic-based cotton desiccants/defoliants for 30 years. Mean arsenic concentrations of 7900 and 3200 µg/litre were found. During the dry season total dissolved arsenic concentrations (< 0.45 µm) of up to 250 µg/litre were recorded near industrial discharges to the Xiangjiang river (China); however, maximum levels during the rainy season were generally less than 30 µg/litre (Chunguo & Zihui, 1988).

High levels of arsenic have been recorded in thermal waters. Tanaka (1990) found a mean concentration of 570 µg/litre in geothermal waters throughout Japan, with a maximum level of 25.7 mg/litre.

Table 5. Background concentrations of As in seawater


Sampling period

Sampling details and/or species

Concentration (µg/litre)a


Gulf of Mexico


0.2 µm filtered


Chakraborti et al. (1986)

Pacific Ocean


no arsenite detected

1.8 (1.6–2.1)

Bodewig et al. (1982)


Andreae (1978, 1979)

Coastal waters, South Australia


dissolved; particulate As below limit of detection (0.6 ng/litre)

1.3 (1.1–1.6)

Maher (1985a)

Continental shelf waters, south-eastern USA


depth 30 m and 500 m

1.1 and 1.5

Waslenchuk (1978)

Coastal waters, south-east Spain


below surface

1.5 (0.45–3.7)

Navarro et al. (1993)

Baltic Sea


0.45 µm filtered

0.76 (0.45–1.1)

Stoeppler et al. (1986)

Coastal waters, Malaysia


0.45 µm filtered; 66% arsenate; 33% arsenite

0.95 (0.65–1.8)

Yusof et al. (1994)

Bohai Bay, China


39º10’–38º40’N; 117º37’–180º00’E

1.4 (0.56–2.1)

Tan et al. (1983)

a Mean and ranges of total As unless stated otherwise

NS = not stated

Table 6. Concentrations of As in estuarine waters


Sampling Period

Sampling details and/or species

Concentration (µg/litre)a


Tamar estuary, UK


Glass fibre filtered

2.7–8.8 (range)

Howard et al. (1988)

Rhone estuary, France


4–24% arsenite; surface

1.3–3.8 (range)

Seyler & Martin (1990)

Gironde estuary, France


4–14% arsenite; surface

0.7–2.5 (range)

Seyler & Martin (1990)

Loire estuary, France


4–25% arsenite; surface

1.5–3.0 (range)

Seyler & Martin (1990)

Schelde estuary, Belgium


2–16% arsenite

1.8–4.9 (range)

Andreae & Andreae (1989)

Huang He river estuary, China


dissolved As, surface water

Total = 3.6 (2.8–4.3)

Li et al. (1989)

Organic = 2.3 (1.3–2.9)

Inorganic. = 1.4 (0.7–2.3)

Arsenite = 0.5 (0.3–0.8)

Arsenate = 0.8 (0.2–1.4)

a Mean and ranges of total As unless stated otherwise

NS = not stated

Table 7. Concentrations of As in surface freshwaters


Sampling period

Sampling details and/or As source

Concentration (µg/litre)a


Brazos river, Texas, USA


0.2 µm filtered, arsenite


Chakraborti et al. (1986)

Madison river, Montana, USA




Sonderegger & Ohguchi (1988)

Finfeather lake, Texas, USA


near manufacturing plant for As-based cotton defoliants

7900 (6000–8600)

Crearley (1973)

Municipal lake, Texas, USA


as above

3200 (1700–4400)

Crearley (1973)

Maurice river, NJ, USA


upstream of pesticide plant

3.3 (1.05–4.4)

Faust et al. (1987a)


0.6 km downstream

2222 (1320–4160)

Faust et al. (1987a)


4.2 km downstream

266 (118–578)

Faust et al. (1987a)

Union lake, NJ, USA


14–17 km downstream

86.1 (27.1–267)

Faust et al. (1987a)

Bowron lake, British Columbia, Canada


reference lake; no mining activity

0.26 (<0.2–0.42)

Azcue et al. (1994a)

Lake water, British Columbia, Canada


near abandoned gold mine

0.25 (< 0.2–0.3)

Azcue et al. (1994a)

Asososca lake, Nicaragua


volcanic crater; includes surface, intermediate and bottom samples

5.9 (0.85–15.8)

Cruz et al. (1994)

Moira lake, Ontario, Canada


past mining activity; 15% particle sorbed

43 (4–94)

Diamond (1995)

Lakes, Northwest Territories, Canada


gold mining activity

700–5500 (range)

Wagemann et al. (1978)

Subarctic lakes, Northwest Territories, Canada


gold mining activity

270 (64–530)

Bright et al. (1996)

Yangtze river (source area), China


filtered water (< 0.45 µm)

3.1 (0.1–28.3)

Zhang & Zhou (1992)

Antofagasta, Chile


Toconce river, Andes mountains

< 800

Borgono et al. (1977)

Mutare river, Zimbabwe


near gold/As mine dumps

(range of means)

Jonnalagadda &

Odzi river, Zimbabwe


2.2 km downstream from gold/As mine dumps (after confluence with Mutare river)

(range of means)

Nenzou (1996b)

Xolotlan lake, Nicaragua


volcanic crater; range of means

(range of means)

Lacayo et al. (1992)

Waikato river, New Zealand


volcanic source

32.1 (28.4–35.8)

McLaren & Kim (1995)

Lake water, Lapland, Finland


0.1 m below surface

0.17 (median)

Mannio et al. (1995)

Nakhon Si Thammarat province, Thailand


mining activity

217.5 (4.8–583)

Williams et al. (1996)

a Mean and ranges of total As unless stated otherwise

NS, not stated

5.1.4 Groundwater
Levels of arsenic in groundwater are summarized in Table 8. Arsenic levels in groundwater average about 1–2 µg/litre, except in areas with volcanic rock and sulfide mineral deposits where arsenic levels can range up to 3400 µg/litre (Page, 1981; Welch et al., 1988; Robertson, 1989). In some mining areas arsenic concentrations of up to 48 mg/litre have been reported (Welch et al., 1988). Korte & Fernando (1991) reported that arsenic levels in arsenic-contaminated water supply wells in southern Iowa and western Missouri (USA) ranged from 34 to 490 µg/litre. The authors state that the arsenic appears to be of natural origin. Similarly, Matisoff et al. (1982) found no evidence for an anthropogenic source contributing to elevated groundwater levels of arsenic (< 1 to 100 µg/litre) in north-eastern Ohio (USA). Arsenic levels in groundwater were found to exceed 10 µg/litre in 5.6–9.5% of samples collected in Germany during the period 1992–1994 (Umweltbundesamt, 1997). Varsanyi (1989) found arsenic concentrations in deep groundwater in Hungary to range from 1 to 174 µg/litre with an average value of 68 µg/litre. High arsenic levels originating from arsenic-rich bedrock were found in drilled wells in south-west Finland, with concentrations ranging from 17 to 980 µg/litre (Kurttio et al., 1998). Del Razo et al. (1990) monitored groundwater in the Lagunera region of northern Mexico. Total arsenic concentrations ranged from 8 to 624 µg/litre with over 50% of samples > 50 µg/litre. The predominant arsenic species in 93% of samples was arsenate, although in 36% of samples 20–50% arsenite was found. Chen et al. (1994) report that arsenic levels in the groundwater of south-west Taiwan contained mean dissolved arsenic levels of 671 µg/litre. Arsenic levels in the well-waters of Hsinchu (Taiwan) were less than 0.7 µg/litre.

Table 8. Concentrations of As in groundwater


Sampling period

As source

Concentration (µg/litre)a




deep groundwater

68 (1–174)

Varsanyi (1989)

South-west Finland


well-waters; natural origin

17–980 (range)

Kurttio et al. (1998)

New Jersey, USA



1 (median)

Page (1981)

1160 (maximum)

Page (1981)

Western USA


geochemical environments

48 000 (maximum)

Welch et al. (1988)

South-west USA


alluvial aquifers

16–62 (range of means)

Robertson (1989)

Southern Iowa and western Missouri, USA


natural origin

34–490 (range)

Korte & Fernando (1991)

North-eastern Ohio, USA


natural origin

< 1–100 (range)

Matisoff et al. (1982)

Lagunera region, northern Mexico



8–624 (range)

Del Razo et al. (1990)

Cordoba, Argentina

> 100

Astolfi et al. (1981)


470–770 (range)

De Sastre et al. (1992)

Pampa, Cordoba, Argentina


2–15 m,

100–3810 (range)

Nicolli et al. (1989)

Kuitun-Usum, Xinjiang, PR China



850 (maximum)

Wang et al. (1993)

Hsinchu, Taiwan



< 0.7

Chen et al. (1994)

West Bengal, India


As-rich sediment

(range of means)

Chatterjee et al. (1995)

3700 (maximum)

Chatterjee et al. (1995)

Calcutta, India


near pesticide production plant

< 50–23 080 (range)

Chakraborti et al. (1998)




< 10–> 1000 (range)

Dhar et al. (1997)

Nakhon Si Thammarat Province, Thailand


shallow (alluvial) ground-water; mining activity

503.5 (1.25–5114)

Williams et al. (1996)


deep groundwater; mining activity

95.2 (1.25–1032)

Williams et al. (1996)

a Mean and ranges of total As unless stated otherwise

NS = not stated

Arsenic contamination of groundwater from arsenic-rich sediment has been reported in both India and Bangladesh. Chatterjee et al. (1995) analysed groundwater from six districts of West Bengal (India). Mean total arsenic levels ranged from 193 to 737 µg/litre with a maximum value of 3700 µg/litre. Mean arsenite levels in the groundwater were around 50% of the total arsenic. Mandal et al. (1996) reported that 44% of groundwater samples collected in West Bengal (India) up to January 1996 contained total arsenic levels > 50 µg/litre. Dhar et al. (1997) found that 38% of groundwater samples collected from 27 districts of Bangladesh contained total arsenic levels > 50 µg/litre.

During 1990 and 1991 Chatterjee et al. (1993) sampled groundwater in the vicinity of a chemical plant in Calcutta, India, which had produced the insecticide Paris green (acetocopper arsenite) for 20 years. Groundwater contained total arsenic levels ranging from < 0.05 to 58 mg/litre; the highest total arsenic level included 75% arsenite.

5.1.5 Sediment
Arsenic concentrations in sediments are summarized in Table 9. Sediments in aquatic systems often have higher arsenic concentrations than those of the water (Welch et al., 1988). Most sediment arsenic concentrations reported for rivers, lakes and streams in the USA range from 0.1 to 4000 mg/kg, with higher levels occurring in areas of contamination (Welch et al., 1988). Arsenic concentrations of < 10 000 mg/kg (dry weight) were found in surface sediments near a copper smelter (Crecelius et al., 1975). Sediment arsenic concentrations of < 3500 mg/kg were reported for lakes in the Northwest Territory (Canada) which had received past inputs from gold-mining activity (Wagemann et al., 1978). Mean total arsenic concentrations of 500 mg/kg (dry weight) were measured in sediment near a pesticide plant and at a lake 14–17 km downstream mean concentrations of almost 3000 mg/kg had accumulated (Faust et al., 1987a). Arsenate was the predominant arsenic species, with inorganic arsenic amounting to 70–90% of the total arsenic measured (Faust et al., 1983). Bright et al. (1996) found total arsenic concentrations ranging from 1043 to 3090 mg/kg in the top 10 cm of sediment from subarctic lakes contaminated by gold-mining activity. Total dissolved arsenic levels in porewater ranged from 800 to 5170 µg/litre (0.7% organic arsenic). Ebdon et al. (1987) reported that methylated arsenic species represented 1–4% of the total arsenic in sediment porewater from the Tamar estuary, south-west England (United Kingdom). Similar findings were reported by de Bettencourt (1988) for the Tagus Estuary (Portugal).

Table 9. Concentrations of As in sediment


Sampling period

Sampling details and/or As source

(mg/kg dry weight)a



UK estuaries


100 µm sieved

2–94 (range)

Langston (1980)

Estuaries, south-west England, UK


past mining activity

7–2500 (range)

Langston (1980)

Tamar estuary, UK


inorganic As


Howard et al. (1988)

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