SE 9307/97 KEM Eseoti

No 2/98

received JUN 2 9 m O&T I Arsenic in the aquatic environment - speciation and biological effects

Exemption Substances Project Lars Landner Swedish Environmental Group

DISTRIBUTION OF THIS DOCUMENT IS UNLIMITED FOREIGN SALES PROHIBITED “PL-

THE SWEDISH NATIONAL CHEMICALS INSPECTORATE DISCLAIMER

Portions of this document may be illegible electronic image products. Images are produced from the best available original document. Arsenic in the aquatic environment - speciation and biological effects ISSN: 0284-1185 Order No. 360 599 Printed by: Printgraf, Stockholm, March 1998 Publisher: Swedish National Chemicals Inspectorate© Order address: P.O. Box 1384, S-171 27 Solna, Sweden Telefax 46 8-735 52 29, e-mail [email protected] Preface This report is a contribution to EC Commission's undertaking to review existing EC provisions on the substances for which Sweden has been granted transitional provisions*. The provisions imply that Sweden may maintain more stringent regulations on four substances until the end of 1998. The report is one in a series providing further facts to the Commission and Member states in the review process, and is produced within the Exemption Substances Project at the National Chemicals Inspectorate. Exempted substances are arsenic and organotin compounds (Directive 89/677/EEC), pentachlorophenol (91/171/EEC), cadmium (91/33 8/EEC), and fertilizers with regard to their cadmium content (76/1 16/EEC). The present report deals with speciation and biological effects of arsenic in three types of aquatic environments - marin water, estaurin or brackish-water and freshwater. The similarity between arsenate and phosphate and the interference in phosphorylation reactions is discussed. It is clear that in Scandinavian inland waters the concentration of phosphorus is on average lower than in most inland waters in continental Europe. However, in most inland waters phosphorus is the limiting actor for phytoplancton development and eutrofication, which means that there is a clear risk for detrimental effects in the great majority of inland waters, also eutrophic waters. The author alone is responsible for the contents of the report.

Solna, Lars Gustafsson, Project Manager Exemption Substances Project *Act concerning the conditions of accession and the adjustments to the Treaties on which the Union is founded (OJ 94/C241/08) Table of Contents SUMMARY 1 SAMMANFATTNING 6 I. INTRODUCTION 11 II. SPECIATION OF ARSENIC IN THE AQUATIC ENVIRONMENT 13 1. Marine Environment 13 1.1 Inorganic arsenic species 13 1.2 Organic arsenic species 16 1.3 Biotransformation of arsenic - metabolic cycle 21 1.4 Sinks, mobilization and bioavailability of arsenic species 28 1.5 Uptake in organisms and bioaccumulation 30 1.6 Natural factors affecting speciation and bioavailability 30

2. Estuarine or Brackish Water Environment 32 2.1 Arsenic levels and speciation in water and sediments 32 2.2 Arsenic levels and speciation in biota 39 2.3 Biotransformation of arsenic 41 2.4 Sinks, mobilization and bioavailability of arsenic species 43 2.5 Uptake in organisms and bioaccumulation 46 2.6 Natural factors affecting speciation and bioavailability 51

3. Freshwater Environment(Lakes and Rivers) 53 3.1 Arsenic speciation in water and sediments 53 3.2 Arsenic levels and speciation in biota 57 3.3 Biotransformation of arsenic 58 3.4 Sinks, mobilization and bioavailability of arsenic species 60 3.5 Uptake in organisms and bioaccumulation 64 3.6 Natural factors affecting speciation and bioavailability 67

4. WaterQuality Characteristics in inland European Water Bodies of Relevance for Arsenic Speciation 68 III. EFFECTS OF VARIOUS ARSENIC SPECIES IN THE AQUATIC ENVIRONMENT 71

1. Marine Environment 71 1.1 Effects on microorganisms 71 1.2 Effects on macroalgae and other plants 76 1.3 Effects on invertebrates 76 1.4 Effects on marine fish 77 1.5 Effects at the ecosystem level 78 1.6 Arsenic criteria for protection of marine life 80

2. Estaurine or Brackish Water Environment 81 2.1 Effects on microorganisms 81 2.2 Effects on macroalgae and other plants 82 2.3 Effects on invertebrates 84 2.4 Effects on estuarine fish 86 2.5 Effects at the ecosystem level 86

3. Freshwater Environment (Lakes and Rivers) 87 3.1 Effects on microorganisms 87 3.2 Effects on aquatic plants 90 3.3 Effects on invertebrates 91 3.4 Effects on freshwater fish 92 3.5 Effects at the ecosystem level 93 3.6 Arsenic criteria for protection of freshwater aquatic life 93

4. Mechanisms of Toxicity - Interaction with Phosphorus 94

IV. LITERATURE REFERENCES 96 SUMMARY

1. Predominant Arsenic Species in Aquatic Environments Arsenate is the predominant form of arsenic in marine waters, because it is the thermo-dynamically most stable form in oxic waters. In addition to arsenate, three other arsenic species are commonly encountered in dissolved form in marine as well as brackish water: the reduced incorganic form, arsenite, and two methylated species, methylarsonic acid (MAA) and dimethylarsinic acid (DMAA). Other, more complex organic forms of arsenic may be present as well, but their concentrations in water are low, and they are not generally detected using conventional analytical methods.

In reducing environments, such as anoxic water or sediments, arsenite is the more stable inorganic arsenic species, but it is also found in aerobic systems. However, the oxidation of arsenite to arsenate, in the presence of oxygen, is usually rapid.

The rate of uptake of arsenate into autotrophs (phytoplankton, periphyton and macroalgae) in the marine environment is relatively high. The uptake is followed by a rapid transformation of arsenate to low- molecular, methylated arsenic species, such as MAA and DMAA, which are excreted back to the water phase, or - mainly - to complex organic arsenic species, arseno-sugars (dimethylarsinylribosides and trimethylarsonioribosides), which are stored inside the aquatic plants. The arsenosugars are then transformed in several steps to arsenocholine and arsenobetaine. Arsenobetaine is the predominant arsenic compound found in marine animals, including fish.

The biogeochemistry of arsenic has been thoroughly studied and described in the marine environment, and most of our present knowledge about the metabolic transformations of arsenic in aquatic ecosystems has been obtained from marine studies. However, also in estuarine or brackish-water ecosystems, many similar studies have been conducted and it is now well established that the principle pattern of arsenic turnover is similar in these aquatic environments. Due to the usually higher aquatic productivity of estuaries and land-locked seas, as compared to the open ocean, the production and release by algae of arsenite and DMAA is often much higher than in the open sea.

1 While arsenic speciation has been thoroughly studied in marine ecosystems, and is also relatively well investigated in brackish-water ecosystems, much less data on the exact speciation and transformation of arsenic are available from freshwater ecosystems. In lakes and rivers, it appears that pH and the content of humic substances, as well as the existence of hydrous iron, manganese and aluminium oxides, play a relatively more important role for the form of occurrence and bioavail­ ability of arsenic. For example, in acid lakes with a high content of humic material, arsenic will form stable complexes with humic acid, which tend to increase the mobility of arsenic, but without increasing its bioavailability.

Biomethylation of arsenic by freshwater organisms has been experimentally proven, but the exact nature and the chemical structure of the organic arsenic compounds present in the living cells of freshwater organisms still have to be revealed. However, at least traces of the comparatively non-toxic compound arsenobetaine have been positively identified in freshwater fish. It has also been demonstrated that the arsenic compounds that predominate in freshwater fish are trimethylated, i.e. they have the same number of methyl groups as arsenobetaine.

At the present state of knowledge, it might be concluded that there seems to be a greater variation between freshwater biological species in their capacity to methylate arsenic and to synthesize complex organic arsenic compounds, as compared to marine species.

2. Ecotoxicology of Arsenic in the Aquatic Environment Elevated concentrations of arsenic in surface waters, due to geological anomalies or to pollution from industrial or mining activities, will generate a variety of complex interactions with biotic and abiotic factors which may affect the transport, bioavailability, metabolism and ecotoxicity of arsenic. Consequently, the ecotoxicity and hence the environmental risk of arsenic may be extremely variable depending on the natural factors existing in the water body being exposed to the elevated arsenic concentrations.

As mentioned in the previous sections, the speciation of arsenic and the conversions between the different chemical forms are to a large extent determined by the biotic components of the ecosystem. Thus, the arsenic being released (e g from rocks and minerals) to the aquatic system will change its chemical form - and thereby its toxic properties - along its

2 route through the ecosystem in close relationship with the activity and the metabolic competence of the organisms through which it passes. Furthermore, one of the major inorganic arsenic species in water, arsenate, is chemically so similar to phosphate that organisms may be protected by phosphate from the toxicity of arsenate. It is well-known that phosphate usually exerts an overall control of the production of freshwater ecosystems, where it usually is the limiting nutrient. Consequently, in a natural, oligotrophic freshwater ecosystem with low phosphate levels, the introduction of even low concentrations of arsenate will produce severe detrimental effects, while a naturally eutrophic system with moderate to high phosphate levels is much more tolerant to arsenate exposure.

From the above it is clear that aquatic organism such as algae (phytoplankton, periphyton and macro-algae) which rely upon dissolved phosphate at low concentrations in the water as their main source of phosphorus are particularly vulnerable to arsenate. Interference with phosphate uptake at the cell surface and in phosphorylation reactions within the cell is generally thought to be the main mode of action of arsenate. Uncoupling of oxidative and photo-phosphorylation will occur, since in the presence of arsenate unstable ADP-arsenate is formed, thus regenerating ADP and preventing the formation of ATP. Once within the algal cell, arsenate seems to be transformed to arsenite, MAA, DMAA, arsenosugars and arsenolipids. Whether these compounds are the outcome of a toxic interference with metabolism or products of a detoxification mechanism has not yet been fully clarified. At least arsenite and MAA are (or can be) reduced to sulfhydryl active agents capable of enzyme inhibition at the biochemical level.

Based on a great amount of experimental exposures of selected aquatic organisms and communities to arsenate in microcosms and mesocosms with the aim of providing data necessary for an advance hazard assessment of arsenic in the aquatic environment, a set of "Lowest Observed Effect Levels" (LOEL) were determined (Table IV:2.1).

It is obvious that, at low phosphorus levels in the water (and in the cells), arsenate is by far the most ecotoxic among the arsenic species. The often held view that arsenite is the most ecotoxic arsenic species must therefore be strongly questioned, at least when it comes to toxicity to algae. Although the microalgae were the most sensitive organisms (among all organisms investigated) to arsenate exposure, it should be noted that the macroalga Fucus vesiculosus was severely affected at an

3 arsenate concentration of 8 pg As/1, and the freshwater macrophyte Lobelia dortmanna was inhibited at 5 pg As/1, but only when the phosphate concentration was very low.

Generally, both the invertebrate and the vertebrate aquatic fauna seems to be quite insensitive towards direct effects of arsenic, at least in comparison to green plants. Available toxicity data from short- and long­ term experiments, where animals have been exposed to arsenic compounds dissolved in the water, do indicate that concentrations have to exceed at least 80 pg As/1 before any direct toxic responses can be observed. The lowest reported effect concentrations were in the range 86-100 pg As/1, and the three species, arsenite, arsenate and MAA all fell in this range. However, indirect effects on herbivores and higher trophic levels have been repeatedly reported as a result of e.g. restructuring of the primary producer community caused by low levels of arsenate exposure.

System Environment Parameter LOEL, figAs/l affected Phytoplankton Marine, P-limited Cell growth 2,3 Phytoplankton Freshwater, P- Total cell volume 4.4 limited Macrophyte, Lobelia Freshwater Loss of leaves 5,0 Periphyton Brackish water, Photosynthesis 6,0 P-limited Macroalga, Fucus Brackish water, Production 8,0 P-limited Periphyton Marine, P-limited PICT 8-15 Periphyton Marine, P- Photosynth, 23 limited biomass Table IV: 2.1 Lowest Observed Effect Levels (LOEL) of arsenate in experimental chronic exposures of selected organisms or communities. The systems are arranged in order of increasing tolerance to arsenate.

The above evaluation resulted in the following conclusions:

* Arsenate is the ecotoxicologically most significant of the occurring arsenic species, due to its similarity with phosphate and its ability to interfere with phosphorylation reactions in energy metabolism.

* Algae and other green plants are the primary targets of arsenate exposure, both in marine and freshwater, phosphorus-limited

4 environments, where extremely low levels of arsenate exposure may cause detrimental effects.

* In Scandinavian inland waters, the concentration of phosphorus in the water is on average lower than in most inland waters in continental Europe. Even in large lakes e.g. in Sweden and Norway, the mean concentration of total phosphorus is <10 pg/1. and in small forest lakes, it is usually <5 pg P/I, while the average level in European lakes is reported to be 40-50 pg P/1.

* The major effect of long-term arsenate exposure at the algal community level is a fundamental restructuring of the community such that sensitive algal species disappear, while more tolerant species remain. Consequently, the main effect can be expressed as a reduction in biological (and genetic) diversity. This restructuring of algal communities may have cascading effects at higher trophic levels, such as impairment of the fecundity and the reproduction of herbivore communities.

* Arsenic is not biomagnified in the food chain, but occurs in many different chemical forms that are not fully characterized from an ecotoxicological point of view. However, the predominant arsenic species in marine and brackish-water animals, including fish (and possibly also in freshwater fish) is arsenobetaine, which has been shown to have a relatively low toxicity, both to aquatic organisms and to higher predators.

* The sediments act as a sinks of arsenic, and benthic organisms can be exposed to high arsenic loads in contaminated environments. Reduced species of arsenic are likely to affect biota in such environments, but high levels of iron, manganese and humic material in the sediments tend to reduce the availability and toxicity of arsenic to sediment­ dwelling organisms. SAMMANFATTNING

1. Dominerande forekomstformer av arsenik i akvatisk miljo Arsenat ar den dominerande forekomstformen av arsenik i havsvatten, beroende pa att det ar den termodynamiskt mest stabila formen i syrerika vatten. Forutom arsenat patraffas vanligen ytterligare tre former av lost arsenik i havs- och brackvatten: den reducerade, oorganiska foreningen arsenit och tva metylerade foreningar, metylarsonsyra (MAA) och dimetylarsinsyra (DMAA). Andra, mera sammansatta organiska arsenikforeningar kan aven forekomma, men deras halter i vatten ar laga och de detekteras vanligen inte vid anvandning av konventionella analysmetoder.

I reducerande miljo, t ex i syrefattiga vatten och i sediment, ar arsenit den mest stabila oorganiska arsenikformen, och denna form kan aven patraffas i aeroba system. Dock sker oxidationen av arsenit till arsenat vanligen snabbt i narvaro av syre.

Upptagshastigheten av arsenat i autotrofa organismer (vaxtplankton, perifyton och makroalger) i den marina miljon ar relativt hog. Upptagningen i vaxtcellema foljs av en snabb omvandling av arsenat till lagmolekylara, metylerade arsenikformer (MAA och/eller DMAA), vilka ater utsondras till vattenfasen. Den huvudsakliga omvandlingsreaktionen resulterar dock i komplexa, organiska arsenikformer, arseniksocker (dimetylarsinylribosider och trimetylarso-nioribosider), vilka upplagras i vattenvaxtemas celler. Arseniksockerforeningama omvandlas sedan vidare i ekosystemet till bl a arsenokolin och arsenobetain. Arsenobetain ar den dominerande arsenikforeningen som pavisats i marina djur, inklusive fiskar.

Arsenikens biogeokemi har grundligt studerats och beskrivits i den marina miljon och merparten av den befintliga kunskapen om arsenikens metaboliska omvandlingar i akvatiska ekosystem har forvarvats genom studier i marin miljo. Dock har manga liknande studier bedrivits i estuarie- eller brackvatten-ekosystem och det ar numera val belagt att det huvudsakliga monstret i arsenikens naturliga omsattning ar likartat i dessa vattenmiljoer i forhallande till rent marina miljoer. Emellertid leder den vanligen hogre akvatiska produktionen i estuarier och kusthav till att savSl algemas bildning som deras utsondring av arsenit och DMAA ofta ar betydligt hogre i dessa miljoer an i det oppna havet.

6 Medan arsenikens forekomstformer och omvandlingar ar grundligt studerade i marina eko-system, och aven relativt val undersokta i brackvatten-ekosystem, sa ar tillgangen pa data rorande arsenikens exakta forekomstformer och dess omvandlingar i sotvattensystem betydligt mindre. I sjoar och vattendrag verkar pH-vardet och innehallet av humusamnen, liksom fbrekomsten av hydratiserade jam-, mangan- och aluminiumoxider, spela en relativt sett mera betydelsefull roll for arsenikens forekomstform och biotillganglighet. Till exempel i sura sjoar med hog humushalt bildar arsenik stabila komplex med humussyra, vilket tenderar till att oka arsenikens mobilitet, dock utan att oka dess biotillganglighet.

Biometylering av arsenik har experimentellt visats kunna utforas av sotvattenorganismer, men den exakta arten och den kemiska strukturen av de organiska arsenikforeningar som fore-kommer i cellema hos sotvattenorganismer aterstar att klarlagga. Dock har atminstone spar av den relativt otoxiska foreningen arsenobetain identifierats i fisk fran sotvattenmiljo. Det har ocksa visats att de arsenikforeningar som dominerar i flera sotvattenfiskar ar trimetylerade, d v s de har samma antal metylgruppersom arsenobetain.

Pa nuvarande kunskapsniva kan slutsatsen dras att det verkar finnas en storre variation mellan olika biologiska arter i sotvatten vad galler deras kapacitet att metylera arsenik och att syntetisera komplexa organiska arsenikforeningar, jamfort med vad som galler for marina arter.

2. Arsenikens ekotoxikologi i akvatisk miljo Forhojda koncentrationer av arsenik i sjoar, vattendrag och kustvatten, till fdljd av geologiska anomalier eller kontaminering fran industri- eller gruvverksamhet, kan ge upphov till en mangfald komplexa samspel med biotiska och abiotiska faktorer som kan paverka arsenikens transport, biotillganglighet, metabolism och ekotoxicitet. Foljaktligen kan arsenikens ekotoxicitet och darmed dess miljdrisk vara extremt variabel beroende pa de naturliga faktorer som specifikt rader i den vattenmiljd dar de forhojda arsenikkoncentrationema upptrader.

Som namnts tidigare, bestams arsenikens forekomstformer, liksom omvandlingama mellan dessa, i hog grad av de biotiska komponentema i ekosystemet. Saledes kommer den arsenik som introduceras (t ex fran berggrunden) till det akvatiska systemet att andra sin kemiska form - och darmed sina toxiska egenskaper - langs sin vag genom ekosystemet, i

7 hog grad beroende pa aktiviteten och den metaboliska formagan hos de organismer genom vilka amnet passerar. Dessutom ar den viktigaste oorganiska forekomstformen i vatten, arsenat, kemiskt sa lik fosfat att organismema kan skyddas av fosfat fran arsenats toxicitet. Det ar valkant att fosfat vanligen utovar en overgripande kontroll av produktionen i sotvatten-ekosystem, dar det ofta utgor det begransande makronarings- amnet. I ett naturligt, oligotroft sotvatten-ekosystem med laga fosfat- nivaer kan foljaktligen tillforsel redan av laga arsenatkoncentrationer resultera i kraftiga negativa effekter, medan naturligt eutrofa system med moderata till hdga fosfatnivaer vanligen ar mycket mera okansliga for arsenat-exponering.

Utiffan det ovanstaende ar det uppenbart att akvatiska organismer som alger (fytoplankton, perifyton och makroalger), vilka ar beroende av lost fosfat i laga koncentrationer i vattnet for sin huvudsakliga fosforfor- sorjning, ar sarskilt kansliga mot arsenat. Stoming av fosfatupptagningen vid cellvaggen och av fosforyleringsreaktionema inuti cellen anses vanligen vara arsenatens viktigaste verkningsmekanism. Frikoppling av den oxidativa fosforyleringen och/eller av foto-fosforyleringen kan aga rum, eftersom instabil ADP-arsenat bildas i narvaro av arsenat, vari- genom ADP regenereras och bildningen av ATP forhindras. Sedan arsenat tagits upp av algcellen, reduceras foreningen till arsenit och metyleras till MAA och/eller DMAA eller byggs in i mera komplexa molekyler som arseniksocker och arseniklipider. Huruvida dessa fbreningar ar resultatet av en toxisk paverkan pa cellmetabolismen eller produkter av en avgiftningsmekanism har annu inte blivit fullstandigt klarlagt. Atminstone arsenit och MAA kan reduceras till sulfhydryl- aktiva amnen som har formagan att biokemiskt inhibera enzymer.

Baserat pa ett stort antal experimentella studier, dar utvalda akvatiska organismer och samhallen har exponerats for arsenat i mikrokosmer och mesokosmer med syftet att generera data som ar nddvandiga for en avancerad riskanalys av arsenik i akvatisk miljo, har en lista over "Lagsta Observerade Effekt-Nivaer" tagits ffam (Tabell 1).

Det ar uppenbart att arsenat ar den klart mest ekotoxiska arsenikformen vid laga fosfornivaer i vattnet (och i cellema). Den ofta framforda asikten att arsenit ar den mest toxiska fbrekomst-formen av arsenik maste darfor starkt ifragasattas, atminstone nar det galler toxicitet mot alger. Fastan mikroalger visade sig vara de kansligaste organismema (bland alia testade organismer) mot arsenatexponering, bbr det noteras att makroalgen Fucus vesiculosus (blastang) paverkades kraftigt vid en

8 arsenatkoncentration om 8 pg As/1, och sotvattenmakrofyten Lobelia dortmanna (notblomster) inhiberades vid 5 pg As/1, men endast nar fosfatkoncentrationen var mycket lag.

Generellt sett forefaller bade de akvatiska ryggradslosa och ryggrads- djuren vara relativt okansliga mot arsenikexponering, atminstone i jamfbrelse med grona vaxter. Tillgangliga toxicitetsdata fran saval korttids- som langtidsexperiment, dar djuren bar exponerats for arsenik- foreningar losta i vattnet, indikerar att koncentrationema maste over- skrida atminstone 80 pg As/1 innan nagon direkt toxisk respons kan observeras. De lagsta rapporterade effektkoncentra-tionema lag i inter- vallet 86-100 pg As/1. Alla tre forekomstformema, arsenat, arsenit och MAA, foil inom detta koncentrationsintervall. Emellertid har indirekta effekter pa herbivorer och hogre trofiska nivaer flerfaldiga ganger rapporterats som ett resultat av t ex omstrukturering av primarproducent- samhallet, fbrorsakad av laga nivaer av arsenatexponering.

System Miljd Matparameter Effektniva, pgAs/l Fytoplankton Marin, P-begransad Celltillvaxt 2,3 Fytoplankton Sotvatten, P- Total cellvolym 4,4 begrdnsad Makrofyt, Lobelia Sotvatten Bladforlust 5,0 Perifyton Brackvatten, P- Fotosyntes 6,0 begrdnsad Makroalg, Fucus Brackvatten, P- Produktion 8,0 begransad Perifyton Marin, P-begransad PICT 8-15 Perifyton Marin, P-begransad Fotosyntes, 23 biomassa Tabell 1. Lagsta Observerade Effekt-Nivaer for arsenat vid experimented, kronisk exponering av ndgra utvalda organismer och samhaden. Malorganismerna dr listade i enlighet med okande tolerans mot arsenat.

(PICT = Pollution-Induced Community Tolerance)

9 Utvarderingen leder till foljande slutsatser:

* Arsenat ar den ekotoxikologiskt mest betydelsefulla av arsenikens forekomstformer i akvatisk miljd, pa grand av dess likhet med fosfat och dess formaga att interferera med fosforyleringsreaktioner i energimetabolismen.

* Alger och andra grona vaxter ar de primara malorganismema for arsenatexponering, bade i marina och sdtvattenmiljoer som ar fosforbegransade, dar mycket laga nivaer av arsenatexponering kan framkalla skadliga effekter.

* I Skandinaviska inlandsvatten ar koncentrationen av fosfor i vattnet i genomsnitt lagre an i de fiesta inlandsvatten i kontinentala Europa. Aven i de stora sjoama i t ex Sverige och Norge ar medelkon- centrationen av totalfosfor <10 pg/1, och i sma skogssjoar ar den vanligen <5 pg P/1, medan medelnivan i sjoar i Europa ligger omkring 40-50 pg P/1.

* Den huvudsakliga effekten av langtidsexponering for arsenat pa algsamhallesnivan ax en fundamental omstrakturering av samhallet, sa att kansliga algarter forsvinner och endast toleranta arter kvarblir. Saledes kan huvudeffekten uttryckas som en reduktion i biologisk (och genetisk) diversitet. Omstruktureringen av algsamhallena kan ha foljdeffekter pa hdgre trofinivaer, sasom nedsattning av fekunditeten och fortplantningen hos vissa arter i vaxtatarsamhallet, pa grand av "felaktig" naringstatus hos kvarvarande algarter.

* Arsenik biomagnifieras inte i naringskedjan, men forekommer i manga olika kemiska former, vilka inte alia ar fullt karakteriserade fran ekotoxisk synpunkt. Dock ar den dominerande organiska arsenikformen i marina och brackvattendjur, och troligen ocksa i sotvattenfisk, arsenobetain, som bar visats besitta en relativt lag toxicitet, bade gentemot akvatiska organismer och konsumenter pa hdgre trofisk niva.

* Sedimenten fungerar som en "sink" for arsenik, och bentiska organismer kan exponeras for hoga arseniknivaer i kontaminerade miljoer. Reducerade arsenikformer kan paverka biota i sedimenten, men riklig forekomst av jam, mangan och humusamnen i dessa miljoer tenderar att reducera tillgangligheten och toxiciteten av arsenik gentemot sedimentlevande organismer.

10 I. INTRODUCTION

A very great number of review articles and books have been devoted to arsenic in the environment and, particularly, to discussions about arsenic speciation and transformation in the environment. Most of these reviews were published in the 1970s and 1980s, and are listed by Maeda (1994). As pointed out by this author, there seems to a significant difference in the total concentration of arsenic, between terrestrial and freshwater organisms on one hand and marine organisms on the other. Terrestrial organisms in uncontaminated environments rarely contain more than 1 pg As/g dry weight (d.w.), whereas marine organisms contain from several pg As/g to more than 100 pg As/g (Lunde, 1977, Francesconi and Edmonds, 1993). Also in unpolluted rivers, ponds and lakes, the arsenic concentrations in organisms are generally low, but are greatly affected by local conditions.

The first reports of arsenic in marine organisms were published already in the beginning of the century (e g Bertrand, 1902), and later work (e g Jones, 1922; Cox, 1925; Chapman, 1926) established that marine organisms naturally contain high levels of arsenic. These observations prompted an upsurge of research on the biogeochemical cycling of arsenic in the marine environment, in particular after the first identification, in 1977, of the precise chemical form of one of the organoarsenic compounds found in a marine animal. It was in this year, Edmonds et al. (1977) identified arsenobetaine in the western rock lobster (Panulirus cygnus).

Subsequent work has identified the presence of a large array of other organic arsenic compounds in marine organisms, and the complex chain of biotransformations of arsenic as well as its complete metabolic cycle in marine ecosystems has been fairly well described. However, relatively few studies have been directed to the full identification of chemical species of arsenic in freshwater organisms and the arsenic transforma­ tions along the freshwater food chain have only rarely been reported (Maeda, 1994). The result is that, still today, arsenic compounds in the marine environment are described in many books and reviews, but those in the freshwater environment are usually reported without any details on the exact identity of the chemical forms, and - even so - they are discussed only in few publications. Also with regard to the brackish- water or estaurine aquatic environment, the existing knowledge about arsenic speciation and biotransformation in various organisms is less

11 developed than for the marine ecosystem, although more data are available from estuarine waters than from inland waters.

This general lack of comprehensive information on arsenic speciation in freshwater environ-ments makes the assessment of biological effects more difficult than in the case of marine environments. It is evident that a correct evaluation of the environmental risk associated with arsenic exposure requires good knowledge about the precise chemical form of arsenic to which different organisms are exposed during different stages of their life cycles. Fortunately, a relatively large number of studies have been devoted to the determination of the dissolved arsenic species, occurring in the water column, both in brackish-water and in freshwater. This allows at least an assessment of the biological effects caused by direct exposure via the water to the different, soluble arsenic forms.

In preparing the present overview of arsenic species and effects in the aquatic environment, it was considered most logical to start the presentation of the natural cycling of arsenic in the field where the most comprehensive data base exists. Therefore, a description of the speciation and biotransformations of arsenic in the marine environment .will be used as a background for the subsequent discussion of arsenic in brackish-water and freshwater environments. It is obvious that human impact, including discharges of arsenic, may be greatest in the last mentioned environments. Here, the human activities are causing the strongest increases in arsenic concentrations over the natural background levels, and thereby the highest exposure of organisms to potentially toxic arsenic species. Thus, estuaries, land-locked seas, lakes and rivers are the aquatic environments where anthropogenic sources of arsenic may cause more important ecological impacts than in the marine ecosystem.

12 II. SPECIATION OF ARSENIC IN

THE AQUATIC ENVIRONMENT

1. Marine Environment

1.1 Inorganic arsenicspecies 1.1.1 Marine waters The concentration of total arsenic in the world's seas is fairly uniform, about 2 pg/1 (Andreae, 1978). The concentration range in the Atlantic Ocean has been given as 1.0 - 1.5 pg/1 (Sanders, 1980) and in the English Channel as 2 - 4 pg/1 (Woolson, 1975), while the global mean concentration in surface sea water has been suggested at 1.6 pg/1 (Burton and Statham, 1982).

Arsenate, As(V), is the predominant form of arsenic in seawater, making up about 80% of the total amount (Waslenchuk, 1978). However, arsenite, As(III), occurs at concentrations greater than those expected from purely thermodynamic considerations (Gohda, 1975). The explanation for this apparent anomaly seems to be that arsenate is reduced by marine bacteria (Johnson, 1972) and marine phytoplankton (Johnson and Burke, 1978). The concentration of arsenite is usually higher in the euphotic zone than below, which further supports that algae are important in the production of this species.

1.1.2 Marine sediments / interstitial waters While the mean arsenic content in the continental crust has been reported at 1.5 - 2 pg/g (NAS, 1977) and 3 pg/g (Cullen and Reimer, 1989), respectively, the average level of arsenic in deep sea sediments is much higher. The average for world oceans is about 40 pg/g (Bostrdm and Valdes, 1969). Arsenic levels in sediments from coastal regions and estuaries are generally lower than those from the deep sea, and reported values from uncontaminated coastal regions range from 3 to 15 pg/g (Kennedy, 1976).

Total arsenic concentrations have been found to be lower in the interstitial water than in the overlying water, suggesting removal of arsenic onto the solid phases by adsorption or coprecipitation (Andreae, 1979). Both forms of inorganic arsenic were found in the inter-stitial water, and the proportion of arsenite was always higher than that in

13 seawater, reflecting the different redox potentials of the two environments (Francesconi and Edmonds, 1994).

1.1.3 Marine biota Inorganic arsenic usually constitutes less than 10% of the total arsenic in macroalgae, although a few exceptions to this rule have been reported: Sargassum muticum contained 38% (Whyte and Englar, 1983) and Hizikia fusiforme some 50-60% of inorganic arsenic (Shinagawa et al., 1983). The concentration of total arsenic in the former species was reported to be in the range 40-90 pg/g d.w..

No good data are available on the natural content of inorganic arsenic in unicellular algae, but it may be assumed that the major part of arsenic in these microalgae are organic arsenic compounds, in essence the same as those formed by macroalgae (Francesconi and Edmonds, 1994), see section 1.2.4.

The arsenic content of marine animals shows considerable variability. Among the highest recorded values from uncontaminated areas were >100 pg/g wet weight (w.w.) in plaice (Pleuronectes platessa) and the gastropod Reishia bronni (Luten et al., 1982; Shiomi et al., 1984), and as much as >2000 pg/g d.w. in the polychaete Tharyx marioni (Gibbs et al., 1983). The dominant part of arsenic in soft tissues of marine animals has been demonstrated to be in an organic form (see section 1.2.3). In fact, various authors have shown that inorganic arsenic constitutes a very small percentage, <2%, of the total arsenic in marine animals from uncontaminated waters (see e g Maher, 1983).

In a recent re-evaluation of virtually all studies that have specifically analysed the fraction of inorganic arsenic in marine animals (Edmonds and Francesconi, 1993), it was concluded that the previous assumption that inorganic arsenic constitutes 2-10% of the total arsenic in seafood (GESAMP, 1986; Friberg, 1988) is not justified, see Figure 11:1.1. The regression indicates that the proportion of inorganic arsenic in marine animals falls from around 1% at very low total arsenic concentrations to about 0.5% at total arsenic levels of around 20 pg/g.

However, in shells of marine animals, the inorganic forms seem to predominate. For example, in the shells of four species of gastropods, total arsenic levels of 1.8 - 16 pg/g were recorded, while the shells of four bivalve species generally had lower levels. Both arsenate (66-92%

14 of total) and arsenite were present, and all samples contained low levels of methyl- or dimethylated arsenic compounds (Cullen et al., 1989).

GESAMP

TOTAL ARSENIC (mg/kg)

Figure 11:1.1 Inorganic arsenic versus total arsenic in marine animals. 1 = crustaceans, 2 = bivalve molluscs, 3 = gastropod molluscs, 4= cephalopod molluscs, 5=fish. The atypical dataofLunde (1973) are shown boxed. "GESAMP" denotes the relationship assumed by GESAMP (1986) and Friberg (1988). Linear regressions of inorganic on total arsenic have been fitted, (a) excluding, and (b) including the data of Lunde (1973). After Edmonds and Francesconi, (1993).

15 1.2 Organic arsenic species 1.2.1 Marine waters Usually, small quantities of methylarsonic acid (MAA) and dimethylarsinic acid (DMAA) are found in seawater, with somewhat higher levels in the euphotic zone than below (Braman and Foreback, 1973; Andreae, 1979). Apart from the four simple forms of arsenic (arsenate, arsenite, MAA and DMAA), no other dissolved arsenic species have been identified in marine waters (Francesconi and Edmonds, 1994). However, so-called "hidden arsenic" has been shown to make up an average of 25% of the total arsenic in coastal waters (Howard and Comber, 1989). The usual method for determining arsenic in seawater does not allow detection of arsonium compounds (R4As+) or most other organic arsenic compounds that occur in marine organisms. Therefore, it cannot be excluded that organic arsenic compounds, except MAA and DMAA, may be present at appreciable levels, but remain undetected by standard analytical methods (Francesconi and Edmonds, 1994).

The possibility that arsenobetaine, the predominant arsenic species in marine animals (see below), is also present in seawater has been discussed by the latter authors: On the basis of uptake experiments with mussels and lobsters, it was found that the mussles absorbed the arsenobetaine from the spiked seawater much more efficiently than did the lobsters, a result which is not in accordance with the situation found in the natural populations. Therefore, the experimental outcome was considered to indicate that direct uptake of arsenobetaine from seawater is not a significant contributor to the arsenobetaine burden of mussels or lobsters in their natural environment, and hence, arsenobetaine might not be an important constituent of natural seawater (Francesconi and Edmonds, 1994).

1.2.2 Marine sediments / interstitial waters The methylated arsenic species MAA and DMAA have been found in interstitial water of estuarine sediments (Ebdon et al., 1987) and the same species together with a trimethylated arsenic species were detected in interstitial water of deep sea sediments (Reimer and Thompson, 1988). It is considered most probable that these methylated arsenicals are formed by microbial methylation in situ, rather than from degradation of biological debris in the sediment.

1.2.3 Marine animals Already in 1926, Chapman described that the arsenic in lobster was a nontoxic organic compound soluble in both alcohol and water, and 16 resistant to hot dilute hydrochloric acid. Lunde (1975) showed that the same kind of organoarsenic compound was present in crustaceans, molluscs and fish, and the compound was identified as arsenobetaine by Edmonds et al. (1977). Subsequent work has revealed that arsenobetaine is the major arsenic compound in a large range of marine animals, see Table 11:1.1 (after Francesconi and Edmonds, 1993).

Among other organoarsenic compounds that have been identified in trace amounts in marine animals, the following may be mentioned (after Francesconi and Edmonds, 1994):

- tetramethylarsonium ion - arsenocholine - trimethylarsine oxide - trimethylarsine - dimethylarsinylribosides - phosphatidylarsenocholine.

Animal group Arsenic concentration % of arsenic present (No. of species) (pg/g wet weight) as arsenobetaine Fish Elasmobranchs (7) 3.1 -44.3 94 - >95 Teleosts (17) 0.1-166 48->95 Crustaceans Lobsters (4) 4.7-26 77->95 Prawns /Shrimps (5) 5.5-21 55->95 Crabs (6) 3.5-8.6 79->95 Mollusks Bivalves (4) 0.7- 2.8 44-88 Bivalves (7) 1.0- 2.3* 12-50 Gastropods (6) 3.1-117 58->95 Cephalopods (3) 49 72->95 Echinoderms (1) 12.4 60 Coelenterates (1) 7.5* 15 Sponges(2) 3.2-6.8* 13-15 * = Whole wet tissue.

Table 11:1.1 Arsenobetaine in marine animals (after Francesconi and Edmonds, 1993). Unlessspecified, values refer to muscle tissue only.

The most important of these other organic forms of arsenic in marine animals appears to be the tetramethylarsonium ion, which has been

17 detected in a variety of species (clams, gastropods, sea hare and sea anemone). In a sea anemone, this compound was shown to be the major water-soluble arsenic compound with more than 50% of the total water- soluble arsenic (Shiomi et al., 1988), while in five species of clams, it represented 20-48% of the total arsenic (Cullen and Dodd, 1989). Thus, it may be concluded that the tetramethylarsonium ion is a common constituent of mollusks, where it seems to occur predominantly in the gills. The origin of the tetramethylarsonium ion in marine animals is still unknown (Francesconi and Edmonds, 1994).

The remaining organoarsenicals identified in marine animals have, so far, only been found in trace amounts. The arsenocholine was recently demonstrated to occur, in amounts correspon-ding to 0.25% of total arsenic, in plaice, the shrimp Pandalus borealis and the mussel Mytilus edulis (E.H. Larsen, ref. in Francesconi and Edmonds, 1994).

Trimethylarsine oxide was identified as a natural component of some species of fish, and it is supposed to result from metabolic transformation of arsenobetaine within the animal (Norin et al., 1985a; Hanaoka et al., 1992). It may also arise from methylation of arsenate by the intestinal flora of the fish (Edmons and Francesconi, 1987).

Traces of the toxic compound trimethylarsine have been found in prawn and lobster, and it is hypothesized that it is produced either by microbial methylation of inorganic arsenic within the digestive gland of the animals or from microbial breakdown of arsenobetaine via trimethylarsine oxide (Whitfield et al., 1983; Hanaoka et al., 1992).

The dimethylarsinylribosides detected in some marine animals are probably derived from ingested algal metabolites, since these compounds are common constituents of marine algae (Edmonds et al., 1992).

The general picture emerging from the reviewed work is that arsenobetaine is - by far - the most common and dominant arsenic compound in marine animals. With few exceptions (examples mainly from sponges and coelenterates), arsenobetaine makes up the bulk of the total arsenic found in the marine animals, quite often more than 95%. All the other organic arsenic compounds hitherto identified in marine animals, except the tetramethylarsonium ion, seem to be produced only in trace amounts and almost "by accident" through the action of micro­ organisms associated with the animals.

18 1.2.4 Marine plants Brown algae appear to be the marine plants containing the highest concentrations of arsenic, up to 230 pg As/g d.w.. These levels are considerably higher than those in red algae (<30 pg/g d.w.), in green macroalgae (<23 pg/g d.w.), in unicellular plankton algae (about 9 pg/g d.w.) and seagrasses (<0.6 pg/g w.w.) (Francesconi and Edmonds, 1994).

The precise chemical form of the arsenic in macroalgae was established in 1981, when Edmonds and Francesconi isolated and identified the two major organoarsenic compounds in the brown macroalga Ecklonia radiata. Both turned out to be dimethylarsinylribosides, one acidic and the other a basic compound (Figure II: 1.2). Later work in this area has resulted in the identification of several other dimethylarsinylribosides in various species of brown, red and green macroalgae (Francesconi and Edmonds, 1994). It has been suggested that the distribution of arsinylribosides among the different species of algae may have some chemotaxonomic significance (Shibata et al., 1987), but this still has to be demonstrated.

Figure 11:1.2. Structures of the first two identified dimethylarsinylri­ bosides isolated from algal sources. The compounds were purified by gel permeation chromatography, thin-layer chromatography and HPLC, and were then identified by means of NMR and confirmed by X-ray crystallograpnic analysis (from Francesconi and Edmonds, 1994).

Only one of the arsinylribosides isolated from macroalgae was lipid- soluble. Although water-soluble organoarsenic compounds predominate in macroalgae, unicellular algae often contain the bulk of their arsenic

19 (up to 95%) as lipid-soluble compounds (Cooney, 1981). It was shown that the lipid-soluble arsenic compounds in unicellular algae were susceptible to hydro-lysis by phospholipases, indicating the presence of arsenic-containing phospholipids (Cooney et al., 1978). Subsequent work by Edmonds and Francesconi revealed that unicellular algae, living as symbionts in the mantle of the giant clam (Tridacna maxima), contained five of the arsinylribosides previously found in macroalgae and, in addition, four new arsinylribosides (Figure 11:1.3) and a novel taurine derivative, (Francesconi and Edmonds, 1994).

Figure 11:1.3. Structures of dimethylarsinylribosides identified in unicellular algae (after Francesconi and Edmonds, 1994).

An important observation is that, in addition to the predominant dimethylarsinylribosides, also a trimethylarsonioriboside has been detected (Figure 11:1.4), albeit in small amounts, both in macroalgae and in unicellular algae (Shibata and Morita, 1988; Francesconi et al., 1992a).

20 + Me3As 0S03

HO OH

Figure 11:1.4. Structure of a trimethylarsonioriboside identified in macroalgae and unicellular algae (after Francesconi and Edmonds, 1994).

Thus, the present state of knowledge clearly shows that unicellular and macroalgae contain in essence the same kind of organoarsenic compounds. As a consequence, it may be concluded that the biosynthesis of arsenic-containing ribosides from arsenate, occurring in marine water, seems to be a general property of marine algae, both macroalgae and unicellular algae.

1.3 Biotransformation of arsenic - metaboliccycle

1.3.1 Uptake by algae Marine algae are naturally exposed to arsenic, mainly in the form of arsenate. In uncontami-nated surface waters, where nutrient depletion may occur due to phytoplankton production, and consequently concentrations of free phosphate usually are low, arsenate can be present at levels comparable with or greater than those of phosphate (Johnson and Pilson, 1972).

According to several workers, arsenate and phosphate compete for uptake in marine algae, which was indicated by experimental exposure of the unicellular marine alga Skeletonema costatum to arsenate at different concentrations of phosphate (Sanders, 1979). Further support for this hypothesis comes from the close chemical similarity between the two elements, implying that algae might be unable to distinguish the arsenate ion from its chemical analogue, the essential phosphate ion (Maugh, 1979). In other organisms it has been demonstrated that arsenate is taken up by the phosphate transport system (Rothstein, 1963; Meharg and Macnair, 1992), but the situation in marine algae is not quite clear. At low phosphate concentrations, arsenate uptake by unicellular algae

21 increased with increasing phosphate concentrations (Andreae and Klumpp, 1979), and there was no evidence for a common mechanism of uptake for arsenate and phosphate in two species of brown macroalgae (Klumpp, 1980). Therefore, it may be concluded that arsenate is taken up by marine algae by more than one mechanism.

After having been transported into the algal cells, arsenate may begin to exert a toxic action on the alga, causing inhibition of growth, inhibition of carbon uptake, interference with phosphorus metabolism and changing the algal cell morphology (Sanders, 1979; Bottino et al., 1978). It is assumed that the production by the algae of organoarsenic compounds might, at least in part, be a detoxification process.

1.3.2 Biosynthesis of arsenic-containing ribosides At least three different pathways have been proposed for the formation of arsenic-containing ribosides in marine algae. The scheme proposed by Phillips and Depledge (1985) links the pathway for phospholipid biosynthesis to those for the formation of arsenosugars. A second proposal(Phillips, 1990) is based on the hypothesis that the enzymes of the pentose phosphate pathway could utilize arsenate rather than phosphate as a substrate. Although these proposals have some attractive features, they are, at least partly, based on some speculative assumptions.

The most plausible pathway for the genesis of arsinylribosides in algae was proposed in 1983 (Edmonds and Francesconi, 1983), and is based on the early work by Challenger (1945) and Canton! (1953). The scheme proposed by Challenger required sequential reduction of arsenic followed by oxidative methylation. Subsequent to Cantoni's identification of S-adenosyl-methionine as an active methyl donor in enzymatic systems (Cantoni, 1953), this compound was considered to be the likely source of methyl groups in the methylation of arsenic (Challenger et al., 1954).

Recent work by Francesconi et al. (1992a) has given further support to the idea that transformation of arsenic in algae proceeds according to a similar pathway, with S-adenosyl-methionine donating its methyl group and its adenosyl group to arsenic (Figure 11:1.5). However, it has also been suggested that two different pathways for the algal biosynthesis of methylated arsinylribosides may occur in nature (Phillips, 1994). It has, moreover, been suggested that the dimethylarsinylribosides might be able to participate in cellular redox reactions in the algal cells (Francesconi and Edmonds, 1994).

22 0

0—As—0 Reduction

HO OH

Mtthylarsomc avid Dimvlhylarsinic acid (DiimMhylarsinyludeno.sine) (MAA) (DM AA) (Compound 13)

0 X 1 OR

Reduction then A HO OH HO OH

S-Adenosylmethionine I HO OH B (AdoMel) |

Figure 11:1.5. Proposed pathway for the biogenesis by algae of arsenic- containing ribosides from arsenate. All proposed intermediates have been identified from marine algal sources. AdoMet is the probable source of the methyl and adenosyl groups (after Francesconi and Edmonds, 1994).

1.3.3 Origin of arsenobetaine in marine animals A great deal of research has been devoted to the problem of determining the relative importance of the two possible sources of arsenobetaine, water and food, to marine animals. For the time being, no evidence exists to indicate that aquatic animals are capable of synthesizing arseno­ betaine de novo from arsenate (Francesconi and Edmonds, 1994). Then, the question remains whether or not the "hidden" arsenic species in water may, at least partly, be composed of arsenobetaine and other methylated forms of arsenic, that may be taken up from the water and - directly or after transformation - contribute to the arsenobetaine burden of marine animals. However, as mentioned in section 1.2.1, the results of experiments undertaken are at variance with the situation observed in

23 natural populations. Thus, based on the data available, seawater does not appear to be a significant source of arsenobetaine to marine animals. In stead, the evidence suggests that marine animals accumulate arsenobetaine primarily from their food (Francesconi and Edmonds, 1994).

The problem is that the food of primary consumers does not seem to contain arsenobetaine. Although the levels of arsenic in algae are three to four orders of magnitude higher than the levels in seawater, none of the many studies of the composition of algal arsenic have demon-strated the presence of arsenobetaine in marine algae. The presumed immediate precursor of arsenobetaine, namely arsenocholine, was similarly not detected in algae (Francesconi and Edmonds, 1994).

Since there is no experimental evidence indicating de novo synthesis of arsenobetaine from ingested arsenate by fish (although indications exist that small proportions of ingested arsenate can be converted to an organic form, e g trimethylarsine oxide), the possibility remains that the food of primary consumers contains a precursor that is converted to arsenobetaine at a later stage in the food chain.

Food chain experiments were carried out by Cooney and Benson (1980) with the unicellular alga Dunaliella tertiolecta and the lobster Homarus americanus, and speciation of the arsenic compounds in the two trophic levels was made after exposure of the alga to 74As as arsenate in the seawater. In a similar experiment, Klumpp and Peterson (1981) examined the arsenic transformation in the marine food chain: macroalga (Fucus spiralis) —> herbivorous gastropod (Littorina littoralis) —> carnivorous gastropod (Nucella lapillus), after exposure of the alga to 74 As as arsenate. In none of these experiments it was possible to detect any arsenobetaine in the consumers, which shows that neither the lobster nor the gastropods were able to directly transform the ingested algal arsenic compounds into arsenobetaine. However, in the latter study, most of the 74As lable in the gastropods was associated with a single arsenic compound with unknown identity. The chromatographic and electrophoretic properties of the compound seem to fit those of the tetramethylarsonium ion, which was later shown to be a common constituent of mollusks (Cullen and Reimer, 1989). However, the main conclusion from these experiments is that primary (and secondary) consumers appear to be unable to biosynthesize detectable levels of arsenobetaine (or arsenocholine) from directly ingested algal arsenic.

24 The next possibility to investigate was whether an intermediate (perhaps microbially mediated) stage is necessary to render the algal arsenic into a form that can be accumulated by animals. This view was supported when it was found that under anaerobic conditions (in beach sediments) there was an almost quantitative conversion of the naturally occurring algal arsenic compounds into a single compound, subsequently identified as dimethylarsinylethanol (Francesconi and Edmonds, 1994). No arsenobetaine was detected. Since the new compound, dimethylarsinyle­ thanol only requires further methylation and then oxidation of the primary hydroxyl group for conversion to arsenobetaine via arsenocholine or dimethylarsinylacetic acid, this pathway (Figure 11:1.6) was hypothetically proposed for the natural formation of arsenobetaine.

? MejAs cooh

Dimelhylaf.sinylocetlc acid

OXIDATION METHYLATION

O » ANAEROBIC DECOMPOSITION t Me3As\/C00" (C3-C4 cleavage)

Dimethylarsinylethanol Arsenobetaine

Dimethylarsinyiribosides

METHYLATION OXIDATION

+ X

Arsenocholine

Figure ILL 6. Proposed pathways for the formationof arsenobetaine from dimethylarsinyiribosides via the intermediate compound dimethylarsinylethanol (after Francesconi and Edmonds, 1994).

The remaining problem was that attempts to carry out further methylation of the dimethylated arsinyl intermediates by natural processes outside the algae have been unsuccessful (Francesconi and Edmonds, 1994). However, a new line of reasoning emerged from the fact that, in addition to the dimethylated arsinylribosides that make up the bulk of the algal arsenic, small amounts of the trimethylated analogues were also detected in both macroalgae (Shibata and Morita, 1988) and unicellular algae, cf. section 1.2.4 (Francesconi et al., 1992a). When one of the trimethylarsonioribosides was subjected to conditions

25 of anaerobic microbial activity, it underwent degradation to give a virtually quantitative yield of arsenocholine (Francesconi et al., 1992b). On the basis of these results, an alternative pathway for arsenobetaine was proposed.

In order to confirm the possibility that any of the proposed pathways of arsenobetaine formation does occur in nature, feeding experients were carried out with yelloweye mullet (Aldrichetta forsteri). Separate groups of fish received food containing known quantities of each of the following organoarsenic compounds: (a) dimethylarsinylethanol, (b) dimethyl-arsinylacetic acid, (c) arsenocholine, and (d) arsenobetaine. After 3-4 weeks, the amount of arsenic in muscle tissue of the fish was analysed and the amount of arsenic retained as a percentage of the total arsenic ingested was calculated (Francesconi et al., 1989). It was found that no arsenic was retained in the fish body after ingestion of compounds (a) and (b), while 37-39% of the arsenic was retained after ingestion of either of compounds (c) and (d). The chemical form of the retained arsenic in the two latter groups was, in both cases, arsenobetaine, but no arsenocholine was present (Francesconi et al., 1990).

These results indicate that further methylation of the dimethylated compounds did not occur within the fish, because the products of such a process (arsenocholine and arsenobetaine) would have been retained by the fish. Furthermore, the results support the assumption that arsenocholine is the immediate precursor to arsenobetaine in marine animals, and provide an explanation for the absence of significant amounts of arsenocholine in marine animals: in fish, at least, arsenocho­ line is so readily metabolized that none is detected (Francesconi and Edmonds, 1994).

Whether dimethylarsinylribosides or trimethylarsonioribosides serve as precursors to arseno-betaine in natural marine systems remains to be established. However, the two proposed pathways are not mutually exclusive, and it is assumed that the trimethylated compounds in algae at least contribute to the arsenobetaine content in marine animals. The degradation of arsenic-containing ribosides (both di- and trimethylated compounds) into possible precursors of arsenobetaine has been shown to occur only under anaerobic conditions, in microbially active environ­ ments. Whether this actually takes place only in the sediments or can occur also in anaerobic micro-environments in other marine compart­

26 ments, e g in the water column or in the gut of the fish, is presently not known (Francesconi and Edmonds, 1994).

1.3.4 Complete metabolic cycle of arsenic in the marine environment Philips (1990) has proposed a consolidated scheme for the inter­ conversion of the various forms of arsenic found in marine environ­ ments. This scheme, later modified by Philips (1994), see Figure 11:1.7, was based on work by Philips and Depledge (1985) and by Edmonds and Francesconi (1988), but also included a novel biosynthetic pathway for the production of arsenosugars in macroalgae. In this hypothetical, new pathway, it is assumed that the enzymes of the pentose phosphate pathway utilize DMAA rather than phosphate as a substrate, resulting in the synthesis of the dimethylarsinoyl derivative of 5-phospho-b-D- ribosylamine, and this compound could act as a precursor of all the documented arsenosugars found in marine macroalgae.

i S'-Dimcthylarsinoyl- S'-Dimethylarsinoyl- S'-Dimethylarsinoyl- glucosc ► ci-D-rtbulose —— ------► p-D-ribosylamine

Arscnolipids not containing sugars

Arsenate

I’edenosyl Arsenoeihanolamine

Trimethylarsine / oxide

Tetramcthylarsonium ion

Figure 11:1.7. Proposed pathways for the biosynthesis and catabolism of the different arsenic species found in marine organisms (after Philips,

27 1.4 Sinks, mobilization and bioavailability of arsenic species Arsenic, as many other elements, is taken down to the bottom of oceans as sorptive aggregates with hydrous ferric and manganese oxides or with organic matter. In sediments of the open ocean, the average level of arsenic has been estimated at 40 pg/g (Bostrom and Valdes, 1969). The enrichment of arsenic in iron-manganese nodules and crusts is well known, resulting in concentration of 100 pg As/g to well in excess of 1000 pg As/g (Calvert and Price, 1977).

In coastal waters, it has been observed that intense biological activity in the surface layers, and the subsequent uptake of arsenate into phytoplankton, leads to depletion of dissolved arsenic from these layers, compared to the concentrations measured below the thermocline, where regeneration from particulates occurs (Andreae, 1979; Waslenchuk, 1978). When measuring the concentrations of dissolved arsenate and arsenite in several basins off the California coast, Andreae (1979) found that arsenate in sediment interstitial waters usually was somewhat lower (1.3 pg/1) than in the bottom water (1.8 pg/1), and that the arsenate was depleted near the sediment-water interface. It was suggested that diffusion of arsenate from the water to the sediment may be occurring, followed by adsorption to the solid phase.

On the other hand, Andreae (1979) also found that the arsenite concentrations in sediment interstitial waters were higher (up to 0.25 pg/1 higher) than in the bottom water, indicating that this arsenic species may diffuse out of the sediments. However, it was not clear whether or not the arsenite can escape through the oxidized surface sediment layers into bottom waters. Nonetheless, Sanders (1980) used the above data to calculate the possible maximum flux of arsenite out of the sediments in Georgia Bight, and he concluded that this was much lower than other arsenic sources (river run-off, atmospheric deposition and deep-water intrutions) in the studied marine area.

The mechanisms of arsenic sedimentation and the subsequent rates of reduction and oxidation of arsenic compounds in the interstitial waters of marine sediments are still largely unknown (Sanders, 1980). While arsenic flux out of marine sediments has not been directly measured in the field, this has been suggested to explain excess arsenic in bottom waters in some coastal seas (Sanders et al., 1994). Many more studies have been conducted to measure the arsenic flux from contaminated estuarine sediments (see section 2.4).

28 In a recent experimental study (Ettajani et al., 1996), fine-particle marine sediment was artificially contaminated with arsenate and then submitted to in vitro desorption tests, in which enzymes and pH changes were used to mimic the digestive processes in molluscs. Although different enzymes combined with low pH (down to pH 4) induced the desorption of 3-24% of sediment-bound arsenic, the accumulation of this element in the soft tissues of oysters remained low after exposure to contaminated particles (12.2 pg/g dry weight as compared to the control value of 8.9 pg/g, after 14 d exposure). However, minor cytological effects were noted in the oysters exposed to sediment-bound arsenic, indicating that this arsenic had a certain bioavailability and could be transported into the soft tissues of the animals.

Weis and co-workers (see Weis and Weis, 1996a for references) have conducted a series of studies of the leaching and the effects of metals from chromated copper arsenate (CCA-) treated wood used as pilings and bulkheads in the marine or estuarine environment. In the United States, wood intended for marine use is treated with high amounts of CCA, between 24 and 40 kg/m3. Weis et al. (1991) showed that all three metals leach from the treated wood into sea water. Leaching for 2 weeks in 28 %o sea water was 30 pg Cu/cm2, 0.8 pg Cr/cm2 and 8.0 pg As/cm2, and the leaching rates generally decreased over time. It was clearly shown that the released metals accumulated in epibiota on the CCA-treated wood panels, but the arsenic cencentration decreased steadily with each repeated submersion and removal of the wood (Weis and Weis, 1996b). Also the sediments, particularly the fine fraction, and the benthic animals near the CCA-treated wood were highly contaminated with metals (Weis and Weis, 1994). The highest concentrations in the fine sediments were found for copper, and the lowest for arsenic, while copper and arsenic were both high in benthos. Thus, the use of CCA-treated wood in the marine environment obviously introduces bioavailable arsenic into the ecosystem, but after presoaking of the wood for two months, the leaching and the subsequent accumulation of arsenic in biota are reduced to low levels (Weis and Weis, 1996b). Therefore, a proper presoaking of CCA-treated wood before using it in the marine environment, might reduce the arsenic leaching and the contamination of sediments and biota to insignificant levels, maybe except in the close vicinity of the wooden structures.

29 1.5 Uptake in organisms and bioaccumulation Arsenic uptake by phytoplankton and attached macroalgae forms the most important pathway for arsenic entrance into marine food webs (Sanders et al., 1994). Marine phytoplankton actively take up arsenate occurring at natural concentrations, and seem to regulate arsenic levels in the cells over a large concentration range independently of the phosphate concentrations. Significant differences in regulating capacity exist between different algal species (Andreae and Klumpp, 1979).

Arsenic levels in seaweeds (Fucus spiralis and Ascophyllum nodosum) have been shown to reach a steady-state in 1 to 8 days, depending on the species and external arsenic concentration. The uptake was increased in direct proportion to increasing temperature. The accumulation of arsenate was four times that of arsenite. Arsenic uptake requires energy, which is derived from respiration rather than from photosynthesis (Klumpp, 1980).

Only little uptake of dissolved arsenic from water occurs in marine invertebrates or vertebrates (Edmonds and Francesconi, 1987; Sanders et al., 1989; 1994). The relatively high uptake rate by algae, however, allows a pathway for arsenic into higher trophic levels, as the organo- arsenic compounds formed in algae are available to herbivores (Klumpp, 1980; Sanders et al., 1994). The food chain accumulation of arsenic from Fucus spiralis in the snail Littorina littoralis was primarily in the soft tissues, especially the digestive gland and gonads, while in the carnivore Nucella lapillus, about 85% of the accumulated arsenic was associated with the shell. Compared with the macroalgae, the marine snails exihibit a much greater ability for eliminating arsenic from the tissues (Klumpp, 1980).

Further comments on uptake and bioaccumulation of arsenic in marine animals are given in section 1.3.3 and corresponding phenomena in the brackish water environment are presented in section 2.5.

1.6 Natural factors affecting speciation and bioavailability As previously mentioned, arsenate is the predominant form of arsenic in marine waters. In addition to arsenate, three other arsenic species are commonly encountered in dissolved form: the reduced inorganic form, arsenite, and two methylated forms, methylarsonate (MAA) and dimethylarsinate (DMAA). Other, more complex organic forms of arsenic may be present as well, but their concentrations are low, and they

30 are not generally detected using conventional analytical methods (Sanders et al., 1994).

While arsenate is the thermodynamically more stable form in oxic waters, arsenite is more stable in reducing environments such as anoxic water or sediments, but it is also found in aerobic systems. Oxidation of arsenite to arsenate, in the presence of oxygen, occurs rapidly, with a residence time of 0.5 to 10 days in coastal waters (Sanders and Windom, 1980; Cutter, 1992). The rate of arsenite oxidation in seawater is enhanced by increases in pH, salinity, temperature and arsenite concentration (Johnson and Pilson, 1975). The relatively rapid oxidation rate of arsenite implies that the occurrence of arsenite in marine oxic waters is closely related to the biological processes responsible for its formation through reduction, such as primary production in surface waters and bacterial degradation in the layers below.

Phytoplankton species seem to differ in their ability to produce reduced and methylated arsenic compounds, with some species capable of producing arsenite, MAA and DMAA, and others producing only one or none of these compounds (Andreae, 1983; Sanders and Riedel, 1993).

Arsenite formed in the sediments or at the sediment-water interface through bacterial reducing activity may be rapidly oxidized back to arsenate if the sediment surface is aerobic. It has been suggested that Mn4+ and Fe3+ oxides present on the surface of sediment particles may be responsible for catalyzing the oxidation of arsenite through an electron transfer mechanism (Oscarson et al., 1980). The oxidation by means of manganese seems to be more effective than that by iron. Over long periods of time, however, also iron may play a significant role.

The simple, methylated arsenic species are produced biologically, either by methylation of inorganic arsenic or by degradation of more complex organoarsenic compounds such as arsenocholine and arsenobetaine (Hanaoka et al., 1987; Anderson and Bruland, 1991). Methyl arsenic species are more persistent than arsenite, with much longer degradation times. Therefore, these species can persist in natural systems for relatively long periods of time, and their presence does not depend so much on the processes that led to their formation (Sanders et al., 1994).

The rate of uptake of arsenate in marine algae and invertebrates appears to be higher than that of arsenite, possibly reflecting a higher bioavailability of arsenate. Because of the chemical similarity between

31 arsenate and phosphate, arsenate is taken up by autotrophs along with phosphate. Arsenate also interferes with the biochemical functions of phosphate, particularly phosphorylation and ATP-production (Planas and Healey, 1978; Blanck and Wangberg, 1988a). Therefore, when the ratio of arsenate to phosphate is relatively high, arsenic toxicity to phytoplankton becomes more likely (Wangberg and Blanck, 1990; Sanders et al., 1994). Presumably, the arsenic transformation reactions inside the algal cells have been evolved to minimize this toxicity by altering and inactivating arsenate.

Arsenate adsorbed onto the surface of fine, suspended particles is not entirely non-bioavailable. As was shown by Ettajani et al., (1996), a certain fraction of the arsenic could be released from the particles under pH and other conditions that mimic the environment inside the gut of mollusks. It was also shown that such arsenate-laden particles could induce slight toxic reactions in the mollusks (oysters).

In summary, in the relatively stable environment existing in marine systems, arsenic is mainly present in a form, arsenate, which is readily bioavailable. Since the concentration of phosphate in open marine waters usually is low or very low, the rate of uptake of arsenate into autotrophs is relatively high. The uptake is followed by a rapid transformation of arsenate to complex organic arsenic species, which generally have a lower toxicity than arsenate and arsenite.

2. Estuarine or Brackish Water Environment

2.1 Arsenic levels and speciation in water and sediments 2.1.1 Estuarine waters Elements like arsenic, that are biologically and chemically reactive can undergo considerable changes in chemical form and bioavailability, particularly in estaurine areas (Sanders et al., 1994). The main difference between truly marine and estaurine systems is that, while the former are relatively stable, the latter are dynamic systems, subject to wide, temporal and spatial, variations in physical, chemical and biological parameters. Because of their high physical variability, estuaries and brackish water sea ecosystems tend to be biologically simple, characterized by high biomass and high productivity, low species diversity and large, natural fluctuations in species abundance (Day et al., 1989). The biological composition and dynamics of an aquatic ecosystem determine, to a large extent, the various chemical forms of arsenic and their relative concentration. Therefore, in temperate,

32 estuarine ecosystems, arsenic undergoes dynamic seasonal and spatial changes in concentration and speciation.

The most typical feature of arsenic speciation in estuaries, as opposed to the open sea, is the seasonal occurrence of relatively high concentrations of reduced and methylated arsenic species, particularly during warm summer months. For example, in English estuaries, the appearance of reduced and methylated arsenic occurred after periods of high phytoplankton productivity (Howard et al., 1982; 1984).

The temporal dynamics of arsenic speciation in a productive estuary has been studied in Patuxent River, a subestuary of the Chesapeake Bay, Maryland, U.S., and it was clearly shown that not only the total arsenic concentration in the water greatly varied, but also the relative distribution between different arsenic species (Riedel, 1993), Figure 11:2.1. A summer maximum of total arsenic was observed, which was largely composed of arsenate (1 pg/1). The arsenate concentration fell during winter to about 0.1 pg/1. Arsenite was irregularly present at low concentrations. Peaks of DMAA occurred at various times throughout the year, and appeared to be closely linked to dense dinoflagellate blooms and resulting phosphate limitation. A peak of MAA occurred at the same time as the summer peak of arsenate, following the spring DMAA peak; thus, it is possibly a degradation product of DMAA (Sanders et al., 1994). As an alternative, the MAA peak may be a result of methylation by specific algal taxa (Sanders, 1985). These transforma­ tions are further discussed in section 2.6.

33 1.2 B-Q MMA ■ ■ DMA O O orsenite *-# arsenate

A M J F M A M

DATE (1988-1989)

Figure 11:2.1. Arsenic concentration and speciation in the Patuxent River from June 1988 to May 1989 (after Riedel, 1993 and Sanders et a/., 7%%).

Within the Chesapeake Bay, a similar seasonal pattern was observed, coupled with a spatial variability (Sanders, 1985). Almost all arsenic entering the head-waters of the Bay is in the form of arsenate. Reduction and methylation take place during the warmer months, resulting in changes in the pattern of arsenic speciation down the main stem of the Bay (Figure 11:2.2). Arsenite is produced in the upper regions with low salinity (1-10 %o), while methylated species reach high concentrations in the lower regions of the Bay (10-20 %o).

Quite similar patterns were observed in several other estuaries, for example in Florida, England, Portugal and in the Baltic Sea. In almost all cases, the appearance of enhanced concentrations of reduced and methylated, dissolved species was well correlated with phytoplankton production (Sanders, 1983; Froelich et al., 1985).

34 SALINITY (°/oo)

Figure 11:2.2. Arsenic speciation in the waters of Chesapeake Bay during productivesummer months. From bottom to top: arsenate, arsenite and methylated arsenic species, (after Sanders et ah, 1994).

In the Baltic Sea, with its anoxic bottom water, a substantial reduction of arsenate to arsenite was observed below the oxic-anoxic interface, at a depth of about 170 m in the Gotland Deep (Andreae and Froelich, 1984). The concentration of total arsenic increased from 0.6 pg/1 in the surface water to almost the double below the redoxcline. In the uppermost water layers, above the seasonal thermocline, almost half of the total arsenic was in the form of arsenite, DMAA and MAA (Figure 11:2.3). Thus, the arsenate was depleted in the euphotic zone. However, the methylated species remained low in the whole water column from the depth of 40 m down to the bottom. On the contrary, arsenite sharply increased in concentration at the redoxcline and remained the dominating species in the anoxic layers.

An interesting pattern was observed at the measuring stations in the Gulf of Finland, where the dominating arsenic species, above the seasonal thermocline, belonged to the methylated forms DMAA and MAA, together making up 83% of total dissolved arsenic. The samples were taken in June 1981, during the summer peak production period and the

35 high level of methylarsenicals in this area may depend on the relative rate of methylation by algae.

The vertical distribution of different arsenic species in the Baltic Sea water suggests that arsenate is taken up by phytoplankton, similarly to phosphate, in the surface mixed layer. Arsenic is then, at least partially, regenerated at depth, especially near the anoxic interface (Andreae and Froelich, 1984). The arsenic-salinity relationships are plotted in Figure 11:2.4, from which it can be seen that the profile for each station begins (at the surface) at low salinities and low arsenic concentrations. Then there is an increase of arsenic with depth at constant salinity, indicating rapid regeneration of arsenic in the seasonal thermocline. Below the redoxcline (shown be an arrow) at stations BY11 and BY 15 in the Gotland Deep, arsenic continues to increase at nearly constant salinity as a consequence of regeneration of arsenic in the anoxic layers (Andreae and Froelich, 1984).

AM "M O O.S 1.0 2.5 3.0

Figure 11:2.3. Concentration profiles of different arsenic species in the Baltic Sea water, at the Gotland Deep. Asj = total dissolved inorganic As; Asj = total dissolved As (after Andreae and Froelich, 1984).

36 20

15-

1 10- f

i i i i i . i i—.i i i i 10 15 SALINITY (%.)

Figure 11:2.4. Plots of total dissolved arsenic versus salinity at five stations in the Baltic Sea. BY5 is in the Bornholm Deep, BY11 and BY15 in the Gotland Deep, BY23 and BY26 in the Gulf of Finland (after Andreae and Froelich, 1984).

The average concentration level of total dissolved arsenic in the surface waters of the Baltic Sea (0.6-0.7 jxg/1) is intermediate between the level in rivers entering into northern Bothnian Bay (0.2 jj.g/1) and the level in the North Sea (1.5 pg/1). In the Bothnian Bay, concentrations of total arsenic were somewhat enhanced (0.65 pg/1), in 1987, due to local contamination from a smelting complex (Hotter, 1993).

Arsenic flux from contaminated estuarine sediments has been examined under controlled conditions in laboratory microcosms (Riedel et al., 1987; 1989). The fluxes from undisturbed, oxic sediments were low or undetectable. However, in the case of mechanical disturbance of the bed by resuspension or by the activity of benthic infauna, substantial fluxes of arsenate, arsenite, MAA and DMAA were observed (Sanders et al., 1994). Sediments covered by anoxic waters showed even greater fluxes of arsenic.

37 2.1.2 Estuarine sediments Speciation studies of arsenic in estuarine sediments have not been found in the literature, but it can be assumed that the relative amount of arsenite is higher in the interstitial waters of sediments than in the overlying water, particularly if the sediments are anoxic. However, in one of the measuring profiles in the Baltic Sea, where the bottom water was anoxic and the sulphide concentration relatively high (20 pM), the arsenite concentration sharply decreased close to the sediment-water interface and this was accompanied by an increase in the arsenate concentration (Andreae and Froelich, 1984). Also the methylated species MAA and DMAA may be higher at the sediment-water interface, due to decomposition of the complex, methylated algal arsenic compounds.

The concentration of total arsenic in sediments of the Baltic Sea and the Skagerrak is shown in Table 11:2.1.

Basin/ Sea Surface sediment (0-1 cm) Older sediment (>20 cm) Bothnian Bay 109 7 Bothnian Sea 27 9 Baltic Proper 15 9 Skagerrak 2 9

Table 11:2.1. Mean concentrations of total arsenic in sediments from different basins of the Baltic Sea and Skagerrak, in jug/g dry matter (after Natter, 1993).

The data in Table 11:2.1 clearly indicate the influence of discharges from the metal smelting complex on the western coast of the Bothnian Bay. At a distance of 5-30 km from the smelter, arsenic levels in organic, surface sediments ranged from 160 to 4,600 pg/g dry matter, in 1989 (Walterson and Landner, 1996). However, the sediments of the Baltic Proper have a lower arsenic content than ocean sediments (40 pg/g) and also lower than the sediments in Skagerrak.

38 2.2 Arsenic levels and speciation in biota Both the levels of total arsenic and the distribution between various arsenic species in tissues of organisms living in brackish water are much less known than in true marine organisms. However, a set of relevant data, covering several trophic levels of the brackish-water ecosystem in the Baltic Sea, was reported from a model ecosystem experiment (Notini et al., ref. in Blanck et al., 1989). This experiment allowed studies of the variation of the natural levels of arsenic over a complete annual cycle in various benthic organisms from the littoral zone of the Baltic Sea. The data were obtained from the control systems, consisting of large outdoor pools fed with uncontaminated Baltic Sea water, containing 0.5-0.8 pg/1 of total, dissolved arsenic (Blanck et al., 1989). Samples of macroalgae and invertebrates were analyzed for total arsenic as well as for the inorganic arsenic fraction (Table 11:2.2).

Species Group Total arsenic % inorganic

Fucus vesicolosus Brown alga 20 5-7 d:o, young shoots 33 5-7 Ectocarpus siliculosus Brown alga 25 - Cladophora sp. Green alga 3-9 - Ceramium sp. Red alga 3-9 - Mytilus edulis (soft tissue) Bivalve 12,5 <10 Cardium sp. - ” - Bivalve 14,5 <10 Mya arenaria - ” - Bivalve 5-15 55-65 Idothea balthica Crustacean 4,9 20-40 Nereis diversicolor Polychaete 10 60

Table 11:2.2 Arsenic concentration (jug/g dry weight) in the tissues of various benthic organisms from the Baltic Sea, after exposure to uncontaminated brackish water in large outdoor pools. Adapted after Blanck et al., 1989.

As demonstrated in Table 11:2.2, the highest arsenic levels were found in the benthic brown algae, which also contained the greatest amount of organic arsenic compounds. Especially the detritus-feeding bivalve Mya arenaria and and the predator polychaete worm, Nereis diversicolor, contained a large fraction of inorganic arsenic (as recovered by HC1 distillation). This was in contrast with the (phytoplankton) filtering bivalves Mytilus and Cardium, which exhibited low amounts of inorganic arsenic.

39 A few field samples from the Baltic Sea have been investigated, mainly in areas where exposure to anthropogenic arsenic occurs. A survey of spring samples of zooplankton, in 1979, showed that the average content of total arsenic in the animals was 15, 20 and 10 pg/g dry weight in the Bothnian Bay, the Bothnian Sea and the Baltic Proper, respectively. The reference value, in uncontaminated areas of the Baltic Sea is supposed to be <10 pg/g d.w.

Gastropod snails (Lymnaea palustris) from the littoral zone of the Bothnian Bay held a background level of 6 pg/g d.w. of total arsenic, and in areas close to the arsenic-emitting smelter, the concentrations rose to about 110 pg/g, in 1989, but only to about 12 pg/g d.w., in 1994, after a sharp reduction of the arsenic emissions (Walterson and Landner, 1996).

Fish from uncontaminated parts of the Bothnian Bay were also analyzed, and the muscle tissue of perch and pike contained 0.14-0.28 pg/g d.w., while the liver held 0.24-1.5 pg/g (Norin and Vahter, 1984). Baltic herring from the same area had muscle tissue concentrations of total arsenic of about 1.2 pg/g d.w., which was five to ten times less than the arsenic concentrations in fish from the Swedish west coast (Ljunggren et al., 1971). The mean relative amount of inorganic arsenic in fish from this brackish-water area was between 5 and 10% of the total arsenic content (Norin et al., 1985b), indicating that the distribution between different arsenic species was similar to what has been found in marine fish. It was furthermore indicated that some 5-10% of the organic arsenic in the fish tissues was in the form of arsenocholine, while most of the remainder consisted of arsenobetaine.

Based on the relatively scarce data on arsenic content and speciation in organisms living in brackish water, it may be concluded that, generally, the concentrations of total arsenic are lower than in marine organisms. The difference appears to be particularly great when fish from the two environments are compared. In the case of invertebrates, it appears that bivalves and crustaceans from marine and brackish water areas show less difference in the content of total arsenic, while gastropods have very variable levels (cf. Table 11:2.2 and Table 11:1.1, where it should be noted that concentrations are related to wet weight). Fish from both habitats seem to contain very small fractions of inorganic arsenic, while the picture becomes more complex when it comes to the different inverte­ brate groups. Obviously, the feeding habits of invertebrates largely determine the speciation of arsenic in their bodies, which means that no general conclusions can be made regarding the relative amount of

40 inorganic arsenic occurring in a whole taxonomic group. However, the available data do not indicate that there should exist any systematic differences between invertebrates living in brackish water seas and those living in marine areas.

Macroalgae, particularly brown algae, from marine ecosystems seem to contain much higher levels of total arsenic than the corresponding groups of algae in brackish water ecosystems. This difference may be due to the generally higher concentrations of dissolved arsenic in marine waters, but possibly also to a more efficient uptake of arsenate (which might be related to the usually lower phosphate concentration) and a subsequent biosynthesis of arsenic- containing ribosides in marine algae.

2.3 Biotransformation of arsenic The rate of reduction and transformation of arsenate by phytoplankton can undergo very dramatic variations in estuaries, due to the great fluctuations in natural conditions in such habitats. For example, correlations have been observed in the field between reduced and methylated arsenic species and phytoplankton densities, chlorophyll-a concentrations, and primary productivity (Sanders and Riedel, 1993). Furthermore, phytoplankton species apparently differ in their ability to produce arsenite, MAA and DMAA. This differential ability may produce temporal and spatial trends in arsenic speciation within estuaries, and these trends may be tied to changes in phytoplankton species composition and the succession of dominant species (Sanders et al., 1994). For example, Sanders (1985) noted a very strong correlation between a dominant cryptophyte, Chroomonas sp., and the occurrence of MAA in Chesapeake Bay (Figure 11:2.5). Another example is the association of high concentrations of DMAA with a dramatic bloom of a dinoflagellate (Riedel, 1993).

However, all algal blooms are not associated with elevated levels of reduced and methylated arsenic. In the low-salinity regions of the Tamar estuary in England, no increase of methylated arsenic was observed in connection with the dominance of the freshwater centric diatom Cyclotella atomus (Howard et al., 1988). A similar absence of arsenic reduction during the spring diatom blooms was seen in other rivers in England, but arsenic reduction was observed later in the spring (Howard et al., 1984). The explanation might be either that diatom species do not produce significant amounts of methylated arsenic or that the lack of

41 arsenic reduction is caused by the still high levels of phosphate in the water early in spring, when the diatom blooms occur. At a later stage, when the phosphate is depleted, the transformation of arsenic into reduced and methylated forms can start (Sanders et al., 1994).

400-

300-

200-

100-

SAUNiTY (°/oo)

Figure 11:2.5. Relationship between the occurrence of methylarsonic acid (A) and the densities of a dominant cryptophyte, Chroomonas sp., in the waters of Chesapeake Bay (after Sanders et al, 1994).

42 The rates of arsenic reduction during phytoplankton blooms can be very rapid. For example, in the Patuxent River, the maximum rates varies from 130 ng/1, day in spring to >330 ng/1, day during midwinter dinofla- gellate blooms (Sanders and Riedel, 1993). Viewed in relation to the bio ­ mass, the overall reduction rates varied from 2 to about 20 fg/cell, day at arsenate concentrations between 0.5 and 2.0 pg/1 (Sanders et al., 1994).

However, cell densities are not the only deciding factor. Periods of arsenate reduction are, in general, associated with periods of rapid decline in phosphate concentrations (Sanders and Riedel, 1993). This may reflect the existance of a causal relationship between phosphate concentration in the surrounding medium and rate of uptake and transformation of arsenic in phytoplankton cells. Thus, it may be concluded that the most important controlling factors of the biotrans ­ formation of arsenic in estuaries and brackish water seas are the phytoplankton species composition and arsenic-phosphorus dynamics.

One consequence of the prospensity of arsenic to change forms after its introduction to the aquatic environment is a greater degree of uncertainty about the effects of a release. Transformation of arsenate to arsenite or DMAA, while an apparent benefit to the phytoplankton, may be more detrimental to fauna components, which may be more sensitive to the reduced and methylated forms (cf. Nissen and Benson, 1982), see further discussion in section III.

2.4 Sinks, mobilization and bioavailability of arsenic species The capacity of retaining arsenic being introduced into model ecosystems of the type described in section 2.2 was investigated after one year of continuous exposure to concentrations of arsenate in the incoming water of 8 pg/1 and 75 pg/1, respectively (Notini et al., 1987). The total amount of arsenate added to the systems was 22.5 g and 225 g, and the percent retained was 5.3 and 1.4, respectively. The distribution of arsenic between the different main components of the system is shown in Table 11:2.3.

Added Sediment Fucus Filamentous Inverte­ Fish Total %of amoun algae brates system added A. 22,500 1,165 21 0 6 0 1,190 5,3 B. 225,000 3,140 0 27 22 1,4 3,190 1,4 Table 11:2.3. Distribution of added arsenic in model ecosystems after 1 year of exposure to 8 pg/l (A) and 75 pg/l (B) of arsenate. Values in mg total arsenic (after Notini et al, 1987).

43 The reason why no arsenic was retained by the Fucus in the high concentration pool was that the algae were killed at this level of arsenate exposure. In both cases, about 98% of the retained amount was found in the sediment, indicating that the sediment is the major sink for arsenic being introduced into a littoral ecosystem.

In a similar type of experiment, using small, microcosm-type containers with uncontaminated soft sediments from the deep areas of the Baltic Sea and benthic organisms from the profundal community (Monoporeia affinis and Macoma balthica), arsenic retention was studied over a period of 222 days (Blanck et al., 1989). The microcosms received a continuous flow of uncontaminated bottom water from the Baltic Sea and to the incoming water, the following forms of arsenic were added: (a) arsenic- contaminated phytoplankton (Skeletonema costatum), (b) dissolved arsenate at concentrations of 20 and 100 pg/1, and (c) a mixture of arsenate and ferric hydroxide (20 pg As + 40 pg Fe/1, and 100 pg As + 200 pg Fe/1). Of the total amount of arsenic added, about 20% was retained in series (a), 1% in series (b), and 3% in series (c). Thus, iron hydroxide is an efficient scavenger of aisenate in the water, increasing the rate of arsenic deposition in the sediment. However, at least in sediments containing benthic invertebrates, the most efficient retention of arsenic was obtained when it was carried (probably in the form of organic arsenic compounds) by algal cells.

It might not be possible to directly interpret these experimental results in terms of sedimentation or fixation rates in natural systems. However, it may be concluded that, in relative terms, arsenic bound to organic material is more readily retained by the sediment compartment than is arsenic in inorganic form, either dissolved or particulate. Furthermore, a littoral system has a limited capacity to retain arsenic: with higher amounts introduced, higher amounts are flushed through the system. As far as the different species of dissolved arsenic are concerned, it has also been demonstrated that arsenate and arsenite are readily sorbed to sediment and suspended solids, while DMAA is less particle-reactive, and is therefore more easily flushed through an estuary, where it is formed as a result of phytoplankton production and methylation (Sanders and Riedel, 1993). Arsenate introduced in relatively low concentrations to a littoral system with dense communities of macroalgae is effectively retained by the algae, and is transformed into arseno-sugars, which are stored within the algal tissue.

44 During phytoplankton blooms in the Baltic Sea, arsenate is readily taken up by the algae, methylated within the cells and partly excreted in the form of MAA and (mainly) DMAA. This transformation is clearly seen as a depletion of arsenate in the water of the euphotic zone and an increase of DMAA (Figure 11:2.2). However, it can be assumed that most of the arsenic taken up by the phytoplankton is transported towards the bottom waters together with sedimenting dead algal cells. This is reflected by the sharp increase in total dissolved arsenic, and particularly of arsenite, in the bottom waters (Andreae and Froelich, 1984). This sharp increase in total dissolved arsenic with depth has been observed both in the Bornholm Basin (with low oxygen levels, but without any hydrogen sulfide) and in the Gulf of Finland (without oxygen depletion in the bottom water), see Figure 11:2.6. The lack of methylated arsenic species in the bottom water also indicates that methylation of arsenic by bacteria at the sediment-water interface is not probable. The occurrence of MAA and DMAA is predominantly a result of the activity of phytoplankton.

nM 0 I 2 0 2 4 6 8 10 12 (« <) T4~r

8Y26

Figure 11:2.6. Concentration profilesof different arsenic species in the Bornholm Basin (BY5) and in the Gulf of Finland (BY26), in June 1981. Asi = total dissolved inorganic As; As( = total dissolved arsenic (after Andreae and Froelich, 1984).

It is not clear to what extent arsenite is remobilized at the sediment- water interface under natural, undisturbed conditions. Andreae (1979) has shown that arsenite in sedimentary pore waters may diffuse into the overlying water column, but other investigators (e g Carpenter et al.,

45 1978) have failed to demonstrate the existence of any significant arsenic flux from the sediments. The high arsenite concentrations near the sediment in the Bornholm Basin (BY5) are probably a result of in situ reduction of arsenate (Andreae and Froelich, 1984). Arsenite concentrations in the anoxic bottom water in the Gotland Deep (Figure 11:2.2), rich in hydrogen sulfide, are very high, but also arsenate occurs under these conditions, which is thermo-dynamically unexpected. The explanation might be that arsenate may be present in the form of thiocomplexes (thioarsenates), as suggested by Cotton and Wilkenson (1972). Evidence for the formation of thioarsenates when sulfide is added to seawater has been presented (Bertine and Lee, 1983).

In summary, the predominant arsenic species in surface waters of the Baltic Sea, arsenate, is readily available for and taken up by phytoplankton and macroalgae. Part of the arsenic taken up and methylated by the algae is released in the form of arsenite, MAA and DMAA. Part of the organic arsenic is brought by the dead algal cells to the bottom water, where it is remobilized from the detritic material and shows up mainly as arsenite. Arsenite is a relatively unstable species in the oxic surface water layers, but is stable in the anoxic bottom water. Thus, pelagic invertebrates and fish residing in the surface waters are exposed to DMAA and MAA (in addition to arsenate), in particular during and soon after algal blooms. The same might be true for benthic animals in the macroalgal zone of the littoral. Benthic animals residing on deep, oxic bottoms may be exposed to arsenite during the main part of the year.

2.5 Uptake in organisms and bioaccumulation The previously described model ecosystem experiments, in which Baltic Sea littoral ecosystems were continuously exposed to arsenate (and transformation products) for about one year, have provided a great amount of data on the bioaccumulation of arsenic (Notini et al., 1987). When exposed to an arsenate concentration of 75 pg As/1, the dominant phytal component of the ecosystem, Fucus vesiculosus, reached a steady- state concentration in its tissues of 140 pg As/1, already after 2 days exposure. However, this arsenate level was lethal to the brown alga, causing complete elimination after 5-7 months. On exposure to 8 pg As/1, the steady-state level in the algal tissue was 60-80 pg As/g d.w. (Figure 11:2.7).

46 150 - High dese

Low dose

Control

Day

Figure 11:2. 7. Total arsenic concentrations in tissues of Fucus vesiculosus, exposed to arsenate at 75 pgAs/l (X), 8 pgAs/l (O), and 0.5 pgAs/l (O) in brackish-water littoral mesocosms (after Notini et ah,

The green macroalga Cladophora reached arsenic concentrations of 260 pg/g d.w. in winter, but only about 50 pg/g d.w. in summer, when exposed to 75 pg As/1, added as arsenate.

The fraction of organic arsenic in Fucus was >95% of total arsenic, under natural conditions, but only about 60% of the total, when the alga was exposed to 75 pg As/1, indicating that the ability of the alga to synthesize organic arsenic compounds might reach a saturation level at increasing exposure (Blanck et al., 1989).

The time required for arsenate-exposed blue mussels, Mytilus edulis, to reach a steady-state concentration of arsenic (45 pg/g d.w.) in its soft tissues was about 3 weeks (Notini et al., 1987). However, the body- burden of arsenic in the mussles showed strong seasonal variations with considerable increases during the spring phytoplankton bloom and the

47 maximum growth period of the macroalgae, Figure 11:2.8. The speciation of arsenic in the water of the mesocosms was not determined, but it may be assumed that the sharp increase in the body-burden of arsenic in the mussels during spring reflects an increase in organic arsenic in food particles (e g phytoplankton) as well as in the water. This is supported by the observation that the other filter feeder or herbivorous species in the model ecosystems (Cardium and the crustacean Idothea) also exhibited a similar seasonal variation in arsenic content, with spring peaks. Furthermore, the deposit-feeding bivalve Macoma balthica and the carnivore polychaete Nereis diversicolor both showed maximum arsenic concentrations in their body tissues at a somewhat later stage, in June- July, after sedimentation of particles, probably containing organic arsenic.

75 jug As/L

0.5 /jg As/L

Figure 11:2.9. Seasonal variation in total arsenic (pg/g d.w.) in the soft tissues of Mytilus edulis exposed to arsenic (added as arsenate) in brackish-water littoral mesocosms (after Notini et a!., 1987).

The importance of the food route for arsenic bioaccumulation was further demonstrated in a 4-week experiment with the herbivore/ omnivore genera Lymnea and Gammarus. When arsenate was supplied only via the water, no increase in arsenic body burden of the animals took place, but when arsenic-laden algae were added, the animals reached arsenic concentrations in the soft tissues comparable to the levels of long-term exposed animals in complete model ecosystems (Blanck et al., 1989).

48 The composition of total arsenic in the soft tissues of three mollusc species after exposure to arsenate for one year in the model ecosystems was determined by means of the HC1 distillation method. The inorganic, HCl-distillable fraction was generally <20% of the total arsenic content in Lymnea, Macoma and Cardium, but increased with higher arsenate exposure in Cardium, see Figure 11:2.10.

Organic As

Inorganic As

OS 8 75.5 /ugAs/l Cardium Macoma Lymnea

Figure 11:2.10. Concentrations of inorganic and organic arsenic (jxg/g d.w.) in soft tissues of three mollusc species after one year arsenate exposure (after Blanch et al. 1989)

When the benthic, soft bottom crustacean Monoporeia affinis was exposed to arsenate in microcosms, it reached a steady-state in 220 days. The arsenic concentrations in the animals at exposure to 100 pg As/1 was about twice the control level of 4 mg As/g d.w. However, at simultaneous exposure to arsenate and ferric iron, the bioaccumulation was about double that when only arsenate was administered. In the same experiment, it was found that the bivalve Macoma balthica accumulated about five times more arsenic than M. affinis, with a tendency for higher uptake in smaller animals (Sundelin, ref. in Blanck et al., 1989).

The observations made in the macrocosm and microcosm experiments with Baltic Sea organisms were largely confirmed in studies of the relative importance of different pathways of arsenic bioaccumulation (via the water or the food-chain) in estuarine animals in the Chesapeake Bay (Sanders et al., 1989). The uptake of arsenic from water and from

49 phyto-plankton was followed in the copepod Eurytemora affinis, the barnacle Balanus improvisus and the oyster Crassostrea virginica. It was found that dissolved arsenic was readily taken up by phytoplankton and by shell material of the barnacle and the oyster. However, no dissolved arsenic, at concentrations up to 56 pg/1, was incorporated into the soft tissues of the animals during exposure periods of 3-4 weeks. When fed phytoplankton containing elevated arsenic contents, all three animals accumulated significant amounts of arsenic. Juvenile barnacles, mainly feeding on phytoplankton, accumulated relatively more arsenic than adults.

Compared to the 100-200% increase in arsenic content by phytoplankton exposed to dissolved arsenic, the 25-50% increase in the investigated invertebrate species via trophic transfer was relatively small (Sanders et al., 1989). Thus, it may be concluded not only that dissolved inorganic arsenic is relatively unavailable to estuarine invertebrates and vertebrates, but also that there is no food-chain magnification of arsenic in estuarine ecosystems. For many organisms, the arsenic uptake rates will decline throughout the individual's development.

The possible biotic interactions and the pathways of arsenic uptake in a simplified estaurine ecosystem are represented in Figure 11:2.11, based on the results obtained by Sanders et al. (1989). The trophic pathway of arsenic uptake is the major one affecting higher levels of the ecosystem. As a result of the relatively low efficiency of arsenic bioaccumulation from the food, the potential impacts from elevated arsenic levels in the water are usually not important to trophic levels other than phyto­ plankton. However, arsenic can also be incorporated into eggs and be passed on to the next generation. Thus, under conditions of chronic exposure, the second generation may start at higher arsenic concentration and then, the youngest stages feeding on phytoplankton, may accumulate arsenic more efficiently. For example, in experiments with copepods, it was found that second-generation individuals showed a decreased ability to develop to adults when cultured under conditions of chronic arsenic exposure (Sanders et al., 1988). In conclusion, it is obvious that bioaccumulation and the potential for impact of arsenic can be extremely complex in an estuarine ecosystem.

50 DIRECT UPTAKE TROPHIC TRANSFER

NON-FEEDING EGGS, LARVAE

SHELL DEPOSITION- ^ ADULT r BARNACLES / BARNACLES

DISSOLVED ZOOPLANKTON. ARSENATE PHYTOPLANKTON INCLUDING LARVAE

SHELL DEPOSITION- JUVENILE OYSTERS OYSTERS

ADULT OYSTERS

Figure 11:2.11. Pathways of arsenic uptake in a simple estuarine ecosystem. Thickness of lines denotes the relative amount of arsenic incorporated (after Sanders et al, 1989).

2.6 Natural factors affecting speciation and bioavailability The data presented in the previous sections clearly show that the thermodynamically most stable form of dissolved arsenic in brackish water, i.e. arsenate, is readily available to aquatic plants, both phytoplankton and attached micro- and macroalgae, because of its chemical similarity to phosphate (Sanders and Windom, 1980). On the other hand, dissolved arsenate has a low bioavailability to invertebrates and vertebrates. However, arsenate can be readily transformed under the influence of various natural factors. The most important factor in productive estuaries or brackish water seas, such as the Baltic Sea, seems to be the activity of the plant community, i.e. the primary production of the attached macroalgae in coastal or shallow areas, and the phyto­ plankton production in the open sea.

All examined phytoplankton communities have the capacity of producing and releasing reduced arsenic, arsenite, as well as a variety of methylated arsenicals into the water column, but the rate of reduction can vary. The rate and extent of arsenate reduction and methylation apperas to be dependent upon the concentration of arsenate, the dominant phytoplankton species present, the season, and the degree of decline in phosphate concentrations during bloom development (Sanders and Riedel, 1993). For example, in the Chesapeak Bay, arsenite - a relatively unstable chemical form - was present in significant quantities in the surface waters only in late spring. Significant concentrations of MAA

51 and DMAA coincided with blooms of certain phytoplankton species. The MAA detected in natural systems may be a degradation product of DMAA. However, the latter compound is more stable than arsenite, and may persist for some time in the water column. Arsenate and arsenite are more readily sorbed to sediment and suspended particles than DMAA, and may therefore be more rapidly transported to the bottom, together with the organic arsenic compounds incorporated in dead algal cells.

In the anoxic bottom waters, arsenite usually is the predominant dissolved arsenic species, and in waters containing hydrogen sulfide, dissolved thioarsenates may also occur. Release of arsenic from sediments to bottom water, in general, seems to be of minor importance for the arsenic cycle.

In the pelagic communities, during periods without any major phytoplankton blooms, arsenic is almost exclusively taken up by invertebrates and vertebrates by food-chain transfer, most efficiently from phytoplankton to herbivores. No food-chain biomagnification has been demonstrated. The benthic communities may be exposed to dissolved arsenic species such as arsenite and DMAA/ MAA, which might be more bioavailable than arsenate. Benthic invertebrates may therefore take up dissolved arsenic in addition to the complex organic arsenic compounds contained in their food.

The main risk of arsenic exposure of pelagic, brackish-water animals occurs during dense phytoplankton blooms involving specific algal species, when concentrations of DMAA may be high in the euphotic zone. Diatom blooms, induced at relatively high phosphate concentrations in the water, apparently do not produce high concentrations of methylated arsenicals. On the other hand, blooms of the cryptophyte Chroomonas sp. and the dinoflagellate Katodinium rotundatum have been shown to be associated with significant production of dissolved methylated arsenicals (Sanders et al., 1994). The latter phytoplankton species started its bloom in the winter, at low temperature and low phosphate concentration (3 pg P/1), but high nitrogen concentration (140-210 pg N/l) (Sanders and Riedel, 1993).

If we translate these data to the conditions in the Baltic Sea, it might be assumed that the important spring blooms of diatoms (in March-May) do not generate large quantities of reduced or methylated arsenic species, while phytoplankton blooms later in the year, involving other algal groups, may produce significant amounts of arsenite and DMAA.

52 Unfortunately, the phytoplankton species involved in the high production of methylated arsenicals in the Gulf of Finland, in June 1981 (Andreae and Froelich, 1984), were not reported. Nor is it known how efficient cyanobacteria, e g Nodularia sp., which often bloom during the summer and autumn in the Baltic Sea, are in reducing and methylating arsenate.

3. Freshwater Environment (Lakes and Rivers)

3.1 Arsenic speciation in water and sediments 3.1.1 Lake and river waters While arsenic speciation has been thoroughly studied in marine ecosystems, and also relatively well investigated in the brackish-water environment, comparatively less data are available from inland waters. However, as far as the dissolved arsenic species in the water column of lakes are concerned, a series of studies have been carried out, mainly in the U S. and Japan (see Maeda, 1994, for references).

The main difference between freshwaters and marine/ brackish-waters that influences the speciation of arsenic is that in freshwaters, the pH and the content of humic substances seem to play a relatively more important role. The importance of pH for the speciation of inorganic arsenic is clearly demonstrated in the classical Eh-pH diagram, presented by Ferguson and Gavis (1972), Figure 11:3.1. As can be seen in the diagram, the dominating inorganic species in aerobic waters at pH close to neutrality is arsenate, which may be reduced to arsenite both abiotically and biotically, at lower Eh values. Increased acidity enhances the formation of arsenite, while high pH and high oxygen content favour the oxidation to arsenate. Sulfides of arsenic - both solid and dissolved species - predominate under reducing conditions in the presence of reduced forms of sulfur (Ferguson and Gavis, 1972). In highly anaerobic environments, volatile arsines may be formed (Woolson, 1977).

53 H,AsO,

HAsO* - H.AsO,

i - HAsS-

-0.25

-0.50

-0.75

Figure 11:3.1 Eh-pH diagram for inorganic arsenic at 25 °C and 1 atm. Solid species are enclosed in parentheses in cross-hatched areas (from Ferguson and Gavis, 1972).

Arsenate forms strong complexes with dissolved humic material, the stability of which increase with decreasing pH. Thus, in humic and acid freshwaters, the mobility (but not the bioavaila-bility) of arsenic may be higher than what would be expected from pureley thermodynamic considerations (Reuther, 1989). On the other hand, arsenate can form insoluble precipitates with calcium, iron and aluminium compounds in natural water, although several of these reactions are slow in nucleation and have slow growth rates (Lemmo et ah, 1983). It is generally assumed that iron is a possible candidate for controlling dissolved arsenate in natural freshwaters. However, since the formation of arsenic precipitates with iron is so slow, any dissolved arsenic species are more likely to be adsorbed on the surface of inorganic, suspended particles (Blanck et ah, 1989).

54 In uncontaminated freshwaters, the concentration of total dissolved arsenic is generally lower than in estuarine and marine waters. This has been clearly demonstrated in studies where the total dissolved arsenic has been followed in estuaries of big rivers, from areas with a salinity of 0%o to areas influenced by seawater with a salinity of about 30%o. In such estuaries in Canada, the arsenic concentrations gradually increased from 0.08 to 0.5 pg As/1 in the freshwater area to about 1.4 fig As/1 in the marine area (Tremblay and Gobeil, 1990). However, there are many examples of high to very high levels of total dissolved arsenic in some inland waters, where the geological conditions result in very high arsenic leaching. Among the highest levels of dissolved arsenic in uncontaminated river water, so far recorded, were found in River Mauri in the Bolivian Altiplano (about 1,000 pg As/1) as well as in the lower reaches of River Desaguadero, draining Lake Titicaca, where levels between 360 and 580 pg As/1 were repeatedly recorded (PPO, 1996).

The natural background concentrations of total arsenic in the water of Swedish lakes and streams are shown in Table 11:3.1.

Area Average (jug/l) Range (jug/l) Reference Lakes SW Sweden 0.10-0.36 0,10-0,36 Borg, 1984 Lakes, Norrland 0,22 0,11-0,40 Borg, 1984 Lakes, Norrland 0,25 0,06-1,20 Bjorklund et al., 1982 Streams, Norrland 0,2-0,4 Landstrom and Wenner, 1965 Svartalven, W 0,27 Borg, 1984 county Storan, Z county 0,08 Borg, 1984 Petikan, AC 1,7 Borg, 1984 county Table 11:3.1 Background levels of total arsenic in naturalfreshwaters in Sweden.

In small forest lakes in the area of influence of atmospheric fallout, in the vicinity of the large copper and lead smelter, Ronnskarsverken, in northern Sweden, the arsenic concentrations in the lake water were (in 1984) >5 pg As/1 wihtin 10 km from the source, and most of the lakes within 40 km had >2 pg As/1, while lakes situated at more than 100 km distance from the smelter usually had <0.5 pg As/1 (Rosen and Lithner,

55 1986). During a sampling campaign in August 1984, water samples from 10 lakes in the Ronnskar area were fractionated by means of filtration (0.4 pm) and in situ dialysis (0.002 pm). Arsenic in the filtrate varied between 61% and 100%, and the dialyzable arsenic ranged from 35% to 94% (mean 57%) of the amount found in unfiltered samples (Borg, 1986). The dialyzable fraction, which is supposed to be directly bioavailable, varied as a function of the pH value and the content of humic material. It was highest in lakes with low humus content and high (>7) pH, and lowest in a lake with a pH-value of 5.9 and very high humus content.

The distribution between inorganic and organic species of dissolved arsenic in freshwater seems to be very much the same as in brackish- water, and the same kind of natural factors appears to affect speciation. The general picture emerging from several studies in lakes is that the biological activity (phytoplankton production) in the surface euphotic zone during summer is responsible for the formation of dissolved methylated arsenicals, such as MAA and DMAA (Maeda, 1994). For example, in lakes in California, the methylated species represented 1- 59% of the total arsenic in the lake water, with DMAA reaching high concentrations only in the epilimnion, but MAA showing almost uniform concentrations throughout the whole water column (Anderson and Bruland, 1991). During the winter, when complete mixing of the lakes occurred, the arsenate concentration was restored at the expense of the methylated forms.

3.1.2 Freshwater sediments Inorganic arsenic species predominate in freshwater sediments, but traces of MAA and DMAA are generally present, usually with higher levels of MAA than of DMAA, the latter being opposite to the situation in the water column (Maeda, 1994). It is therefore assumed that DMAA is demethylated to form MAA by sediment microorganisms. Both DMAA and MAA have been shown to be demethylated by anaerobic river sediments (Holm et al., 1979).

In the interstitial water of freshwater sediments, the arsenic concentration was highest at pH values <4 and >9. It was suggested that the mobilization of arsenic at low pH values is due to the dissolution of hydrous oxides of manganese and iron, while the mobilization at high pH values most probably depends on a competitive ligand exchange between arsenate and phosphate (Clement and Faust, 1981). The competition between arsenate and phosphate for sorption sites on suspended and

56 sediment particles was also observed in experiments carried out by Reuther (1992), indicating that phosphorus-poor sediments would be a more efficient sink for arsenate than phosphorus-enriched ones.

The sorptive behaviour of inorganic arsenic has been the subject of many studies (see e g Gupta and Chen, 1978). Arsenate reaches adsorption maxima for iron, manganese and aluminium hydroxides, and clay minerals, at pH values between 4 and 7, while adsorption of arsenite increases with pH, reaching a maximum at pH around 9. Thus, there are several mechanisms contributing to the immobilization and incorporation of arsenic into the bottom sediments of lakes and ponds, where it remains at least as long as the overlying water remains aerobic (Maeda, 1994).

3.2 Arseniclevels and speciation in biota Much less is known both about the total levels and the distribution of arsenic species in freshwater biota than in biota from marine and brackish-water ecosystems. However, available data seem to indicate that arsenic levels in both plants and animals from uncontaminated fresh­ water environments, in general, are lower than those found in waters with higher salinity.

In oligotrophic lake water (conductivity = 3.4 mS/m; phosphate-P = 3-4 pg/1; arsenic = 0.1-1.0 pg/1; and pH = 6.2-6.S), the total arsenic content in leaves and roots of the aquatic plant Lobelia dortmanna was 0.2-0.3 pg/g d.w., in the soft tissues of the mussel Anodonta cygnea, about 4 pg/g d.w., and in the hepatopancreas of crayfish (Astacus astacus), 3 pg/g d.w. (Reuther, 1992). The aquatic moss, Fontinalis spp., sampled from uncontaminated lakes in northern Sweden, contained 1-3 pg As/g d.w., but in mining areas, elevated levels of arsenic (11^40 pg/g d.w.) were detected in the tissues of the moss (Bjorklund et al., 1982). The iron content in the lake water was found to strongly influence the arsenic level of the moss (Blanck et al., 1989).

Zooplankton sampled in various lakes in the Ronnskar area in northern Sweden, where the water contained 0.4—3.7 pg As/1, also showed relatively low concentrations of total arsenic, 2.3-4.8 pg/g d.w., but in a lake with 32 pg As/1 in the water, the zooplankton exhibited an elevated level of 12 pg As/g d.w. (Blanck et al., 1989).

Fish from the North American Great Lakes are reported to contain 0.03- 0.12 pg As/g d.w. (Traversy et al., 1975; Pillay et al., 1973), while in

57 small, uncontaminated forest lakes in northern Sweden, muscle tissue from roach, perch and pike was reported to contain 0.06-0.09 pg As/g d.w. (Norin and Vahter, 1984). This is less than half the concentration of total arsenic in fish (perch and pike) from the Bothnian Bay (cf. section 2.2), and one to several magnitudes of order less than what is usually found in marine fish (cf. section 1.2.3). Perch from the arsenic contaminated lakes in the Ronnskar area were found to contain between 0.1 and 1.4 pg As/g d.w. in the muscle tissue, and these levels correlated well with the concentration of arsenic in the water of lakes, where the pH was >5.8. Arsenic in the liver of the same fish varied with the season, showing the highest levels in summer (Blanck et al., 1989).

The information is very scarce about the occurrence of organic arsenic species in freshwater organisms, including fish. However, DMAA, trimethylarsine oxide, arsenocholine and traces of arsenobetaine were identified in fish from lakes in the Ronnskar area (Norin and Vahter, 1984; Norin et al., 1985b). Unfortunately, due to the low to veiy low levels of total arsenic in these samples, it was not possible to quantify the various organic arsenic species. However, the fraction of inorganic arsenic in most of the fish species was determined to be 5-12% of the total arsenic content, and this fraction tended to decrease with increasing total arsenic concentrations (Norin et al., 1985b). Thus, in spite of the lack of confirming data, there is little reason to believe that the composition of the arsenic in freshwater fish should be radically different from the composition found in marine and brackish-water fish.

3.3 Biotransformation of arsenic While the biotransformation of arsenic compounds by organisms in marine ecosystems has been thoroughly investigated (see section 11:1.3), only few experimental or field studies have been made to elucidate the biotransformation of arsenic in the freshwater environment. Most of these few studies were conducted by Maeda and co-workers (Maeda et al., 1990; 1992a; b; Maeda, 1994; Kuroiwa et al., 1994; 1995). Unfortunately, no chemical identification of the methylated trans­ formation products was carried out; the organic arsenicals were only characterized as monomethylated, dimethylated and trimethylated arsenic compounds, respectively. No other studies have been found in the literature, where conclusive data on the identity of the organic arsenicals formed by freshwater organisms are presented. Therefore, it is still not possible to draw any definite conclusions about the possible discrepancies between the metabolic cycle of arsenic in marine and

58 freshwater environments. However, available studies provide some clues regarding the similarities, as will be shown in the following.

Already in the beginning of the 1970s, Lunde (1972) found that various freshwater algae contained arsenic in the lipid phase, suggesting that this arsenic was organically bound. Later, Nissen and Benson (1982) exposed the blue-green alga Rhizoclonium sp. to 74As-arsenate for one week and then extracted the cells with ethanol. It was found that almost all the arsenic in the cells was present as lipid- or water-soluble "lipid-related" compounds, which had chromatographic characteristics very similar to those extracted from marine algae. Similar treatment of charophytes, water ferns and higher plants revealed that the fraction of "lipid-related" arsenicals was somewhat smaller, but always at least 60% of the total (Cullen and Reimer, 1989). These results indicate that arsenate taken up by various limnetic aquatic plants may be transformed into arsinylribosides.

Maeda et al. (1992a) showed that the green alga Chlorella vulgaris, when exposed to MAA, the cells contained both mono- and dimethylated arsenic, while when exposed to either DMAA or arsenobetaine, the cells contained only dimethylated or trimethylated arsenic, respectively. Thus, this alga was capable of biomethylating MAA (but not DMAA), but was incapable of demethylating any of the methylarsenicals.

In a food-chain experiment, where arsenic was accumulated via the food from Chlorella to the herbivore Moina sp. and further to the carnivore Poecilia sp. (guppy), almost all the arsenic in the algae was in inorganic form, but in the herbivore some 15% was in dimethylated form, and in the fish about 85% was in the trimethylated form (Maeda et al., 1992b).

In a similar food-chain experiment, including Chlorella, the freshwater shrimp Neocaridina denticulata and the killifish (Oryzias latipes), Kuroiwa et al. (1994) measured not only the amount and form of arsenic in the different trophic levels, but also the amount of arsenic excreted into the water by the shrimps when exposed to each of the following compounds: arsenate, MAA, DMAA and arsenobetaine.

The algae in the food-chain experiment were cultured in the presence of arsenate, but contained, after one week, 11% of methylated arsenicals. The shrimps received arsenic for 7 days, only via the algae and contained 39% methylated arsenic, mainly dimethylated. Finally, the fish, after 7 days feeding with shrimp powder, contained 80% of trimethylated

59 arsenic. The concentration of total arsenic in the organisms decreased by more than an order of magnitude for each trophic level, but steady-state was probably not achieved.

The accumulation and excretion experiments indicated that both methylation and demethylation occurred within the shrimps, and that the excreta contained both inorganic arsenic and mono-, di- and trimethy- lated compounds, regardless of the dosed compound.

To summarize the experiments dealing with freshwater biotransformation of arsenic, it is clear that, in spite of the incomplete and scattered data, the occurrence of biomethylation of arsenic has been proven. The fact that great variations have been found in the capacity of freshwater algae and higher plants to methylate arsenic indicates that the difference between species in this respect might be greater in the fresh­ water than in the marine environment. Whether or not the trimethylated arsenic which predominates in freshwater fish is arsenobetaine, like in marine fish, has not been proven, but at least traces of arsenobetaine have been positively identified in some species of freshwater fish (Norin et al., 1985b).

3.4 Sinks, mobilization and bioavailability of arsenic species As mentioned in section 3.1.2, the fate of arsenic introduced into the water column of a lake depends on the presence of precipitating and sorbing agents, such as hydrated iron and aluminium oxides, and on the affinity for complexation with inorganic and organic ligands, e.g. amino acids, humic constituents and sulfide ions (cf. Reuther, 1986).

However, the most efficient mechanism for stripping arsenic from the water column probably is related to the incorporation of arsenic into phytoplankton and the subsequent removal by sedimentation from surface waters of lakes. The efficiency of this mechanism for the removal of arsenic from the water column was shown to depend on both the concentration of arsenic and phosphate in the water (Reuther, 1992). In large mesocosms continuously dosed with arsenate (final concentra­ tions were 5 and 50 pg As/1) for 65 days, without and with simultaneous addition of phosphate (final concentrations, 3 and 5 pg P/1), the amount and the arsenic concentration of the settling material was measured. It was found that the low arsenate dose without phosphate addition generated about twice the amount of sediment as was formed with phosphate addition, and in the latter case, the arsenic concentration was

60 twice as high (Table 11:3.2). This result may be explained by an increased mortality of the phytoplankton in the absence of phosphate addition, which generated a higher volume of settling material. In the high arsenate dose, this effect was much stronger: a very high sedimenta­ tion of dead algal cells at low, but almost no sedimentation at somewhat higher phosphate concentration. The small amount of settling material in the latter case with very high arsenic content can be assumed to consist of dead cells of arsenic-tolerant algal species that were able to survive long enough to accumulate high amounts of arsenic due to the higher phosphate concentration (Reuther, 1992).

No. Treatments Settling material Arsenate (pgAs/l) Phosphate (ug P/l) Dry weight (g) Tot. As (pg/g) 1. 0,1-1,0 3 1,8 1,5 2. 5,0 3 3,6 4,4 3. 5,0 5 2,0 10,6 4. 50 3 4,4 40,5 5. 50 5 0,5 76,0 Table 11:3.2 Amount of settling material and its arsenic content in five flow-through mesocosms exposedfor 65 days to different concentrations of arsenate and phosphate (after Reuther, 1992).

Leaching rates of arsenic from CCA-treated wood has been measured, for 7 days, in water with different salinity (0, 6 and 12 %o, respectively). It was found that the leaching rate in the freshwater was more than double (70 pg/cm2) that recorded in the brackish-water test (about 30 pg/cm2) (Sanders et al., 1994).

The bioavailability of the various arsenic species to organisms at different trophic levels has been studied in several uptake experiments in the laboratory. Unfortunately, these experiments have, in general, been conducted at high to very high exposure concentrations, and for rather short exposure times, which make the results difficult to extrapolate to the natural environment. However, some conclusions on the relative bioavailability of arsenicals to different organisms may be justified, based on these experiments.

Marin et al. (1992) investigated the relative bioavilability of different arsenic species to rice (Oryza sativa), grown in nutrient solutions, inter alia containing 10 mg P/1, enriched with arsenic compounds at concentrations of 50,200 and 800 pg As/1. The bioavailability of arsenic to rice followed the trend DMAA < arsenate < MAA < arsenite, regardless of the concentration added, Figure 11:3.2. Upon absorption, 61 DMAA was readily translocated to the shoot, while the other three arsenic species largely accumulated in the roots (Marin et al., 1992).

0.05 0.2 0.8 Arsenic concentration in nutrient solution (mg / L)

Figure 11:3.2. Arsenic concentrations in shoots of rice, cultivatedfor 4 weeks in nutrient solutions containing different forms and concentrations of arsenic (after Marin et al., 1992).

Maeda et al. (1992a) compared the availability and uptake of four arsenic species to the green alga Chlorella vulgaris at nominal concentrations as high as 10 mg As/1. After 7 days, the cell content of total arsenic was about 5 times higher, when arsenate was added, than when either MAA, DMAA or arsenobetaine was added. The three methylated species had the same degree of bioavailability to the alga.

62 200 -

As(V) concentration in water (jig As/mL) DMAA concentration in water (jig As/mL)

2000

ee 1000

1 s 10 MAA concentration in water (jig As/mL) AB concentration in water (jig As/mL)

Figure 11:3.3. Total arsenic accumulated by the shrimp Neocaridina denticulata from the water phase containing one of four different arsenic species ( Kuroiwa et al., 1994)

In a similar experiment, the shrimp Neocaridina denticulata was exposed for 7 days to the same four arsenic forms, dissolved in the water. At an exposure concentration of 1 mg As/1, arsenate and MAA were equally bioavailable, while DMAA was less available for uptake (Figure 11:3.3). Arsenobetaine was taken up most efficiently (Kuroiwa et al., 1994), but this compound usually does not occur dissolved in water at elevated concentrations in natural waters.

Experiments with another shrimp, Macrobrachium rosenbergii), which were conducted at extremely high arsenic concentrations, confirmed the above results. The availability of arsenate and MAA to the shrimps was equally high, but DMAA was taken up at a rate about 20% of the rate determined for MAA (Kuroiwa et al., 1995).

63 Finally, dissolved arsenate, occurring in high concentrations, is readily bioavailable to killifish (Oryzias latipes), which accumulates arsenic mainly in inorganic form, but 20-40% of the total arsenic in the fish was recovered as methylated species, mainly monomethylated (Kuroiwa et al., 1994).

Probably the main factor, in natural limnetic systems, determining the bioavilability of the predominant arsenic species in lake water, arsenate, to the most important organisms for the flux of arsenic into the food web, is the actual concentration of phosphate. At low ambient concentrations of both arsenate and phosphate, the complex interaction between the two ions has important consequences both for the flux of arsenic into cells and for the effects on populations and communities. These aspects will be further discussed in section 111:3.

3.5 Uptake in organisms and bioaccumulation Although there are great differences between different algal species in their capacity to accumulate arsenic from the water phase (Andreae and Klumpp, 1979), it appears that there are no systematic differences in this respect between freshwater and marine algae. Lunde (1973) cultivated three species of freshwater and three species of marine unicellular algae in freshwater and saltwater media, respectively, containing radioactive arsenic ions. Arsenic enrichment in the freshwater algae was 240 to 2,800 times the concentration in the medium, while in the marine algae, the enrichment was 710 to 2,900 times. Thus, freshwater algae seem to have the same capacity to concentrate arsenic from the water phase as the marine algae.

A culture of the green alga Chlorella vulgaris was isolated from a site with high natural content of arsenic, and was found to tolerate arsenate levels in the medium up to 10 g/1 (Maeda, 1994). When this culture was grown in a medium containing 100 mg As/1, the arsenic level in the cells reached 20 mg/g d.w., and when grown in the highest concentration tolerated, 10 g As/1, a level of 50 mg/g d.w. was recorded. Maeda and coworkers also investigated the role of pre-acclimation to arsenate of Chlorella cultures on the degree of arsenic accumulation of the cells. It was found that the higher the pre-acclimation concentration, the higher was the degree of arsenic bioaccumulation in the cells at the stationary growth phase.

64 No arsenate was bioaccumulated by Chlorella cells that had been pretreated with the respiratory inhibitor dinitrophenol or with heat, while only little effect of pretreatment with the photo-synthesis inhibitor sodium azide on arsenic bioaccumulation was observed (Maeda, 1994). Phosphate competitively inhibited arsenic accumulation by C. vulgaris, and in some other algae, such as Chlamydomonas sp. and Phormidium sp., arsenate reduced phosphate uptake. The cyanobacterium Synechococcus leopoliensis was found to be almost insensitive to arsenate, showed low arsenate uptake, and high discrimination for phosphate (Planas and Healey, 1978; Budd and Craig, 1981; Maeda et al., 1988).

In field studies in a series of lakes, contaminated with arsenic from mining effluents, the bioconcentration in various vascular plants was investigated (Dushenko et al., 1995). It was found that bioaccumulation of arsenic usually was much higher in submerged plants, such as Potamogeton pectinatus and Myriophyllum sp. than in emergent species, such as Typha latifolia. Concentration factors (plant tissue/sediment) were in the range of 1.3-3.4 in the first category of plants, while the average concentration factor for Typha latifolia shoots was only 0.04, and for roots, 0.59. It is well known that the two first mentioned plant species primarily take up nutrients from the water phase. No relationship could be demonstrated between tissue concentrations of arsenic and environmental concentrations of phosphorus. It was concluded that the lack of arsenate/ phosphate interaction in the studied system might be due to the relatively high levels of arsenic, caused by the mining effluent discharges (200-500 pg As/1 in the water column), and that such interactions might only be observed at low, natural arsenate concentrations.

Also in the controlled mesocosm experiments (designed as shown in Table 11:3.2) run at relatively low arsenate concentrations (5 and 50 pgAs/1), no competitive inhibition of arsenic bioaccumulation in plants caused by phosphate was observed (Reuther, 1992). On the contrary, somewhat increased phosphate concentrations, in basins 3 and 5 (see Figure 11:3.4), resulted in higher arsenic accumulation, accompanied with enhanced biomass of both leaves and roots of Lobelia dortmanna.. The stimulation of arsenic bioaccumulation in the leaves of the plant by the addition of small amounts of phosphate to the water was quite conspicuous. On the other hand, phosphate addition did not seem to stimulate the accumulation of arsenic in periphyton (Figure 11:3.4).

65 ppm dw

250

zee

150 As ppm dw 100

50

0 1 has in

Figure 11:3.4. Bioaccumulation of arsenic in leaves of Lobelia dortmanna (upper part) and in periphyton (below) as a function of arsenate and phosphate concentrations. For explanation of treatments, see Table 11:3.2 (after Reuther, 1992).

Based on the many freshwater food-chain experiments carried out by Maeda and coworkers (e g Maeda et al., 1990; 1992b; Kuroiwa et al., 1994; and review in Maeda, 1994), it can be concluded that the main route for bioaccumulation of arsenic in higher trophic levels of freshwater ecosystems is via the food, i e much the same as has been

66 observed in marine and brackish-water ecosystems. It is generally found that organisms at lower trophic levels (unicellular algae as well as aquatic plants) have a greater ability to accumulate arsenic than the herbivorous or carnivorous organisms. Thus, arsenic is not biomagnified in the freshwater food-chain, which is in striking contrast to the situation for e g mercury.

Biomethylation of arsenic by freshwater organisms has been experimentally proved, but the exact nature or the chemical structure of the organic arsenic compounds present in the living cells of freshwater organisms still have to be revealed.

3.6 Natural factors affecting speciation and bioavailability As mentioned before, many observations support the existence of an interaction between phosphate and arsenate in the water phase, controlling the bioavilability and the uptake of arsenate into the cells of aquatic plants. However, results presented by different researchers are conflicting: in some cases, a competitive inhibition of arsenate uptake by phosphate has been reported, while in other cases, phosphate seems to have a stimulatory effect on the uptake and bioaccumulation of arsenic. There are also some reports indicating that no relationship at all should occur between phosphate and arsenate. These conflicting results will be further discussed in Section III of this report, where effects of arsenic exposure are described, because the assessment needs to take into account the interaction of arsenate with the phosphorus metabolism of cells, including effects on photosynthesis, respiration and other parts of the energy metabolism.

In spite of the conflicting observations described in previous sections, it appears that - in one way or another - the amount of phosphate available to the aquatic plants determines the fate of arsenate introduced to a freshwater aquatic system. Since the phosphate concentration can show extremely great variations in freshwater ecosystems, both between lakes or rivers and between seasons as well as between different water layers and sediment strata, it is pertinent to devote a great interest to the possible interactions between phosphate and arsenate in the assessment of the environmental risks of arsenic.

In freshwater environments, other important natural factors determining speciation and bioavailability of arsenic are the content of humic material and the occurence of hydrous iron, manganese and aluminium oxides in the sediment as well as in the water column. The adsorption of

67 arsenic to sediments, suspended solid particles and humic material is related to the pH and the redox-potential. Once adsorbed, arsenicals are not easily removed (Mok and Wai, 1994). Arsenate and arsenite differ in adsorption characteristics: arsenate is charged and adsorbed over a wide pH range, whereas arsenite is not, which is consistent with the greater mobility of arsenite. The increase in mobility of arsenate under more reducing conditions is generally attributed to the reduction of Fe(III) to Fe(II), with subsequent release of arsenate, and to the reduction of arsenate to arsenite. The arsenic-humic acid interactions may, at certain pH values, especially at low pH, be as important as adsorption to hydrous oxides. Increased acidity during sediment-water interactions will result in decreased stability of the iron oxy-hydroxides, and this changes the binding efficiency for arsenicals, thus mobilizing the arsenic (Mok and Wai, 1994). However, in acid lakes with a high content of humic material, the released arsenic is most probably complexed by humic acid, resulting in an increased mobility of the arsenic, but without any notable increase in its bioavailability.

It may be impossible to propose a general model for arsenic biogeochemistry. Different ecosystems have specific blends of both geochemical and biological controls over arsenic speciation, and contain different types of biological communities. Once a given system has been described, the patterns of arsenic speciation are explainable and potential impacts can be identified, but they cannot necessarily be transferred to another system. However, the continuing accumulation of information concerning different types of systems does lead to a greater overall understanding of arsenic biogeochemistry. It is clear that successful predictions will require not only a background understanding of arsenic biogeochemistry, but also a thorough understanding of the trophic relationships of the particular system under examination (Sanders et al., 1994).

4. Water Quality Characteristics in Inland European Water Bodies of Relevance for Arsenic Speciation Because of the chemical similarity between arsenate and phosphate, the water quality characteristic that is most important for the uptake and transformation of arsenic in living cells is phosphorus. Concentrations of total phosphorus in lakes and rivers vary over a very broad range, depending on the geological background and the type of soils existing in the catchments, but even more important for determining the actual phosphorus concentration are the type and the scope of human activities: land use, amount of phosphate fertilizers and manure used in agriculture,

68 degree of phosphorus removal in sewage treatment, type of industrial effluents and degree of treatment.

As a part of the large OECD Project on Eutrophication of Waters, the concentration of phosphorus in more than 100 lakes and reservoirs in Europe and North America was measured, mainly during the 1970s. Based on these data, the mean annual concentration of total phosphorus in 115 lakes and reservoirs has been reported as 47 pg P/1 (range: 3-750 pg P/1) and the mean annual concentration of phosphate in 99 lakes and reservoirs as 16 pg P/1 (range: 0.2-890 pg P/1) (OECD, 1982). Although some 35 North American lakes were included in the data base, the average values are quite representative for European lakes and reservoirs. The highest values of phosphorus are usually found in shallow lakes in the plain districts of Central Europe. In these areas, the average phosphorus concentrations in lake water are much higher than the average concentrations in the Scandinavian lakes. However, the results of the OECD programme on lakes and reservoirs in Europe and North America have shown that, in most cases, phosphorus is the limiting factor which determines the development of eutrophication (OECD, 1982).

Generally speaking, in Europe, the situation is the following: in low-land areas where a high percentage of land is used for agricultural purposes, and the population density is high, the highest concentrations of phosphorus are found in relatively shallow lakes, which therefore, tend to be eutrophic. In mountainous, hilly and forested areas of Europe with low degree of agricultural exploitation, the oligotrophic lakes are found with low levels of dissolved phosphorus in the water. In some forested areas, especially in Finland and Sweden, the lakes furthermore contain high amounts of humic material dissolved in the water.

Below are given a few examples of the general levels of phosphorus in some great lakes in Europe. Data are given as concentrations of total phosphorus. This way of reporting phosphorus concentrations in lake water is more justified than giving concentrations of reactive phosphorus or phosphate, because the turnover of phosphorus in the lake water is quite rapid, mainly due to the activity of bacterioplankton. Therefore, a speciation of phosphorus at any given point in time is usually not meaningful, at least not for the present purposes. The data given below are compiled from Eurostat (1994).

69 Lake Ysselmeer (NL) 200 gg P/1 Lake Geneva (Leman) 52 gg P/1 Lake Peipus 40 gg P/1 Lake Ladoga 30 gg P/1 Lake Malar (Granfjarden) 30 gg P/1 Lake Vanem 10ggP/l Lake Mjosa (N) 7 gg P/1 Lake Vattem 5 gg P/1

In small forest lakes in Scandinavia, the phosphorus concentrations are generally in the range 5-10 gg P/1, in some cases <5 gg P/1.

Organisms, such as algae, that rely upon dissolved phosphate at low concentrations in the water as their major source of phosphorus are particularly susceptible to simultaneous exposure to the chemical analogue arsenate. Interference with phosphate uptake at the cell surface and in phosphorylation reactions wihtin the cell is generally supposed to be the main mode of toxic action of arsenate. Furthermore, higher concentrations of phosphorus in the water seem to counteract the toxicity of arsenate to several aquatic plants. Another characteristic of lake water that might interfere with the availability and toxicity of arsenate is its content of humic material. Arsenate tends to adsorb or bind to humic material, complexes which are quite stable at low pH values. Increasing levels of humic material in the lake water tend to decrease the availability and toxicity of arsenate to phytoplankton, but also the impoverished light climate in brown-water lakes may reduce the uptake of arsenate by algae, due to the lower rate of photosynthesis.

Therefore, the most critical conditions with respect to bioavailability, uptake and toxicity of arsenate in the freshwater environment seem to occur in small, dear-water lakes which are oligotrophic (low phosphorus concentration), i e where there might be a strong competition between arsenate and phosphate for uptake in algae.

70 EFFECTS OF VARIOUS ARSENIC SPECIES IN THE AQUATIC ENVIRONMENT

1. Marine Environment

1.1 Effects on microorganisms 1.1.1 Sensitivity to arsenate The sensitivity of different species of marine microalgae (phytoplankton and periphyton) to arsenate varies over several orders of magnitude (Blanck et ah, 1984). The toxicity of arsenate does not only vary with the algal species or the algal phylum, but also with the concentration of phosphate in the medium. The interactions between phosphate and arsenate will be discussed in further detail in chapter 111:4. In this section, only some general considerations about the mechanisms determining the sensitivity of algae to arsenate will be mentioned.

It is generally assumed that phytoplankton communities in the open sea are dominated by algal species with high affinity for nutrients, and are adapted for the low and stable nutrient concentrations usually found in open oceans (Sanders and Riedel, 1987). Phytoplankton species that have a high affinity for phosphate appear to be better able to discriminate between the necessary nutrient ion and the competing arsenate ion than do species exhibiting lower nutrient affinity. Increased discrimination should result in reduced arsenate uptake. However, even if marine algae, in general, might be relatively efficient in keeping down the uptake of arsenate, the differential sensitivity to arsenate exposure may have important consequences for species succession within a natural phyto­ plankton community (Sanders and Cibik, 1985).

There are two possible strategies by which algal species in a community may be successful and dominate after exposure to arsenate stress:

(i) In order to be successful, a species may have a lower incorporation rate of arsenate because of a higher degree of specificity for the uptake of phosphate. Although slower growing, these algal forms would not be susceptible to increased concentrations of arsenate and would be able to dominate over more rapidly growing cells that were incorporating larger amounts of arsenate.

71 (ii)Some algal species are very efficient at transforming arsenate into less toxic, methylated forms, which are either incorporated into the cell or released to the water, other species are not. By using the strategy of transforming the arsenate, these species may be able to dominate. Because the transformations require energy, cell growth may be somewhat inhibited, but the species can continue to exist within the community at lowered cell densities (Sanders and Riedel, 1987).

When cultures of the centric diatom Skeletonema costatum were exposed to arsenate at low phosphate concentrations (1.6-2.5 pg P/1), the arsenate concentration causing a 50% growth inhibition was found to be 5.0 pg As/1 (Sanders, 1979). When the phosphate concentration was greater than 9.3 pg P/1, the growth rate was not affected by small additions of arsenate. However, although enough phosphate was present to reduce the arsenate toxicity, the algal population was still hampered by an increased requirement for phosphate. This could reduce the species' ability to compete with more resistant species.

Sanders and co-workers have tested a number of other marine micro­ algae for sensitivity to arsenate (see Sanders and Riedel, 1987, for references). It was found that the most sensitive species were the diatom Rhizosolenia fragilissima and the chrysophyte Isochrysis galbana, which suffered a 50% reduction in growth rate at arsenate additions of 2.3 pg As/1 in low-phosphate seawater. Other diatoms, two dinoflagellates and one chlorophyte species were found to respond to arsenate concen­ trations in the range 6 to >98 pg As/1.

The relative sensitivity of different marine phytoplankton groups was investigated in natural assemblages cultivated outdoors in large-volume cultures for 15 days (Sanders and Vermersch, 1982). Three series with the following arsenate concentrations were used: control: 1.1 pg As/1, low dose: 7.7 pg As/1 and high dose: 20 pg As/1, in seawater with the ambient phosphate level of 15.5 pg P/1. The total cell numbers were reduced, compared to the control, by 57 and 67% in the low and the high dose, respectively, and the chlorophyll a concentration was reduced by 30% in the treated cultures. The relative response of the different groups of algae to the enhanced arsenate exposure is shown in Figure 111:1.1.

72 levels groups dose; stress. Figure between Blanck periphyton OF TOTAL CELL NUMBER

60

C of

"Centrics" III: of and

=

8 phosphate -

algae high and 1.1

communities Wangberg

Changes 20

dose over 4

and pg

(<3

(after time As/1

:

in 8 pg

to

initiated in relative

Sanders P/1), arsenate

large-volume DAYS

12 studied

exposure

refer

abundance a

73 and

stress. restructuring OF

the

to

Vermersch,

STRESS

to

diatoms.

cultures sensitivity It

arsenate

was

of

the

found of

CENTRICS A DINOFLAGELLATES OTHER PENNATES exposed

1982). various

the

concentrations = of KEY

control;

that, marine community,

FLAGELLATES

to

taxonomic at

arsenate

low B

=

low

because some species were eliminated already at this low arsenate exposure. Various responses were examined, and the results, given as IC20 values, are presented in Table 111:1.1.

Test parameter IC20 (PgAs/l) Species composition 15 Chlorophyll a accumulation 23 Total carbon accumulation 38 Total nitrogen accumulation 60 Photosynthesis per unit area 23 Table I1T.1.1. Sensitivity of marine periphytoncommunities to arsenate. The communities were established during 3 weeks at a phosphate concentration of <3 pg/l. All effects are expressed as IC20 values, i e the arsenate concentration required to produce a 20% change in each parameter, compared to the control (after Blanch and Wangberg, jpagg).

When the periphyton communities were established under continuous arsenate stress (750 pg As/1), they developed a tremendous increase in tolerance to arsenate. The pre-exposed communities were >16,000 times more tolerant to arsenate than control communities established at background levels of arsenate. This increase in tolerance (PICT = Pollution-Induced Community Tolerance) was obviously caused by arsenate exerting a strong selection pressure on the algae, thus restructuring the community and eliminating the arsenate-sensitive components (Blanck and Wangberg, 1991). It is important to note that this selection starts to eliminate the most sensitive components already at low arsenate levels (about 8-15 pg As/1).

No data on the toxicity of arsenate to marine bacteria have been encountered, but it might be assumed that bacteria do not belong to the most sensitive taxa with regard to arsenate exposure.

1.1.2 Sensitivity to other dissolved arsenic species The relative composition of the dissolved arsenic, i e the proportion of reduced and methylated arsenic species, in the water phase is important for the response of the ecosystem as a whole to arsenic exposure. For many marine organisms, with the exception of algae, it is generally assumed (based on laboratory tests) that arsenite and simple methylated arsenic species (MAA and DMAA) are more toxic than arsenate. Therefore, it is important to assess the overall effects of the release, by growing algae, of arsenite, DMAA and MAA (by at least some species).

74 Since arsenite is unstable under oxidizing conditions and is rapidly reconverted to arsenate, while both MAA and DMAA are stable in marine systems, it is most important to evaluate the toxicity of the methylated species to various marine organisms, including micro-algae. Unfortunately, reports dealing with the toxicity of these species are scarce.

Sanders (1979) carried out short-term toxicity tests with Skeletonema costatum, which were exposed to either arsenite or DMAA at low phosphate concentrations. He found that small addtions of arsenite (0.98 pg As/1) caused a slight inhibition of growth, although an addition of 20 pg As/1 resulted in almost the same inhibition. He concluded that arsenate and arsenite caused inhibition at similar levels, and also noted that it is difficult to know exactly whether the algal cells were exposed to arsenite or arsenate, because of the rapid oxidation of arsenite. In contrast to these results, DMAA caused no growth inhibition at a concentration of 9.8 pg As/1, and even at 26 pg As/1; this arsenic form had little or no effect on phytoplankton productivity. The reason for the non-toxicity of DMAA was proposed to be its chemical dissimilarity with phosphate and that it is not actively taken up by the algal cells (Sanders, 1979).

Also Blanck et al. (1989) report that DMAA was non-toxic to all the algal communities they tested, and that the EC50 value for inhibition of photosynthesis was >75 mg As/1. However, MAA was foud to be considerably more toxic, with a EC50 value close to that for arsenate, or 5 pg As/1. At low nutrient concentrations, arsenite was found to be at least one order of magnitude less toxic than arsenate to all communities tested, and the toxicity of arsenite did not show any obvious correlation to the phosphorus concentration.

Thus, it appears that although arsenite is generally considered to be more toxic than arsenate to higher marine life, arsenate has a more profound effect on the growth and other cell functions of marine algae than does arsenite. It has been speculated that marine algae erect a barrier against the absorption of arsenite (Eisler, 1994), much the same way as they exclude DMAA.

Only scattered data have been found on the sensitivity of marine bacteria to arsenite and methylated arsenic species. However, when bacterial cultures from the Sargasso Sea and from a site off the US east coast were grown in arsenite-enriched media, the bacteria reduced all available

75 arsenate and utilized the arsenite during the exponential growth phase, presumably as an essential trace nutrient (Johnson, 1972).

1.2 Effects on macroalgae and other plants The red macroalga Champia parvula was exposed to various concen­ trations of arsenate and arsenite in seawater and the toxic responses were recorded (Thursby and Steele, 1984). On exposure to arsenate, concentrations as high as 10 mg As/1 allowed normal growth, but the sexual reproduction of the alga was impaired. The reproduction was normal at arsenite concentrations in the range 65-95 pg As/1, but at 300 pg As/1, the alga did not survive.

The development of sporelings of another red macroalga, Plumaria elegans, was arrested after short-term exposure to arsenite concentrations of 580 pg As/1 (Eisler, 1994).

These marine macroalgae appear to be far more tolerant to arsenic exposure than macroalgae in the brackish-water environment (see section 111:2.2), but available information is limited and no general conclusions can be drawn on the relative sensitivity of marine macroalgae to arsenic exposure.

1.3 Effects on invertebrates Generally speaking, invertebrates appear to be quite resistant to dissolved arsenic, which is in contrast with the often high sensitivity of algae. Concentrations necessary to cause inhibition or death of invertebrates, 100 - 1,000 pg As/1, are very much higher than those found in natural marine systems (Sanders et al., 1994).

Results from conventional toxicity tests using endpoints such as survival of adults, juveniles or larvae, condition index or fecundity, usually show that direct effects of dissolved arsenate or arsenite occur only at concentrations >100 pg As/1, as can be inferred from the summary of toxicity data compiled by Eisler (1994), see Table 111:1.2.

No data have been found from toxicity tests examining the effects of direct exposure of marine invertebrates to MMA or DMAA, dissolved in the water. On the other hand, the main pathway of arsenic to inverte­ brates is through the food, and therefore, they are predominantly exposed to complex, organic arsenic compounds, such as dimethylated arsinylribosides, arsenocholine and arsenobetaine.

76 Taxon Svecies As (us/l) Effect O o

Copepod Eurytemora affinis I Reduced juvenile survival Copepod Acartia clausi As(III): 510 LC50 (96 hr) Crab Cancer magister As(V): 230 Larvae, LC50 (96 hr) Oyster Crassostrea gigas As(III): 330 Embryo, LC50 (96 hr) Mysid Mysidopsis bahia As(III): 630-1,270 MATC, lifecycle test Mysid d:o As(V): 2,300 LC50 (96 hr) Table 111:1.2. Toxicity data for arsenite and arsenate, selectedfrom tests with the most sensitive marine invertebrates (after Eisler, 1994)

As indicated in Table 111:1.2, when studying the conventional endpoints, marine invertebrates can be considered as remarkably resistant to arsenic. However, when attention is directed towards biological mechanisms at the cellular level, changes may be observed. This was clearly indicated in oysters (Crassostrea gigas) exposed to arsenate at a concentration of 10 pg As/1. Both the gill epithelium and the tissues of the digestive tract showed cytological abnormalities, such as structural alterations of mitochondria and nuclei, suggesting a disturbance of both the cellural respiratory metabolism and nucleotide incorporation (Ettajani et al., 1996). However, it was not possible to decide if these modifications were pathological or adaptive.

1.4 Effects on marine fish When testing the toxicity of arsenic to marine fish, most interest has been devoted to arsenite. Available data indicate that various fish species are quite resistant to short-term exposure to arsenite, the LC50 values being in the range 3.8-27 mg As/1 (Eisler, 1994). Such levels of arsenite do not occur in the natural marine environment. Thus, it may be concluded that arsenic dissolved in the seawater does not constitute a risk for fish.

As was the case with invertebrates, also fish are predominantly exposed to arsenic via the food, i e most of the arsenic intake is in the form of organic arsenic compounds. Marine teleosts seem to be unaffected at arsenic levels in the muscle tissue of more than 100 pg As/g fresh weight. Most of this arsenic is presumably in the form of non-toxic arsenobetaine.

77 1.5 Effects at the ecosystem level From a purely toxicity perspective, the most sensitive link in the marine ecosystem is the phytoplankton. Some phytoplankton species can suffer from growth reduction at arsenate concentrations as low as 3 pg As/1, and concentrations of 5-10 pg As/1 have been shown to cause significant reduction in growth as well as shifts in phytoplankton species composition, because of the different sensitivity of various phytoplankton species (Sanders and Vermersch, 1982; Sanders et al., 1994; Blanck and Wangberg, 1988a).

The relative tolerance of higher trophic levels, described in the previous sections, does not necessarily mean that these levels are free from potential harmful effects. Arsenic may be used as an example of a contaminant that has its primary effect at the base of the food chain. This primary effect can have cascading, indirect impacts upon higher trophic levels (Sanders et al., 1994). The potential effects on higher trophic levels may not be limited to arsenic transfer through feeding. Also rapid changes in phytoplankton species dominance and in community structure can produce significant effects on the grazing community, if the herbivores are feeding selectively with respect to phytoplankton species, size or shape. Different types of food can have quite different nutritive value. Therefore, the assessment of effects of arsenic cannot be limited to one single trophic level; the ecosystem must be considered in its entirety.

The existence of such integrated effects can be illustrated by the results of several experiments, in which phytoplankton communities were exposed to chronic, low-level arsenic doses (Sanders and Cibik, 1988; Sanders et al., 1988; 1991; 1994). A common result in these experiments was a replacement of larger, arsenic-sensitive centric diatoms with smaller, more resistant species, both other diatoms and flagellated algae, yielding a community dominated by smaller phytoplankton cells. For example, at arsenate concentrations of 3 pg As/1 and higher, the growth of the large, centric diatom Cerataulina pelagica was greatly inhibited, and it was replaced by the smaller arsenate-tolerant Thalassiosira pseudonana. In subsequent feeding experiments with the copepods Acartia tonsa and Eurytemora affinis, it was found that when the diet was based on smaller, arsenate-resistant algae, the fecundity of the copepods was reduced, relative to the case when the diet consisted of larger, arsenate-sensitive species (see Figure III: 1.2. Fecundity was reduced even more when the copepods were fed on resistant species, that had been exposed to high arsenate concentrations, and therefore had an

78 elevated arsenic content. Thus, although no direct effect of low arsenic exposure could be detected on the copepods, an indirect effect was observed, due to the arsenate-induced shift in phyto-plankton species composition.

#--# Resistant + As #-# Resistant A A Sensitive 150-

100-

DURATION (days)

Figure 111:1.2. Fecundity of Acartia tonsa, fed different mixtures of phytoplankton species. The sensitive treatment means feeding on algae sensitive to arsenate, the resistant treatment = feeding on arsenate- resistant species, and resistant + As means feeding on resistant species that had accumulated high levels of arsenic. All treatments received the same amount offood, based on carbon content of cells (after Sanders et al, 1994).

However, not all herbivores respond in the same way. For example, rotifers and tintinnids exhibited higher growth rates when fed on the arsenate-exposed phytoplankton community than when fed on the unaltered community. Also larger herbivores, such as barnacles and oysters, are essentially unaffected by arsenate dosing (Sanders et al., 1988). It may therefore be concluded that herbivores which feed on-

79 larger species of phytoplankton may be hampered by the arsenate- induced shift in the phytoplankton community structure, while micro­ zooplankton which thrive on small phytoplankton cells react in an opposite way and herbivores that can feed on a wide range of particle sizes (barnacles and oysters) do not react to shifts in the size of dominant phytoplankton (Sanders et al., 1994).

1.6 Arsenic criteria for protection of marine life Only a few countries have established water quality criteria for arsenic in the marine environment, although many countries have issued criteria for freshwater. These criteria, usually expressed as the highest allowable concentration of dissolved arsenic in the water phase, should not be exceeded if harm to marine life is to be avoided. However, the data base available to those who have developed such criteria apparently varies from country to country and so do the political considerations that have influenced the criteria-setting.

The most stringent criteria are those proposed by Norway (in Sweden no arsenic criteria for marine waters exist). The main objective of the Norwegian system of water quality criteria is to provide protection for salmon, and is built upon a classification of water areas into four quality classes. The different classes allow somewhat different types of utilization or provide different level of protection. The criteria for total arsenic are the following (Lindestrom, 1992):

Class 1: <2 pg As/1 No harm to any organisms and no limitations in using fish as human food. Class 2: 2-6 pg As/1 Sensitive organisms may be harmed, but no problem for fish. Only small effects on communities. Class 3: 6-20 pg As/1 Harmful effects on salmon, reduced communities with tolerant species dominating. Class 4: >20 pg As/1 Unacceptable conditions for salmon. Large damages on communities.

Marine water quality criteria in the U.K. include total arsenic, and refer to the annual mean in samples that have been filtered through a 0.45 pm filter. The criterion value is 25 pg As/1 (Lindestrom, 1992).

In the USA, the Environmental Protection Agency (EPA) has issued limit values for many metals, including arsenic in seawater. However,

80 the criteria are explicitly given for arsenite, because it was considered that data were insufficient for criteria formulation for arsenate or for any organoarsenical (EPA, 1985a). Two different criteria have been established for arsenite:

(i) The four-day average not to exceed more than once every 3 years = 36 pg As/1; (ii) The one-hour mean not to exceed more than once every 3 years = 69 pg As/1.

It might be noted that it is highly improbable that such high concen­ trations of arsenite would ever occur in marine waters, at least if the water contains dissolved oxygen at levels that will allow marine fauna to survive.

In conclusion, the only seawater quality criteria, among the three sets presented above, that are of any practical use are those proposed by the Norwegian authorities. They seem to be based on a realistic assessment of the potential risks associated with long-term exposure to low levels of total arsenic in the marine environment. They also seem to provide sufficient protection - in the case of Class 1 criterion - to the most sensitive among the marine organisms.

2. Estaurine or Brackish Water Environment

2.1 Effects on microorganisms In the southern part of the Baltic Sea, the toxicity of arsenate to a periphyton community was measured in terms of a 50% inhibition of the photosynthesis. The effect concentration was determined at 6 pg As/1 at phosphate concentrations varying between 1 and 5 pg P/1 (Blanck et al., 1989).

No direct measuresments of arsenate toxicity to phytoplankton in natural brackish-water from the Baltic Sea have been found in the literature. However, in laboratory cultures of Scenedesmus obliquus, grown in a diluted, synthetic seawater medium of l%o salinity, the toxicity of arsenate and arsenite was determined at different phosphate concentrations (Hofslagare et al., 1994). At concentrations of arsenate or arsenite less than 7.5 pg As/1, no toxic effects were observed. Increasing the arsenate concentration to 75 pg As/1 with no or medium (31 pg P/1) phosphate addition resulted in a 22-34% reduction in carbon uptake of

81 the cells, but this inhibition was eliminated at phosphate levels of 310 pg P/1 and higher. Arsenite at a concentration of 75 pg/1 did not reduce the carbon uptake, but arsenite levels in the range 380-750 pg As/1 yielded an 11% reduction of the carbon uptake. Thus, arsenite was clearly less toxic to this green microalga than was arsenate (Hofslagare et al., 1994).

2.2 Effects on macroalgae and other plants The mesocosm experiments described in sections 11:2.2 and 11:2.4, with model ecosystems simulating the Baltic Sea littoral zone, also allowed the toxicity of arsenate to the brown alga Fucus vesiculosus to be determined. After one year of arsenate exposure in brackish-water, a nominal arsenate concentration of 8 pg As/1 yielded a 50% reduction in Fucus biomass, while exposure to 75 pg As/1 resulted in death and complete elimination of this alga (Blanck et al., 1989). The tissue concentrations of arsenic in algae exposed to 8 pg As/1 were in the range of 70-80 pg As/g, and in algae grown at background levels of arsenate (0.5 pg As/1), the tissue concentrations were about 20-30 pg As/g.

The apical growth of Fucus vesiculosus was measured in relation to the tissue content of arsenic, and was found to decline with higher content of arsenic in the tissues (Figure 111:2.1). This relationship was then used as a basis for an assessment of the possible inhibiting effect of arsenic in the field. In a coastal zone of the southern Baltic Sea, close to the outlet from a municipal sewage plant, samples of Fucus were collected and the growth over the last year was measured together with the content of arsenic in the tissues (Blanck et al., 1989). It was found that the apical growth was inhibited at tissue levels of arsenic from about 40-50 pg As/g and upwards (Figure 111:2.1).

Also the rate of phosphate uptake in Fucus was affected by elevated concentrations of arsenate in the water. In short-term experiments, the phosphate uptake decreased to 40% of the control rate at arsenate concentrations higher than 20 pg As/1, and at 100 pg As/1, the rate of P- uptake went down to 15% of the control rate (Rosemarin, ref. in Blanck et al., 1989).

In the mesocosms, the elimination of the Fucus community in the high arsenate dose opened up this niche for oppurtunist species that can survive the arsenic stress. An invasion of Spirogyra resulted in a dense spring bloom in the high-dose mesocosm, which was replaced, in the summer, by a variety of green and red macroalgae species as well as by a

82 periphyton community consisting of bluegreen and diatom microalgae (Blanck et al., 1989). Thus, among the macroalgae, the brown algae, and particularly the keystone species Fucus vesiculosus, appeared to be the most arsenate-sensitive components, being inhibited already at low levels of arsenate exposure, i e at concentrations below 8 pg As/1.

200 1

Zone affected by sewage treatment plant effluent

Fucus tissue arsenic ( /jg/g dw ) (winter max)

Figure 111:2.1. Apical growth (mm/year) of Fucus vesiculosus as a function of the arsenic content in tissues (fig/g dry weight). Data from the field (o) compared to data from mesocosm experiments (x) (from Blanck et al, 1989).

In studies of the macroalgae communities that were able to grow on CCA-treated wood in brackish-water, large differences in community composition were found between the treated wood and control wood.

83 The most tolerant species were the red alga Ceramium sp., the green algae Ulva lactuca and Enteromorpha intestinalis, while other species, such as Ectocarpus sp., Polysiphonia sp. and Cladophora sp., usually did not settle on the CCA-treated wood (Weis and Weis, 1992). However, it is not possible to discriminate between the effect of the three metal salts (copper, chromium and arsenic) and tell which one had the greatest inhibiting effect on the algae.

Nonetheless, it has been clearly shown that low concentrations of arsenate in brackish-water environments have a structuring influence on the macroalgae communities, an effect which is exerted by the difference in sensitivity between different species of algae. In particular, the brown algae group seems to be very sensitive to arsenate exposure, among which Fucus vesiculosus, which is - by far - the most important structuring component of the littoral ecosystem in the Baltic Sea, is severely affected at arsenate concentrations in the water as low as 8 pg As/1 or maybe even lower (Notini et al., 1987). 2.3 Effects on invertebrates It has generally been found that estaurine invertebrates are remarkably resistant to dissolved arsenic. Usually, the arsenic concentrations must exceed 100 to 1,000 pg As/1 before significant direct impact occurs (Sanders et al., 1988). This resistance probably results from the relatively low availability of dissolved inorganic arsenic to invertebrates. Furthermore, the incorporation of complex organic arsenic species into invertebrates also seems to be relatively low, e.g. arsenobetaine does not seem to be available for assimilation and is eliminated unmetabolized (Sanders et al., 1989). Thus, the potential for direct impact from arsenic ingested via food is not likely to be significant in most estuarine or coastal systems.

The relatively low degree of incorporation of arsenic compounds into herbivores, and the apparent absence of food-chain magnification suggests that for many organisms, the arsenic-uptake rates will decline throughout an individual's development. For example, smaller individuals principally feed on phytoplankton, which may result in high arsenic uptake rates, but later, when these animals switch to feeding on zooplankton, the arsenic uptake rates should decline (Sanders et al., 1989). This declining exposure would also result in a reduced risk for detrimental effects.

84 In the previously described mesocosm experiments, not even the high- dose arsenic exposure (75 pg As/1) resulted in any reduction of the total biomass of invertebrates, although the Fucus habitat disappeared from the mesocosms (Notini and Rosemarin, ref. in Blanck et al., 1989). However, different groups of invertebrates reacted in different ways. Both gastropods and crustaceans declined in total biomass (Figure 111:2.2), while the biomass of Mytilus edulis almost doubled in the high dose. The most conspicuous shift in the invertebrate community was a decrease in herbivore biomass and a simultaneous increase in detritovore and filter-feeder biomass. This probably was an indirect effect of the drastic shift in habitat and ecosystem structure, rather than a direct toxic effect of arsenic.

Sediment Wall

g Stone Fucus

yug As/l /ug As/L

Figure 111:2.2. Total biomass (g dry weight per pool) of the gastropods Lymnea sp. and Theodoxus sp. (left) and of crustaceans (right) in different habitats of brackish-water littoral mesocosms, exposed to arsenate for one year (after Notini and Rosemarin, ref. in Blanck et al, 798(0.

85 2.4 Effects on estuarine fish The mesocosms referred to above also contained a few specimens of fish: stickleback and flounder. The two levels of exposure to arsenate, 8 and 75 pg As/1 had no deleterious effects on these two fish species. The total fish biomass in the high-dose pool was the same as in the controls, after one year of exposure (Blanck et al., 1989). However, the growth rate of the stickleback was lower, which was compensated by a greater number of individuals. The reduced individual growth rate of the new generation of fish in the high-dose pool might be explained by the reduction in biomass of prey organisms (invertebrates). No similar effect on the growth of flounders, mainly feeding on bivalves in the sediment, was observed. The biomass of bivalves was not reduced.

2.5 Effects at the ecosystem level Exposure of a brackish-water littoral ecosystem to arsenate (8 or 75 pg As/1) for one year produced major changes in the structure of the ecosystem, due to toxic effects on the keystone alga Fucus vesiculosus. These changes had many features in common with the community shifts caused by eutrophication and consisted in a major shift in the composition of primary producers: from Fucus to short-lived, opportunistic filamentous algae. The habitat-structuring Fucus, strongly dominant on rock or stone substrates in shallow parts of the Baltic Sea, normally functions as a nutrient store, providing a slow but steady flow of carbon, nitrogen and phosphorus to higher trophic levels. Filamentous algae, on the other hand, have a short life-span and a very much faster nutrient turnover. When these algae become dominant, e g due to inhibition or elimination of the arsenate-sensitive Fucus, the result is an increased channeling of material through the detritus pathway. In the experiments, this was observed as a greater content of organic matter in the sediment and as a shift from herbivore to detritivore invertebrates, but no change in the total invertebrate biomass over a period of one year.

In the long term, the disappearance of the Fucus habitat from a natural bay of the Baltic Sea will probably result in a drastic impoverishment of the invertebrate and the fish fauna. Under normal circumstances, many species of invertebrates and fish migrate to the Fucus belt for feeding and reproduction, and use this habitat as a nursery and sheltering area (Blanck et al., 1989).

Thus, also in the brackish-water environment, the sensitivity to arsenic greatly differs between organisms and trophic levels. Even between

86 primary producers, the range of sensitivity to produce toxic responses on exposure to arsenic is very large. However, it turns out that the keystone species Fucus vesiculosus, which is the totally dominant biotic element of littoral ecosystems in the Baltic Sea, is particularly sensitive to arsenate exposure. Therefore, even at low concentrations of arsenate in the water column, the whole system begins to restructure, causing a whole array of cascading effects on higher trophic levels, effects that - in many respects - resemble those occurring as a result of eutrophication of coastal waters.

3. Freshwater Environment (Lakes and Rivers)

3.1 Effects on microorganisms Laboratory tests of the toxicity of arsenate and arsenite to various freshwater phytoplankton species have pointed out the great variability in sensitivity between species. For example, in continuous-culture experiments with Asterionella formosa, arsenate levels of up to 160 pg As/1 neither caused reduced growth rate nor affected the uptake of nutrients (Conway, 1978). In tests of five species of freshwater phytoplankton in the presence of arsenate, 75 pg As/1 depressed the growth rate of two of the species, while a third species required 750 pg As/1 for the same degree of depression, and the two remaining species were unaffected at concen-trations up to 7,500 pg As/1 (Planas and Healey, 1978). Other workers have reported depression of the growth rate of Chlamydomonas at an arsenate concentration of 75 pg As/1 (Christensen and Zielski, 1980) and of Scenedesmus obliquus at an arsenate level of 48 pg As/1 (Eisler, 1994).

These highly variable results, even among species within the same phylogenetic group, indicate that the effects of arsenate on freshwater phytoplankton vary not only between species, but also within species, depending on both the physiological status of the cells at the start of the experiment, and the supply of nutrients in the culture medium (Blomqvist and Heyman, 1990).

In similar laboratory tests with arsenite, the toxicity was usually found to be lower: toxicity values of 1.7-2.3 mg As/1 have been reported for 4 different species (Eisler, 1994).

Relatively few experiments on the effects of arsenate on natural mixed populations of fresh-water phytoplankton have been performed, and even

87 fewer on the response of natural periphyton communities to arsenate exposure. In 1986, phytoplankton and periphyton communities from the mesotrophic brownwater Lake Sormogen were tested with respect to the inhibition of photosynthesis as a result of arsenate exposure (Wangberg et al., 1991). The communities were sampled either directly from the lake or from a (control) limnocorral in the lake without any treatment (total arsenic concentration in the lake water was 1.0-1.6 pg As/1). The results of the tests are summarized in Table 111:3.1.

Source PhytoplanktonPeriphyton Lake, August 19 44 - Lake, August 20 14 - Lake, October 6,0 9,8 Limnocorral, August 44 - Limnocorral, October 9,8 1,5 Table 111:3.1. EC50 values (pgAsfl) for inhibition of photosynthesis by arsenate in phyto-planktonand periphyton communities from Lake Sormogen (after Wangberg et al., 1991).

The difference in toxicity recorded between the two consecutive days in August was not statistically significant. The lowest effect concentration found for phytoplankton was 6.0 pg As/1. The difference between the effect concentrations obtained for periphyton in October has not been explained by the authors.

The main purpose of the experiments in Lake Sormogen was to investigate the effects on the development of the phytoplankton community of low additions of arsenate to various limnocorrals during the whole ice-free season (Blomqvist and Heyman, 1990). Two of the limnocorrals received an arsenate concentration of 14 pg As/1 in the beginning of June, whereupon the arsenate concentration declined to levels between 3 and 6 pg As/1 in July and August. A third limnocorral received an addition of arsenate which gave a peak concentration of 4 pg As/1 in the beginning of June, and after that, the concentration was equal to that in the control limnocorral (1.0-1.6 pg As/1). The phytoplankton biomass declined in the two high-dose limnocorrals to levels between 16 and 28% of the control level in June, and remained at levels between 21 and 74% of the control level throughout the summer months. In the third limnocorral, there was a decline in June to 56% of the control level, but later, the biomass was restored to levels of 81-85% of the control level. The originally very high species diversity was not affected by the

88 arsenate additions. Thus, the results did not show that arsenate had any clear selective effect on the community, since negative effects were found in virtually all common groups and species of algae (Blomqvist and Heyman, 1990).

However, a more detailed investigation of the phytoplankton community from the experiments in Lake Sormogen, carried out by Wangberg et al. (1991), showed both that the arsenate treatments induced a decrease in algal cell volume (the no-effect-concentration being 4.4 pg As/1) and that a slight change in taxonomic composition of the community did occur.

The overall conclusion from these experiments in a brownwater mesotrophic lake is that the lowest concentration of arsenate inducing significant changes in the phytoplankton community, mainly depression of total biomass, but without inducing any major changes in the community structure, was about 4 pg As/1. The low effect concentration (1.5 pg As/1) indicated for inhibition of photosynthesis in a periphyton community must be regarded with precaution, since it could not be replicated with periphyton from the lake itself.

Wangberg (1995) also investigated the phytoplankton communities in eight different lakes being exposed to long-term arsenic contamination through atmospheric fallout originating from the nearby metal smelter in Ronnskar. A complicating factor is that these lakes were also contami­ nated with several other metals, such as cadmium, copper, lead and zinc. Both the number of species, the total cell volume and the possible induction of increased tolerance to arsenic in the phytoplankton communities (PICT) were investigated during three consecutive years. Increased tolerance to arsenate, compared to the communities in four reference lakes, was found only in the two most contaminated lakes, Snesviken (28-48 pg As/1) and Burotjam (about 8 pg As/1). Snesviken also exhibited the lowest number of phytoplankton species and both lakes had higher total cell volumes than other lakes investigated. Snesviken was one of the lakes with the lowest concentration of phosphorus (5-7 pg P/1). Thus, the only lakes where a clear "pollution- induced community tolerance" (PICT) was observed were the lakes having arsenic concentrations in the water mass >8 pg As/1. At least the most contaminated lake had a severe restructuring of the phytoplankton community. However, it is difficult to conclude whether or not the arsenic load on this lake was mainly responsible for these changes or to what extent other metals contributed to the impact.

89 3.2 Effects on aquatic plants The freshwater mesocosm experiments conducted by Reuther (1992) allowed the determination of effects on the aquatic plant Lobelia dortmanna caused by long-term exposure to arsenate with and without phosphate addition (see section 11:3.4). After an exposure time of 65 days, the biomass of Lobelia , exposed to 50 pg As/1, was clearly depressed, but a simultaneous addition of phosphate (to 5 pg P/1) eliminated this inhibition. The average number of Lobelia leaves being lost during the experiment was enhanced by exposure to arsenate levels of both 5 and 50 pg As/1, and - again - the simultaneous addition of phosphate completely eliminated this effect.

Studies of the effect of four different arsenic compounds, dissolved in the water, on the growth and dry matter production of rice were conducted by Marin et al. (1992). The results of these growth tests after a four-week exposure period, are shown in Figure 111:3.1.

c o <3 D T3 O

CO E

*u

DMAA AS (V) As (III) MMAA Arsenic chemical form in solution

Figure III: 3.1. Rice dry matter productionas a function of arsenic concentration and chemical form (after Marin et al., 1992).

90 When arsenic was added as DMAA, at levels of 50 and 200 pg As/1, there was a slight increase in total dry matter production, and no toxic response was observed even at the highest concentration (800 pg As/1). Arsenate did not affect the dry matter production of rice at any of the tested concentrations. However, both arsenite and MAA were clearly phytotoxic to rice. The most phytotoxic compund was MAA, giving a significant effect already at a level of 50 pg As/1. Rice plants grown in the highest concentration of MAA were stunted with necrosis in leaf tips and margins. These symptoms indicated a limitation in the movement of water into the plant resulting in death. The total dry weight of these plants was only 34% as compared to the control (Marin et al., 1992).

In field surveys of arsenic-contaminated lakes, it was found that phytotoxic symptoms (stunted shoot growth and necrosis) generally occurred in Typha latifolia, when the arsenic concentrations in sediment and in water exceeded 300 pg As/g and 400 pg As/1, respectively (Dushenko et al., 1995).

Summing up, it appears that arsenate concentrations of 5 pg As/I, in combination with very low phosphate levels, can produce slight detrimental effects on some freshwater macrophytes. Other macrophytes can tolerate arsenate concentrations up to at least 800 pg As/1. Among the four most common dissolved arsenic compounds in freshwater, MAA seems to be the most phytotoxic one, followed by arsenite, as shown in growth tests with rice.

3.3 Effects on invertebrates A great amount of toxicity data have been reported from laboratory tests with various freshwater invertebrates being exposed to one or several of the common, soluble arsenic compounds (see Eisler, 1994, for overview). However, very few studies have been conducted in which invertebrates have been exposed, via the food, to well specified, complex, organic arsenic compounds. Such di- or tri-methylated arsenicals are, however, generally considered to be of low toxicity.

In this section, only a small sample of toxicity data will be presented and discussed, mainly data from tests with chronic exposure, or data from series of tests in which the same test organism was exposed to different arsenic compounds under comparable experimental conditions.

In Table 111:3.2, some toxicity data on freshwater invertebrates are summarized (Eisler, 1994; Kuroiwa et al., 1994; 1995).

91 Species Compound As cone, jug As/l Type of effect Daphnia magna As(III) 630-1,320 MATC, life-cycle test As(V) 520 Reprod. impairtment of 16% MAA >830 No deaths in 28 days DMAA >1,100 No deaths in 28 days Gammarus As(III) 88 LC20, 28 days pseudolimnaeus As(V) 970 LC20, 28 days MAA 86 LCiQ, 28 days DMAA 850 LCo, 28 days Neocaridina As(V) 1,500 LC50 , 7 days denticulata MAA 10,000 LC50 , 7 days DMAA 40,000 LC50 , 7 days As-betaine 150,000 LC50 , 7 days Macrobrachium As(V) 30,000 LC50 , 5-7 days rosenbergii MAA 100,000 LC50 , 5-7 days DMAA 300,000 LC50 , 5-7 days Table 111:3.2. Toxic effects of various arsenic compounds to freshwater invertebrates.

The most sensitive among the tested organisms was the amphipod G. pseudolimnaeus, which showed mortality at arsenic concentrations <100 pg As/1. Arsenite and MAA were, in this case, equally toxic, but in tests with other organisms, it was generally found that the higher the number of methyl groups in the molecule, the lower the toxicity. When comparing the toxicity of arsenite and arsenate, it turned out that in one case, arsenite was more toxic, while in another case, arsenate was the most toxic form. Generally speaking, however, the most common dissolved arsenic species, including arsenate, are much less toxic to invertebrates than to some algae. Thus, invertebrates cannot be considered as a critical group when assessing the direct toxicity of arsenic to freshwater ecosystems.

3.4 Effects on freshwater fish Also fish is quite tolerant to arsenic dissolved in the water. The maximum acceptable toxicant concentration (MATC) when exposure goes on for the whole life-cycle has been determined for a few fish species. In the case of flagfish (Jordanella floridae), the MATC value for arsenite was 2.1^1.1 mg As/1 (EPA, 1985b). Almost the same range was obained for fathead minnow (Pimephales promelas), when exposed to

92 arsenite: MATC = 2.1-4.8 mg As/1, while arsenate was more toxic to the latter species: MATC = 0.53-1.5 mg As/1 (EPA, 1985b).

The lowest effect levels found for freshwater fish in the literature were 100 pg As/1 as arsenate, which was reported to induce a 15% behavioral impairment in gold fish, and 540 pg As/1 as arsenite, which after 28 days exposure caused 50% mortality in rainbow trout embryos (Eisler, 1994). Arsenic trioxide at a concentration of 300 pg As/1, caused slight biochemical changes in plasma and gills and affected the migratory behavior of coho salmon, but had no effect on growth or survival of the smolts (Nichols et al., 1984).

When rainbow trout were exposed to arsenic compounds via the food, the no-observed effect concentration for dietary arsenate was determined at 13-33 pg As/g diet (Cockell et al., 1992). A dietary dose of arsenate in the range 55-60 pg As/g diet caused gallbladder inflammation and a mild to moderate responsive anemia in the fish. However, rainbow trout fed a diet containing organic arsenic (DMAA or arsenobetaine) for 8 weeks, did not exhibit any toxic responses at doses of 120-1,600 pg As/g diet (Cockell and Hilton, 1988).

3.5 Effects at the ecosystem level Much the same as in the marine and the brackish-water environment, also in freshwater ecosystems it is the primary producers, that are the organisms most susceptible to toxic effects induced by the predominant arsenic species, arsenate. However, the great variability in relative susceptibility to arsenate, between different species and groups of primary producers, found in the two first cases, does not seem to be as conspicuous in the freshwater ecosystems so far examined. Therefore, the typical shift in composition of the phytoplankton community under arsenate stress might not be as common in lakes, and consequently, the cascading, indirect effects on the higher trophic levels may also be less obvious. However, so far, the number of freshwater ecosystems studied in this respect is too small to allow any valid generalisations to be made.

3.6 Arseniccriteria for protection of freshwater aquatic life Generally speaking, it must be concluded that when authorities in various countries have established freshwater quality criteria for arsenic, their way of reasoning has probably not been based on a thorough understanding of the biochemistry and toxicology of arsenic in limnic systems. The arsenic limit values in those countries where such exist, are

93 usually set at 50 (ig As/1 or higher (Lindestrom, 1992). No limit value for arsenic in freshwater has been given by the Norwegian authorities.

In the Netherlands, there exists - in addition to the "standard" limit value for arsenic, 50 pg/1 - a recently proposed, lower value, which represents the "Maximal Acceptable Risk" (MAR) level. This value is based on the 95-percentile of NOEC recordings, and is set at 8.6 pg As/1 (Lindestrom, 1992). This value is not very different from the lowest arsenate levels (4—8 pg As/1), which in Swedish field or mesocosm experiments caused toxic responses to various green plants.

4. Mechanisms of Toxicity - Interaction with Phosphorus As pointed out several times before, arsenic is an element with an active biochemistry. Because of the chemical similarity between arsenate (the predominant inorganic arsenic form in aquatic systems) and phosphate, arsenate is taken up by autotrophic organisms along with phosphate. However, arsenate interferes with the important biochemical functions of phosphate, in particular, uncoupling of oxidative and photo­ phosphorylation may occur by a competitive interaction with phosphate in the formation of ATP (Avron and Jagendorf, 1959). In the presence of arsenate, unstable ADP-arsenate is formed, thus regenerating ADP, but the arsenylated ADP is stable enough for arsenosugars to be formed (Blanck et al., 1989). The rapid regeneration of ADP in the presence of arsenate leads to a consumption of the electro-chemical gradient without any ATP formation (Blanck and Wangberg, 1991).

In the water phase, when the ratio of arsenate to phosphate is relatively high, arsenic toxicity to phytoplankton becomes more likely (Sanders et al., 1994). Presumably, the transformation reactions, by which methylated arsenic species are formed by algae, have evolved to minimize this toxicity by altering arsenate and either eliminating it from the cell in a less toxic form or storing it within the cell in the form of non-toxic, complex molecules.

Arsenate can be transported by phosphate carriers in algal cells, although it has been shown that the phosphate uptake mechanism of algae and cyanobacteria, to some extent, can discriminate between phosphate and arsenate (Blanck and Wangberg, 1991). In addition, arsenate can be reduced to arsenite by algae and cyanobacteria, a compound that also

94 inhibits photosynthesis, but by a completely different mechanism (Blanck and Wangberg, 1991).

Arsenite is also an intermediate in the detoxification pathway leading to the excretion of methylated arsenic acids.

When comparing the relative toxicities of arsenic compounds, it is necessary to consider the modifying effect of phosphate on the toxicity of arsenate. At excess phosphorus concentration, the toxicity of arsenate to algae is insignificant, while at low P/As quota in the cells, arsenate is by far the most toxic among the arsenic species (Blanck et al., 1989). The traditional view that arsenite is the most toxic arsenic species, certainly does not hold for algae, and it may be questioned if it is generally valid for other aquatic organisms. Rather, it seems that there is a great variability in the relative sensitivity of different animal species and taxa to the two inorganic dissolved arsenic forms.Of the two methylated arsenic acids, MAA may have a relatively high toxicity to aquatic organisms other than algae, while DMAA seems to have a low toxicity to all organisms tested. The lack of toxicity of DMAA to algae is consistent with the idea of the methylation pathway as a means of detoxification of arsenate.

95 IV. LITERATURE REFERENCES

Anderson, L.C.D. and Bruland, K.W. (1991) Biogeochemistry of arsenic in natural waters: Importance of methyl species. Environ. Sci. Technol., 25: 420-427.

Andreae, M.O. (1978) Distribution and speciation of arsenic in natural waters and some marine algae. Deep-Sea Res., 25: 391-402.

Andreae, M.O. (1979) Arsenic speciation in seawater and interstitial waters: The influence of biological-chemical interactions on the chemistry of a trace element. Limnol. Oceanogr., 24: 440-452.

Andreae, M.O. (1983) Biotransformation of arsenic in the marine environment. In: Lederer, W.H. and Fensterheim, R.J. (Eds.) Arsenic: Industrial, Biomedical, Environmental Perspectives. Van Nostrand- Reinhold, New York. pp. 378-392.

Andreae, M.O. and Froelich, P.N. (1984) Arsenic, antimony, and germanium biogeochemistry in the Baltic Sea. Tellus, 36B: 101-117.

Andreae, M.O. and Klumpp, D. (1979) Biosynthesis and release of organo-arsenic compounds by marine algae. Environ. Sci. Technol., 13: 738-741.

Avron, M. and Jagendorf, A T. (1959) Evidence concerning the mechanism of adenosinetriphosphate formation by spinach chloroplasts. J. Biol. Chem., 234: 967-972.

Bertine, K.K. and Lee, D.S. (1983) Antimony content and speciation in the water column and interstitial waters of Saanich Inlet. In: Wong C.S. et al. (Eds.) Trace Metals in Sea Water. Plenum Press, New York. pp. 21-38

Bertrand, G. (1902) Sur l'existence de l'arsenic dans la serie animate. C.R. Hebd. Seances Acad. Sci., 135: 809-812.

Bjorklund, I., Borg, H. and Johansson, K. (1982) Environmental impact by airborne mercury and other metals in middle and northern Sweden. Swedish environmental Protection Agency, SNV PM 1568. (In Swedish).

96 Blanck, H. and Wangberg, S.-A. (1988a) The validity of an ecotoxi- cological test system. Short-term and long-term effects of arsenate on marine periphyton communities in laboratory systems. Can. J. Fish. Aquat. Sci., 45: 1807-1815.

Blanck, H. and Wangberg, S.-A. (1988b) Induced community tolerance in marine periphyton established under arsenate stress. Can. J. Fish. Aquat. Sci., 45: 1816-1819.

Blanck, H. and Wangberg, S.-A. (1991) Pattern of cotolerance in marine periphyton communities established under arsenate stress. Aquat. Toxicol., 21: 1-14.

Blanck, H., Wallin, G. and Wangberg, S.-A. (1984) Species dependent variation in algal sensitivity to chemical compounds, Ecotoxicol. Environ. Safety, 8: 339-351.

Blanck, H., Holmgren, K., Landner, L., Norin, H., Notini, M., Rosemarin, A. and Sundelin, B. (1989) Advanced hazard assessment of arsenic in the Swedish environment. In: Landner, L. (Ed.) Chemicals in the Aquatic Environment: Advanced Hazard Assessment. Springer- Verlag, Berlin, Heidelberg, New York. pp. 256-328.

Blomqvist, P. and Heyman, U. (1990) Effects of arsenate additions to pelagic mesocosms on phytoplankton biomass and composition. Limnologica (Berlin), 20(2): 329-338.

Borg, H. (1984) Background levels of trace metals in Swedish fresh- waters. Swedish Environmental Protection Agency, SNV PM 1817. (In Swedish).

Borg, H. (1986) Speciation of metals in lake waters in the Ronnskar area. Swedish Environmental Protection Agency, Report No. 3124. (In Swedish).

Bostrom, K and Valdes, S. (1969) Arsenic in ocean floors. Lithos, 2: 351-360.

97 Bottino, N.R., Newman, R.D., Cox, E.R., Stockton, R., Hoban, M., Zingaro, R.A. and Irgolic, K.J. (1978) The effects of arsenate and arsenite on the growth and morphology of the marine unicellular algae Tetraselmis chui (Chlorophyta) and Hymenomonas carterae (Chrysophyta). J. Exp. Mar. Biol. Ecol., 33: 153-168.

Braman, R.S. and Foreback, C.C. (1973) Methylated forms of arsenic in the environment. Science, 182: 1247-1249.

Budd, K. and Craig, S.R. (1981) Resistance to arsenate toxicity in the blue-green alga Synecho-coccus leopoliensis. Can. J. Bot., 59: 1518- 1521.

Burton, J.D. and Statham, P.J. (1982) Occurrence, distribution and chemical speciation of some minor dissolved constituents in ocean waters. Environ. Chem., Vol. 2. R. Soc. Chem., London, pp. 234-265.

Calvert, S.E. and Price, N.B. (1977) Shallow water, continental margin and lacustrine nodules: Distribution and geochemistry. In: Glasby, G.P. (Ed.) Marine Manganese Deposits. Elsevier, New York.

Cantoni, G.L. (1953) S-Adenosylmethionine. A new intermediate formed enzymatically from L-methionine and adenosinetriphosphate. J. Biol. Chem., 204: 403-416.

Carpenter, R., Peterson, M.L. and Jahnke, R.A. (1978) Sources, sinks and cycling of arsenic in the Puget Sound region. In: Wiley, M.L. (Ed.) Estuarine Interactions. Academic Press, New York. pp. 459-480.

Challenger, F. (1945) Biological methylation. Chem. Rev., 36: 315-361.

Challenger, F., Lisle, D.B. and Dransfield, P.B. (1954) Studies on biological methylation. Part XIV. The formation of trimethylarsine and dimethyl selenide in mould cultures from methyl sources containing 14C. J. Chem. Soc, pp. 1760-1771.

Chapman, A.C. (1926) On the presence of compounds of arsenic in marine crustaceans and shell fish, Analyst (London), 51: 548-563.

Christensen, E.R. and Zielski, P.A. (1980) Toxicity of arsenic and PCB to a green alga (Chlamydomonas). Bull. Environ, Contam. Toxicol, 24: 43-48.

98 Clement, W.H. and Faust, S.D. (1981) The release of arsenic from contaminated sediments and muds. J. Environ. Sci. Health, A16: 87-122.

Cockell, K.A. and Hilton, J.W. (1988) Preliminary investigations on the comparative chronic toxicity of four dietary arsenicals to juvenile rainbow trout {Salmo gairdneri R.). Aquat. Toxicol., 12: 73-82.

Cockell, K.A., Hilton, J.W. and Bettger, W.J. (1992) Hepatobiliary and hematological effects of dietary disodium arsenate heptahydrate in juvenile rainbow trout (Oncorhynchus mykiss). Comp. Biochem. Physiol., 103C(3): 453^158.

Conway, H.L. (1978) Sorption of arsenic and cadmium and their effects on growth, micronutrient utilization, and photosynthetic pigment composition of Asterionellaformosa, J. Fish. Res. Board Can., 35: 286- 294.

Cooney, R.V. (1981) The metabolism of arsenic by marine organisms. Ph D. Dissertation, Univ. of California, San Diego.

Cooney, R.V. and Benson, A.A. (1980) Arsenic metabolism in Homarus americanus. Chemo-sphere, 9: 335-341.

Cooney, R.V., Mumma, R.O. and Benson, A.A. (1978) Arsonium- phospholipid in algae. Proc. Natl. Acad. Sci. U.S.A., 75: 4262-4264.

Cotton, F.A. and Wilkenson, G. (1972) Inorganic Chemistry. Wiley & Sons, New York.

Cox, H.E. (1925) On certain new methods for the determination of small quantities of arsenic, and its occurrence in urine and in fish. Analyst (London), 50: 3-13.

Cullen, W.R. and Dodd, M. (1989) Arsenic speciation in clams of British Columbia. Appl. Organomet. Chem., 3: 79-88.

Cullen, W.R. and Reimer, KJ. (1989) Arsenic speciation in the environ­ ment. Chem, Rev., 89: 713-764.

Cullen, W.R., Dodd, M., Nwata, B.U., Reimer, D.A. and Reimer, K.J. (1989) Compounds of arsenic, antimony, and tin in mollusc shells. Appl. Organomet. Chem., 3: 351-353.

99 Cutter, G.A. (1992) Kinetic controls of metalloid speciation in seawater. Mar. Chem., 40: 65-80.

Day, J.W.,Jr., Hall, C.A.S., Kemp, W.M. and Yanez-Arancibia, A. (1989) Estuarine Ecology. Wiley, New York.

Dushenko, W.T., Bright, D.A. and Reimer, K.J. (1995) Arsenic bioaccumulation and toxicity in aquatic macrophytes exposed to gold­ mine effluent: relationships with environmental partitioning, metal uptake and nutrients. Aquat. Bot., 50: 141-158.

Ebdon, L., Walton, A.P., Millward, G.E. and Whitfield, M. (1987) Methylated arsenic species in estuarine porewaters. Appl. Organomet. Chem., 1: 427-433.

Edmonds, J.S. and Francesconi, K.A. (1981) Arseno-sugars from brown kelp {Ecklonia radiata) as intermediates in cycling of arsenic in a marine ecosystem. Nature , 289: 602-604.

Edmonds, J.S. and Francesconi, K.A. (1983) Arsenic-containing ribofuranosides: Isolation from brown kelp Ecklonia radiata and NMR spectra. J. Chem. Soc., Perkin Trans. 1. pp. 2375- 2382.

Edmonds, J.S. and Francesconi, K.A. (1987) Trimethylarsine oxide in estuary catfish (Cnidoglanis macrocephalus) and school whiting (Sillago bassensis) after oral administration of sodium arsenate; and as a natural component of estuary catfish. Sci. Total Environ., 64: 317-323.

Edmonds, J.S. and Francesconi, K.A. (1988) The origin of arsenobetaine in marine animals. Appl. Organomet. Chem., 2: 297-302.

Edmonds, J.S. and Francesconi, K.A. (1993) Arsenic in seafoods: Human health aspects and regulations. Mar. Pollut. Bull., 26: 665-674.

Edmonds, J.S., Francesconi, K.A., Cannon, J.R., Raston, C.L., Skelton, B.W. and White, A.H. (1977) Isolation, crystal structure and synthesis of arsenobetaine, the arsenical constituent of the western rock lobster Panulirus longipes cygnus George. Tetrahedron Lett., 18: 1543-1546.

100 Edmonds, J.S., Shibata, Y., Francesconi, K.A., Yoshinaga, J. and Morita, M. (1992) Arsenic lipids in the digestive gland of the western rock lobster Pcmulirus cygnus: An investigation be HPLCICP-MS. Sci. Total Environ., 122: 321-335.

Eisler, R. (1994) A review of arsenic hazards to plants and animals with emphasis on fishery and wildlife resources. In: Nriagu, J O. (Ed.) Arsenic in the Environment, Part II: Human Health and Ecosystem Effects. John Wiley & Sons, Inc., New York. pp. 185-259.

EPA (1985a) Guidelines for deriving numerical national water quality criteria for the protection of aquatic organisms and their uses. NTIS No. PB 85-227049. 98 pp.

EPA (1985b) Ambient water quality criteria for arsenic - 1984. U.S. Environ. Prot. Agency Rep., 440/ 5-84-033. 66 pp.

Ettajani, H., Amiard-Triquet, C., Jeantet, A.Y. and Amiard, J.C. (1996) Fate and effects of soluble or sediment-bound arsenic in oysters (Crassostrea gigas Thun.). Arch. Environ. Contain. Toxicol., 31: 38 —46.

Eurostat (1994) Europe's Environment. Statistical Office of the Ear. Communities, Luxembourg.

Ferguson, J.F. and Gavis, J. (1972) A review of the arsenic cycle in natural waters. Water Res., 6(11): 1259-1274.

Francesconi, K.A. and Edmonds, J.S. (1993) Arsenic in the sea. In: Ansell, A.D., Gibson, R.N. and Barnes, M. (Eds.) Oceanography and Marine Biology: An Annual Review. UCL Press, Vol. 31. pp. 111-151.

Francesconi, K.A. and Edmonds, J.S. (1994) Biotransformation of arsenic in the marine environment. In: Nriagu, J.O. (Ed.) Arsenic in the Environment. Part I: Cycling and Characterization. John Wiley & Sons, Inc., New York. pp. 221-261.

Francesconi, K.A., Edmonds, J.S. and Stick, R.V. (1989) Accumulation of arsenic in yellow mullet (Aldrichetta forsteri) following oral administration of organoarsenic compounds and arsenate. Sci Total Environ., 79: 59-67.

101 Francesconi, K.A., Stick, R.V. and Edmonds, J.S. (1990) Glycerylphos- phorylarsenocholine and phosphatidylarsenocholine in yelloweye mullet (Aldrichetta forsteri) following oral administration of arsenocholine. Experientia, 46: 464-466.

Francesconi, K.A., Edmonds, J.S. and Stick, R.V. (1992a) Arsenic compounds from the kidney of the giant clam Tridacna maxima : Isolation and identification of an arsenic-containing nucleoside. J. Chem. Soc., Perkin Trans, l.pp. 1349-1357.

Francesconi, K.A., Edmonds, J.S. and Stick, R.V. (1992b) Arsenocholine from anaerobic decomposition of a trimethylarsonioriboside. Appl. Organomet. Chem., 6: 247-249.

Friberg, L. (1988) The GESAMP evaluation of potentially harmful substances in fish and other seafood with special reference to carcinogenic substances. Aquat. Toxicol., 11: 379-393.

Froelich, P.N., Kaul, L.W., Byrd, J.T., Andreae, M.O. and Roe, K.K. (1985) Arsenic, barium, germanium, tin, dimethylsulfide and nutrient biogeochemistry in Charlotte Harbor, Florida, a phosphorus-enriched estuary. Estuarine, Coastal Shelf Sci., 20: 239-264.

GESAMP (1986) Review of potentially harmful substances: arsenic, mercury and selenium. Reports and Studies, GESAMP, No. 28.

Gibbs, P.E., Langston, W.J., Burt, G.R. and Pascoe, P.L. (1983) Tharyx marioni (polychaeta): A remarkable accumulator of arsenic. J. Mar. Biol. Assoc. U.K., 63: 313-325.

Gohda, S. (1975) Valence state of arsenic and antimony in sea water. Bull. Chem. Soc. Jpn., 48: 1213-1216.

Gupta, S.K. and Chen, K.Y. (1978) Arsenic removal by adsorption. J. Wat. Poll. Contr. Fed., 493-506.

Hanaoka, K., Matsumoto, T., Tagawa, S. and Raise, T. (1987) Microbial degradation of arsenobetaine, the major water soluble organoarsenic compound occurring in marine animals. Chemosphere, 16: 2545-2550.

Hanaoka, K., Tagawa, S. and Raise, T. (1992) The fate of organoarsenic compounds in marine ecosystems. Appl. Organomet. Chem., 6: 139-146.

102 Hofslagare, O., Sjoberg, S. and Samuelsson, G. (1994) The effect of arsenate and arsenite on photosynthesis in Scenedesmus obliquus. A potentiometric study in a closed C02-system. Chem. Spec. Bioavail., 6(4): 95-102.

Holm, T.R., Anderson, M.A., Iverson, D.G. and Stanforth, R.S. (1979) Heterogeneous interactions of arsenic in aquatic systems. ACS Symp. Ser., 93: 711-736.

Howard, A.G. and Comber, S.D.W. (1989) The discovery of hidden arsenic species in coastal waters. Appl. Organomet. Chem., 3: 509-514.

Howard, A.G., Arbab-Zavar, M.H. and Apte, S.C. (1982) Seasonal variability of biological arsenic methylation in the estuary of the river Beaulieu. Mar. Chem., 11: 493^198.

Howard, A.G., Arbab-Zavar, M.H. and Apte, S.C. (1984) The behaviour of dissolved arsenic in the estuary of the river Beaulieu. Estuarine, Coastal Shelf Sci., 19: 493-504.

Howard, A.G., Apte, S.C., Comber, S.D.W. and Morris, R.J. (1988) Biogeochemical control of the summer distribution and speciation of arsenic in the Tamar Estuary. Estuarine, Coastal Shelf Sci., 27: 427^-44.

Johnson, D.L. (1972) Bacterial reduction of arsenate in sea water. Nature (London), 240: 44-45.

Jonhson, D.L. and Burke, R.M. (1978) Biological mediation of chemical speciation. II. Arsenate reduction during marine phytoplankton blooms. Chemosphere, 7: 645-648.

Johnson, D.L. and Pilson, M.E.Q. (1972) Arsenate in the western North Atlantic and adjacent regions. J. Mar. Res., 30: 140-149.

Johnson, D.L. and Pilson, M.E.Q. (1975) The oxidation of arsenite in seawater. Environ. Lett., 8: 157-171.

Jones, A.J. (1922) The arsenic content of some of the marine algae. In: Hampshire, C.H. (Ed.) Year Book of Pharmacy, Transact. British Pharm. Conf., July 1922. Churchill, London, pp. 388-395.

103 Kennedy, V.S. (1976) Arsenic concentrations in some coexisting marine organisms from New-foundland and Labrador. J. Fish. Res. Board Can., 33: 1388-1393.

Klumpp, D.W. (1980) Characteristics of arsenic accumulation by seaweeds Fucus spiralis and Ascophyllum nodosum. Mar. Biol. (Berlin), 58: 257-264.

Klumpp, D.W. and Peterson, PJ. (1981) Chemical characteristics of arsenic in a marine food chain. Mar. Biol. (Berlin), 62: 297-305.

Kuroiwa, T., Ohki, A., Naka, K. and Maeda, S. (1994) Biomethylation and biotransformation of arsenic in a freshwater food chain: green alga (Chlorella vulgaris) —» shrimp (Neocaridina denticulata) —» killifish (Oryzias latipes). Appl. Organomet. Chem., 8: 325-333.

Kuroiwa, T., Yoshihiko, I., Ohki, A., Naka, K. and Maeda, S. (1995) Tolerance, bioaccumulation and biotransformation of arsenic in freshwater prawn ([Macrobrachium rosenbergii). Appl. Organomet. Chem., 9: 517-523.

Landstrom, O. and Wenner, C.G. (1965) Neutron-activation analysis of natural water applied to hydrology. AB Atomenergi, AE-204, Studsvik, Sweden.

Lemmo, N.V., Faust, S.D., Belton, T. and Tucker, R. (1983) Assessment of the chemical and biological significance of arsenical compounds in a heavily contaminated water-shed. Part I: The fate and speciation of arsenical compounds in aquatic environments. - A literature review. J. Environ. Sci. Health A, 18(3): 335-387.

Lindestrom, L. (1992) On water quality criteria for metals. Vattenfall, FUD Report U(M) 1992/25. (In Swedish).

Ljunggren, K., Sjostrand, B. Johnels, A.G., Olsson, M., Otterlind, G. and Westermark, T. (1971) Activation analysis of mercury and other environmental pollutants in water and aquatic ecosystems. Nuclear techniques in environmental pollution. Intemat. Atomic Energy Agency, Vienna, pp. 373^105.

Lunde, G. (1972) The analysis of arsenic in the lipid phase from marine and limnetic algae. Acta Chem. Scand., 26: 2642-2644.

104 Lunde, G. (1973) Separation and analysis of organic-bound and inorganic arsenic in marine organisms. J. Sci. Food Agric., 24: 1021— 1027.

Lunde, G. (1975) A comparison of organoarsenic compounds from different marine organisms. J. Sci. Food Agric., 26: 1257-1262.

Lunde, G. (1977) Occurrence and transformation of arsenic in the marine environment. Environ. Health Perspect., 19: 47-52.

Luten, J.B., Riekwel-Booy, G. and Rauchbaar, A. (1982) Occurrence of arsenic in plaice (Pleuronectes platessa), nature of organo-arsenic compound present and its excretion in man. Environ. Health Perspect., 45: 165-170.

Maeda, S. (1994) Biotransformation of arsenic in the freshwater environ­ ment. In: Nriagu, J O. (Ed.) Arsenic in the Environment. Part I: Cycling and Characterization. John Wiley & Sons, Inc., New York. pp. 155-187.

Maeda, S., Fujita, S., Ohki, A., Yoshifuko, I, Higashi, S. and Takeshita, T. (1988) Arsenic accumulation by arsenic-tolerant freshwater blue- green algae (Phormidium sp.). Appl. Organo-met. Chem., 2: 353-357. Maeda, S., Inoue, R., Kozono, T., Tokuda, T., Ohki, A. and Takeshita, T. (1990) Arsenic metabolism in a freshwater food chain. Chemosphere, 20: 101-108.

Maeda, S., Kusadome, K., Arima, H., Ohki, A and Naka, K. (1992a) Biomethylation of arsenic and its excretion by the alga Chlorella vulgaris. Appl. Organomet. Chem., 6: 407^113.

Maeda, S., Ohki, A., Kusadome, K., Kuroiwa, T., Yoshifuko, I. and Naka, K. (1992b) Bioaccumulation of arsenic and its fate in a freshwater food chain. Appl. Organomet. Chem., 6: 213- 219.

Maher, W.A. (1983) Inorganic arsenic in marine organisms. Mar. Pollut. Bull, 14: 308-310.

Marin, A.R, Masscheleyn, P.H. and Patrick, W.H, Jr. (1992) The influence of chemical form and concentration of arsenic on rice growth and tissue arsenic concentration. Plant, Soil, 139: 175-183.

Maugh, T.H. (II) (1979) It isn't easy being king. Science, 203: 637.

105 Meharg, A. A. and Macnair, M R. (1992) Suppression of the high affinity phosphate uptake system: A mechanism of arsenate tolerance in Holcus lanatus L. J. Exp. Bot., 43: 519-524.

Mok, W.M. and Wai, C.M. (1994) Mobilization of arsenic in con­ taminated river waters. In: Nriagu, J.O. (Ed.) Arsenic in the Environment. Part I: Cycling and Characterization. John Wiley & Sons, Inc., New York. pp. 99-117..

NAS (1977) Medical and biological effects of environmental pollutants: Arsenic. Nat. Acad. Sci., Washington D C.

Nichols, J.W., Wedemayer, G.A., Mayer, F.L., Dickhoff, W.W., Gregory, S.V., Yasutake, W.T. and Smith, S.D. (1984) Effects of freshwater exposure to arsenic trioxide on the parr-smolt transformation of coho salmon (Ortcorhynchus kisutch). Environ. Toxicol. Chem., 3: 143-149.

Nissen, P. and Benson, A.A. (1982) Arsenic metabolism in freshwater and terrestrial plants. Physiol. Plant., 54: 446-450.

Norin, H. and Vahter, M. (1984) Organic arsenic compounds in fish. Swedish Environmental Protection Agency, SNV PM 1892, Solna. (In Swedish).

Norin, H., Christakopoulos, A., Sandstrom, M. and Ryhage, R. (1985a) Mass fragmentographic estimation of trimethylarsine oxide in aquatic organisms. Chemosphere, 14: 313-323.

Norin, H., Vahter, M., Christakopoulos, A. and Sandstrom, M. (1985b) Concentration of inor-ganic and total arsenic in fish from industrially polluted water. Chemosphere, 14: 325-334.

Notini, M., Holmgren, K. and Rosemarin, A. (1987) Long-term fate and effects of low levels of arsenate on the Baltic Sea Fucus vesiculosus ecosystem. Proc. 21st Eur. Mar. Biol. Symp., Gdansk, Poland, 1986.

Notter, M. (Ed.) The metals and the environment. Swedish Environ­ mental Protection Agency, SEPA, Report No. 4135. (In Swedish).

106 OECD (1982) Eutrophication of Waters. Monitoring, Assessment and Control. Organization of Economic Cooperation and Development, OECD, Paris. 164 pp.

Oscarson, D.W., Huang, P.M. and Liaw, W.K. (1980) The oxidation of arsenite by aquatic sediments. J. Environ. Qual., 9(4): 700-703.

Phillips, D.J.H. (1990) Arsenic in aquatic organisms: A review emphasizing chemical speciation. Aquat. Toxicol., 16: 151-186.

Phillips, D.J.H. (1994) The chemical forms of arsenic in aquatic organisms and their interrelation-ships. In: Nriagu, J.O. (Ed.) Arsenic in the Environment. Part I: Cycling and Characterization. John Wiley & Sons, New York. pp. 263-287.

Phillips, D.J.H. and Depledge, M.H. (1985) Metabolic pathways involving arsenic in marine organisms: A unifying hypothesis. Mar. Environ. Res., 17: 1-12.

Pillay, K.K.S., Thomas, C.C. and Hyche, C.M. (1973) Neutron activation analysis of some of the biologically active trace elements in fish. In: Hemphill, D.D. (Ed.) Proc. 7th Ann. Conf. Trace Subst. Environ. Health, pp. 415—420.

Planas, D. and Healey, F.P. (1978) Effects of arsenate on growth and phosphorus metabolism of phytoplankton. J. Phycol., 14: 337-341.

PPO (1996) Environmental aspects of metals and metalloids in the Desaguadero hydrologic system. Proyecto Piloto Oruro. Ministerio de Desarrollo Sostenible y Medio Ambiente, Secretaria Nacional de Mineria, Swedish Geological AB. PPO 9607, May 20, 1996. 40 pp.

Reimer, K.J. and Thompson, J.A.J. (1988) Arsenic speciation in marine interstitial water. The occurrence of organoarsenicals. Biogeochemistry, 6:211-237.

Reuther, R. (1986) The occurrence and speciation of arsenic in the aquatic environment. A literature review. Swedish Environ. Research Group (MEG) Report, Fryksta, 20/11,1986. 48 pp.

107 Reuther, R. (1989) Geochemical mobility and availability of arsenic in two flowthrough freshwater model ecosystems. Swedish Environmental Research Group (MEG) Report, Fryksta, Oct. 10 1989. 59 pp.

Reuther, R. (1992) Arsenic introduced into a littoral freshwater model ecosystem. Sci. Total Environ., 115: 219-237.

Riedel, G.F. (1993) The annual cycle of arsenic in a temperate estuary. Estuaries, 16: 533-540.

Riedel, G.F., Sanders, J.G. and Osman, R.W. (1987) The effect of biological and physical disturbances on the transport of arsenic from contaminated estuarine sediments. Estuarine, Coastal Shelf Sci., 25: 693-706.

Riedel, G.F., Sanders, J.G. and Osman, R.W. (1989) The role of three species of benthic invertebrates in the transport of arsenic from contaminated estuarine sediments. J. Exp. Mar. Biol. Ecol., 134: 143— 155.

Rosen, G and Lithner, G. (1986) The response of phytoplankton to metal contamination in the Rdnnskar area. SEPA Laboratory for Monitoring of Products and Emissions. Preliminary Report 1986-05-22.

Rothstein, A. (1963) Interactions of arsenate with the phosphate­ transporting system of yeast. J. Gen. Physiol., 46: 1075-1085.

Sanders, J.G. (1979) Effects of arsenic speciation and phosphate concentration on arsenic inhibition of Skeletonema costatum (Bacillariophyceae). J. Phycol., 15: 424-428.

Sanders, J.G. (1980) Arsenic cycling in marine systems. Mar. Environ. Res., 3(4): 257-266.

Sanders, J.G. (1983) Role of marine phytoplankton in determining the chemical speciation and biogeochemical cycling of arsenic. Can. J. Fish. Aquat. Sci., 40: 192-196.

Sanders, J.G. (1985) Arsenic geochemistry in Chesapeake Bay: Dependence upon anthropogenic inputs and phytoplankton species composition. Mar. Chem., 17: 329-340.

108 Sanders, J.G. and Cibik, S.J. (1985) Adaptive behavior of euryhaline phytoplankton communities to arsenic stress. Mar. Ecol.: Prog. Ser., 22: 199-205.

Sanders, J.G. and Cibik, S.J. (1988) Response of Chesapeake Bay phytoplankton communities to low levels of toxic substances. Mar. Pollut. Bull., 19: 439-444.

Sanders, J.G. and Riedel, G.F. (1987) Control of trace element toxicity by phytoplankton. Recent Adv. Phytochem., 21: 131-149.

Sanders, J.G. and Riedel, G.F. (1993) Trace element transformation during the development of an estuarine algal bloom. Estuaries, 16: 521- 531.

Sanders, J.G. and Vermersch, P S. (1982) Response of marine phyto­ plankton to low levels of arsenate. J. Plankton Res., 4: 881-893.

Sanders, J.G. and Windom, H.L. (1980) The uptake and reduction of arsenic species by marine algae. Estuarine Coastal Mar. Sci., 10: 555— 567.

Sanders, J.G., Osman, R.W. and Brownlee, D C. (1988) Arsenic transport and impact in Chesapeake Bay food webs. U.S. EPA, Chesapeake Bay Liaison Office Publ., CBP/TRS 18/88.

Sanders, J.G., Osman, R.W. and Riedel, G.F. (1989) Pathways of arsenic uptake and incorporation in estuarine phytoplankton and filter-feeding invertebrates, Eurytemora affinis, Balanus improvisus and Crassostrea virginica. Mar. Biol. (Berlin), 103: 319-325.

Sanders, J.G., Riedel, G.F. And Ferrier, D P. (1991) Changes in community structure of Chesa- peake Bay phytoplankton when exposed to low levels of trace metals: Management implications. In: Mihursky, J.A. and Chaney, A. (Eds.) New Perspectives in the Chesapeake System: A Research and Management Partnership. Proc. Conf. Baltimore MD, 1990. Chesapeake Res. Consortium Publ., No. 137, pp. 451—460.

Sanders, J.G., Riedel, G.F. and Osman, R.W. (1994) Arsenic cycling and its impact in estuarine and coastal marine ecosystems. In: Nriagu, J.O. (Ed.) Arsenic in the Environment. Part I: Cycling and Characterization. John Wiley & Sons, Inc., New York. pp. 289-308.

109 Shibata, Y. and Morita, M. (1988) A novel, trimethylated arseno-sugar isolated from the brown alga Sargassum thunbergii. Agric. Biol. Chem., 52: 1087-1089.

Shibata, Y., Morita, M. and Edmonds, J.S. (1987) Purification and identification of arsenic-containing ribofuranosides from the edible brown seaweed, Laminaria japonica (MAKONBU). Agric. Biol. Chem., 51: 391-398.

Shinagawa, A., Shiomi, K., Yamanaka, H. and Kikuchi, T. (1983) Selective determination of inorganic arsenic (III), (V) and organic arsenic in marine organisms. Bull. Jpn. Soc. Sci. Fish., 49: 75-78.

Shiomi, K., Shinagawa, A., Igarashi, T., Hirota, K., Yamanaka, H. and Kikuchi, T. (1984) Contents and chemical forms of arsenic in shellfishes in connection with their feeding habits. Bull. Jpn. Soc. Sci. Fish., 50: 293-297.

Shiomi, K., Aoyama, M., Yamanaka, H. and Kikuchi, T. (1988) Chemical forms of arsenic in sponges, sea anemones and sea-hare. Comp. Biochem. Physiol., 90C: 361-365.

Thursby, G.B. and Steele, R.L. (1984) Toxicity of arsenite and arsenate to the marine macroalgae Champia parvula (Rhodophyta). Environ. Toxicol. Chem., 3: 391-397.

Traversy, W.J., Goulden, P.D., Sheikh, Y.M. and Leacock, J.R. (1975) Levels of arsenic and selenium in the Great Lakes Region. Environment Canada, Canada Centre for Inland Waters, Sci. Ser. 58. 18 pp.

Tremblay, G. and Gobeil, C. (1990) Dissolved arsenic in the St. Lawrence Estuary and the Saguenay Fjord, Canada. Mar. Pollut. Bull., 21:465-469.

Walterson, E. and Landner, L. (1996) The impact of Ronnskarsverken on the external environment - an evaluation of field investigations. Swedish Environmental Research Group (MFG), Report to Boliden Mineral AB, Stockholm, April, 1996. (In Swedish).

Waslenchuk, D.G. (1978) The budget and geochemistry of arsenic in a continental shelf environment. Mar. Chem., 7: 39-52.

110 Weis, J.S. and Weis, P. (1992) Construction materials in estuaries: reduction in the epibiotic community on chromated copper arsenate (CCA) treated wood. Mar. Ecol. Prog. Ser., 83: 45-53.

Weis, J.S. and Weis, P. (1994) Effects of contaminants from chromated copper arsenate-treated lumber on benthos. Arch. Environ. Contam. Toxicol., 26: 103-109.

Weis, J.S. and Weis, P. (1996a) The effects of using wood treated with chromated copper arsenate in shallow-water environments: A review. Estuaries, 19(2A): 306-310.

Weis, J.S. and Weis, P. (1996b) Reduction in toxicity of chromated copper arsenate (CCA)-treated wood as assessed by community study. Mar. Environ. Res., 41(1): 15-25.

Weis, P., Weis, J.S. and Coohill, L.M. (1991) Toxicity to estuarine organisms of leachates from chromated copper arsenate-treated wood. Arch. Environ. Contam. Toxicol., 20: 118-124.

Whyte, J.N.C. and Englar, J.R. (1983) Analysis of inorganic and organic- bound arsenic in marine brown algae. Bot. Mar., 26: 159-164.

Woolson, E.A. (Ed.) (1975) Arsenical Pesticides. ACS Symp. Ser. 7. Am. Chem. Soc., Washington D.C.

Woolson, E.A. (1977) Fate of arsenicals in different environmental substrates. Environ. Health Perspect., 19: 73-81.

Wangberg, S.-A. (1995) Effects of arsenate and copper on the algal communities in polluted lakes in the northern parts of Sweden assayed by PICT (Pollution-Induced Community Tolerance). Hydrobiologia, 306: 109-124.

Wangberg, S.-A. and Blanck, H. (1990) Arsenate sensitivity in marine periphyton communities established under various nutrient regimes. J. Exp. Mar. Biol. Ecol., 139: 119-134.

Wangberg, S.-A., Heyman, U. and Blanck, H. (1991) Long-term and short-term arsenate toxicity to freshwater phytoplankton and perriphyton in limnocorrals. Can. J. Fish. Aquat. Sci., 48: 173-182.

Ill KEMI REPORT SERIES

5/88 Flotation chemicals from ore 1987 dressing plants (in Swedish) 1/87 Supervision project - part I 6/88 Environmental effects of orga- Monitoring of chemical product notin in antifouling paints information from 86 companies (English summary) (in Swedish) 7/88 1-4 -dichlorobenzene 2/87 Formaldehyde - a hazard anal­ (in Swedish) ysis (English summary) 8/88 The use of QSAR for chemicals 3/87 Classification and labelling of screening (in English) preparations containing carcinogenic substances 9/88 Initial assessment of the envir­ ReportfromaNordic working group onmental hazard of chemical (in Swedish) substances An evaluationof the "ESTHER 4/87 Analysis of tris (1,3 -d ichloro- manual" (English summary) propyQphosphate in products and human blood 10/88 Effect catalysis (in Swedish) A concept for assessment of the environmental risk of chemicals 5/87 A strategy for ranking chemi­ (in Swedish) cals for biological testing (in English) 11/88 Introduction into genetic toxi­ cology 6/87 Applied toxicology A background document for Some basic views (in Swedish) assessment of the genotoxicity of chemicals (in Swedish) 7/87 Environmental effects of ma­ rine antifouling paints Survey by the laboratory section 1989 for aquatic toxicology within the 1/89 Trimellitic anhydride (TMA) National Environmental Protec­ A hazard analysis tion Board (in Swedish) (in English)

1988 2/89 Benzene and total hydrocar­ bon exposure during petrol 1/88 Methyl-tert-butylether filling of private automobiles A bibliography (in Swedish) (in Swedish) 2/88 Biological control of pest and pi- 3/89 The algal microtest battery seases on agriculture and A manual for routine tests of horticultural crops in Sweden growth inhibition (in English) (English summary) 4/89 Systems for testing and haz­ 3/88 Formaldehyde emission from ard evaluation of chemicals furniture in the aquatic environment A critical review (in Swedish) A manual for an initial assess ment - ESTHER (in English) 4/88 Fluxing agents (in Swedish) 5/89 Methods to quantify the toxi­ 6/90 Cadmium city and hazard of insecti­ An Analysis of Swedish cides to honeybees Regulatory Experience (Apis mellifera L.) (in Egnlish) (English summary) 7/90 Detergents and cleaners for 6/89 Risk assessment of air pollut­ domestic use (in Swedish) ants The Swedish Toxicological 8/90 Approval of pesticides Council (in Swedish) Assessments, restrictions and information (in Swedish) 7/89 Carcinogenic substances Review of the substances on 9/90 The use of chemicals in the Keml list (in Swedish) Sweden Flows, functions (applications) 8/89 Organotin in the Swedish and effects (in Swedish) aquatic environment (in English) 10/90 Environmental risk reduction A Government Commission 9/89 Comparison of different Report models for environmental (in Swedish) hazard classification of Appendix chemicals Report on risk reduction of Chemicals 10/89 Environmental hazardous (in Swedish) substances List of examples and scientific 11/90 Non-genotoxic carcinogens documentation Report from a seminar. The Swe­ dish Toxicological Council 11/89 Evaluations of carcinogenic (in Swedish) substances I 12/90 Environmentally suited 1990 degreasing of motor vehicles (in Swedish) 1/90 Car cleaning products A pilot study (in Swedish) 13/90 Pattern of lead emissions in Sweden 1880 -1980 2/90 Effects on reproduction of sty­ (in English) rene, toluene and xylene Nordic Council of Ministers 1991 (in English) 1/91 Risk reduction of chemicals 3/90 Effects on reproduction of tri- A Government Commission and tetrachloroethylene Report Nordic Council of Ministers (in English) (in English) 2/91 Overview of allergenic sub ­ 4/90 Use of mortality and morbi ­ stances included in chemical dity registers for detecting products and goods chemical health hazards (in Swedish) A Seminar Report (in Swedish) 3/91 Hazard evaluations of aller­ 5/90 Characterization of industrial genic properties of 20 sub ­ oils (in Swedish) stances (English summary) 4/91 Supervision project on curing 3/92 Car care products- a super­ resins (epoxy, isocyanates vision project and acrylates in products) Promotion of safer products (English summary) (English summary)

5/91 Brominated flame retardants 4/92 Principles for identifying (in Swedish) unacceptable pesticides (in English) 6/91 Environmental hazards of mi­ crobiocides in cooling water 5/92 Carcinogenic substances II (English summary) (in Swedish)

7/91 Data for assessment of health 6/92 The health risks of gasoline and environmental hazards of chemical substances (in Swedish) 7/92 Effects on reproduction of chloroform and 1,2-dibromo- 8/91 Flow analysis of metals 3-chloropropane (in Swedish) Nordic Council of Ministers (in English) 991 The burden of proof intoxicology (in English) 8/92 Lubricants (English summary) 10/91 Effects on reproduction on di- 9/92 Cleaning stations against chloromethane, n-hexane and 1,1,1-trichloroethane growth on pleasure boats (in English) (in Swedish) 11/91 Plasticizers 1993 A survey of plasticizers in Sweden (In Swedish) 1/93 A pilot procedure to select candidate substances for 12/91 The substitution principle under section 5 of the act on general restrictions on use (In English) chemical products (in Swedish) 2/93 Antifouling products Pleasure boats, commercial 13/91 Limit values (in Swedish) vessels, nets, fish cages and other underwater equipment 14/91 Cleaning products and skin (in English) effects (in Swedish) 3/93 Sensitizing substances 15/91 Flame retardants (in English) (in English) 4/93 The occurrence and use of chemicals in Society 1992 (English summary) 1/92 HF-90, substitution of hydrofluoric acid, - a super 5/93 Mutagenic substances vision project (English (in Swedish) summary) 6/93 Boat care products- 2/92 Printing inks a supervision project Composition, occurrance (English summary) and development (English summary) 7/93 Nickel allergy 9/94 Risk assessment of poly- Seminar Report, the Swedish brominated diphenyl ethers Toxicological Council (in English) (English summary) 10/94 Chemical substances lists 8/93 Databases containing A Guide to the lists used in ecotoxicological facts the Swedish Sunset project (in Swedish) Supplement to report 13/94 (in English) 9/93 Gallium, germanium and indium - a toxicological 11/94 Mono and di-substituted survey organotins used as plastic (English summary) additives Vol. 1 Environmental hazard assess­ 1994 ment Vol. 2 Health hazard identification (in English) 1/94 Photochemicals - are they hazardous? 12/94 Phtalatic acid esters used as A supervision project plastic additives (English summary) Vol. 1 Ecotoxicological risk assess­ ment 2/94 Mapping of the chemical and Vol. 2 Comparisions of toxicological toxicological knowledge of effects producers and importers of (in English) chemcial products (English summary) 13/94 Selecting multiproblem chemicals for risk reduction 3/94 Some uses of lead and their A presentation of the Swedish possible substitutes Sunset Project (in English) (in English)

4/94 Supervision of suppliers of 14/94 Seminar on international chemicals aspects on risk assessment (In Swedish) Seminar report the Swedish Toxicological 5/94 Detergent and cleaning Council products (in English) A report of the work of a Governmental Commission 15/94 Chlorine and chlorinated (English summary) compounds A report of the work of a 6/94 New ruts governmental commission - A product study of tyres (English summary) (English summary) 16/94 Chlorine and chlorinated 7/94 Use reduction of compounds in Sweden trichloroethylene for (in Swedish) industrial degreasing (In Swedish) 17/94 Chlorine and chlorinated Mona Olsson Oberg substances in Sweden (English summary) 8/94 Phasing out lead and mercury (in English) 1995 11/95 Risk assessment and risk management in chemicals control 1/95 Chlorine and chlorine compounds 12/95 Hazard assessments Report on a Governmental -Chemical substances assignment selected in the Swedish (in English), Swedish version Sunset project - 15/94 Supplement to Keml report 13/94

2/95 A priority setting scheme 13/95 Natural environment and for scoring hazardous health properties -how to improve exposure Supplement to report 13/94 analysis? (in English) Toxicological Council (in Swedish) 3/95 Chlorine compounds in chemical products 14/95 Chromium and nickel in -description and selection for Sweden further investigation Viveka Palm, Bo Bergback (in English) and Per Ostlund (in English

4/95 Tumour promotion 15/95 Plastics additives - studies with pesticides -final report from the Plastic (in English) additives project (summary in English) 5/95 Chlorine and chlorinated compounds in Sweden 16/95 The flame retardants project (in English), Swedish version Final report 16/94 (in Swedish), English version 5/96 6/95 An Introduction to Health Risk Assessment of 1996 Chemicals Marie Haag Groniund 1/96 Endocrine effects of (in English) chemicals Toxicological Council 7/95 Car wash products (in Swedish) - a supervesion project. Promotion of environmental 2/96 One shade greener information (in Swedish) (summary in English) 3/96 Overview of the chemical 8/95 Car care products and toxicological knowledge - development and incidence and competence of (summary in English) manufacturers and import ers of chemical 9/95 Risk assessment of products in 1995 slimicides (summary in English) Ulf Eriksson, Anders Johnson, Monica Tornlund

10/95 Cadmium and its health risks Toxicological Council (in Swedish) 4/96 Alternatives to persistent 5/97 Chemicals in textiles organic pollutants Report of a Government The Swedish input to the IFCS Commission expert meeting on persistent (in English, Swedish version organic pollutants in Manila, 2/97) the Philippines, 17-19June 1996 6/97 The Phase-Out Project (in English) Report of a government commission 5/96 The flame retardants project (in Swedish) Final report (in English), Swedish 16/95 7/97 Nickel in hand-tools Study to assess the risk of 6/96 Additives in PVC allergy Labelling of PVC (in Swedish) A report of the work of a Governmental Commission 8/97 Hollow Product Information (English summary) on Tooth Filling - an inspection project 7/96 Fargleverantdrer bekanner farg (in Swedish) Tillsynsprojekt farger Johan Hakans 9/97 POPs Karin Rumar - persistent organic pollutants in the environment, 8/96 Behavioural toxicology - to measure effects of Toxicological Council environmental toxins (in Swedish) Toxicological Council (in Swedish) 10/97 Mercury in products - a source of transboundary 1997 pollutant transport 1/97 Hormonal effect from (in English) chemicals - summing upp the state-of- 1998 the-art (in Swedish) 1/98 Cadmium Cadmium Exemption 2/97 Chemicals in textiles Report of a Government Substances Project Commission 1) Cadmium in goods - (in Swedish, English version contribution to environmental 5/97) exposure 2) Cadmium in Sweden - 3/97 Acute poisonings with environmental risks pesticides 3) Cadmium in fertilizers, soil - a comparative study at the crops and foods - the Swedish Swedish Poisons Information situation (in Swedish)

4/97 Additives in PVC Marking of PVC Report of a Government Commission (in English, Swedish version 6/96)