ANTTI VENHO IMMOBILIZATION OF ARSENIC THROUGH pH ...

ANTTI VENHO 

IMMOBILIZATION OF ARSENIC THROUGH pH CONTROL 

DURING BIOLEACHING OF NICKEL FLOTATION 

CONCENTRATE 

Master of Science Thesis 

Examiners: Professor Jaakko Puhakka 

and Dr. Minna Peltola 

Examiners and topic approved in the 

Faculty of Science and Environmental 

Engineering on Dec 5 th 2012

ABSTRACT 

TAMPERE UNIVERSITY OF TECHNOLOGY 

Faculty of Science and Environmental Engineering 

Department of Chemistry and Bioengineering 

VENHO, ANTTI: Immobilization of Arsenic through pH control during Bioleaching 

of Nickel Flotation Concentrate 

Master of Science Thesis, 92 pages 

October 2012 

Major subject: Environmental Biotechnology 

Examiners: Prof. Jaakko Puhakka and Dr. Minna Peltola 

Keywords: Arsenic immobilization, pH control, iron precipitation, bioleaching, nickel 

concentrate, pulp density 

Bioleaching is a biohydrometallurgical method that utilizes the ability of acidophilic 

microorganisms such as Acidithiobacillus ferrooxidans, At. caldus, or Leptospirillum 

ferrooxidans to oxidize mineral sulfides, iron and sulfur and thereby solubilize valuable 

metals from low-grade or refractory ores and concentrates at very low pH (< 3). Arsenic 

is a toxic element found everywhere in the Earth’s crust and a harmful side-product of 

hydrometallurgical processes including bioleaching. Arsenic is currently removed from 

wastewater streams of the mineral processing industry by co-precipitation and sorption 

with iron oxides by using lime neutralization, i.e. raising solution pH. The aim of this 

thesis work was to assess the simultaneous immobilization of arsenic and selective 

bioleaching of nickel and cobalt from nickel flotation concentrate (NFC; Mondo 

Minerals LLC, Sotkamo, Finland) by maintaining a constant elevated process pH of 3.0. 

Bioleaching of NFC was performed in aerated continuously-stirred tank reactors 

(CSTRs) as 30 L batch mode and 20 L semi-continuous mode (10 d retention time) 

processes at 21 °C. The aim of the batch mode experiment was to assess the effect of 

scale-up from previous shake-flask experiments and efficiency of bioleaching and 

arsenic immobilization with different pulp densities (5, 10, 15 % p.d.). The purpose of 

the semi-continuous mode experiment (5 % p.d.) was to assess parameters required for 

developing a continuous mode process. Finally, a shake-flask experiment (100 mL, 150 

rpm, 27 °C) was conducted without NFC with the aim to assess the role of bioleaching 

microorganisms in arsenic speciation and immobilization. 

In the 30 L batch mode experiment, bioleaching was conducted for 94, 58, and 79 

days with 15, 10, and 5 % p.d., respectively. Final nickel yields were 73, 66, and 76 %, 

cobalt yields were 78, 71, and 100 % with 15, 10, and 5 % p.d., respectively. Maximum 

leaching rates (250 mg-Ni/L/d and 18 mg-Co/L/d) were achieved with 15 % p.d. 

Soluble arsenic level remained low (< 1 mg/L, < 0.4 % yield) with all studied pulp 

densities after 21 days of operation. Leach residue analysis by X-ray diffraction (XRD) 

revealed that iron sulfates were the predominant compounds, ranging from 55 to 70 % 

(w/w) with all pulp densities studied. Formation of goethite (α-FeOOH, an effective 

adsorbent for arsenic) increased with increasing pulp density. Bacterial community 

I

analysis from samples collected at the end of the experiment showed that At. 

ferrooxidans dominated the population with all studied pulp densities. Additionally, 

Alicyclobacillus sp. was present in the 5 % p.d. reactor. 

Bioleaching in the 20 L semi-continuous mode experiment was conducted for 38 

days. Dissolved arsenic remained below 0.4 % throughout the experiment. After 14 

days of operation and after reaching the maximum yields of 12 and 22 % for nickel and 

cobalt, respectively, the metal yields started to decline until the end of the experiment. 

The 10 d retention time was possibly too short for the microbial community to establish 

a high bioleaching rate. Further, mass exchanges for the reactor were performed daily as 

a single 2 L exchange. Such abrupt changes in the reactor environment likely affected 

the microbial community development. A more continuous material flow and use of 

multiple reactors in a cascade are suggested for further studies. Leach residue analysis 

showed that the goethite content was approximately 5 times higher than in the 30 L 

batch mode experiment [~ 25 vs. 5 % (w/w), respectively]. More information is needed 

on the conditions that facilitate goethite formation during bioleaching of NFC. 

In the shake-flask experiment, solutions with 25.1 g/L ferrous sulfate, 20 mg/L 

arsenite [As(III)] and 10 % (v/v) inoculant from the 30 L batch experiment as well as 

abiotic controls were incubated for 28 days. Two starting pH values were used, i.e., pH 

1.7 and 2.5. Solution pH in the biotic flasks became and remained similar after 6 days. 

In all flasks, arsenite was completely oxidized to arsenate [As(V)] in one day. After 6 

days of incubation, the amount of soluble arsenic in the biotic solution with starting pH 

2.5 decreased to only 4 mg/L, majority of which (75 %) was arsenite. The 

concentrations of soluble arsenite and arsenate remained low until the end of the 

experiment, finishing at 4 mg/L As(III) and 1 mg/L As(V). Arsenite was also the 

dominant (~81 %) oxidation state in the biotic solution with starting pH 1.7 after 6 days, 

although all of the arsenic was in solution. The concentrations of arsenite and arsenate 

both decreased slowly towards the end of the experiment, finishing at 12 mg/L As(III) 

and 3 mg/L As(V). Thus, the starting pH had a significant effect on arsenic 

immobilization. The role of microorganisms was crucial in arsenic immobilization. 

However the main mechanism of immobilization remains to be determined. Possible 

mechanisms include the following: sorption on biomass or coprecipitation and sorption 

on ferric oxides generated by microbial oxidation. 

In conclusion, this thesis work showed selective bioleaching of NFC for nickel and 

cobalt and efficient immobilization of arsenic in batch and semi-continuous mode by 

maintaining a constant pH of 3.0. The highest leaching rates of nickel and cobalt were 

achieved with the highest pulp density studied, which encourages to further increase the 

pulp density in future studies. This work also demonstrated the importance of sufficient 

mixing: Settled and partially unleached material was present on the bottom of the 15 

and 10 % p.d. reactors at the end of the 30 L batch mode experiment. Similarly, 

unleached material was present on the bottom of the 20 L, 5 % p.d. reactor. Finally, the 

major role of microorganisms in arsenic immobilization in a biomining environment 

was confirmed. The mechanisms of arsenic immobilization and the role of 

microorganisms in arsenic immobilization should be looked into in future studies. 

II

TIIVISTELMÄ 

TAMPEREEN TEKNILLINEN YLIOPISTO 

Luonnontieteiden ja ympäristötekniikan tiedekunta 

Kemian ja biotekniikan laitos 

VENHO, ANTTI: Arseenin hallinta nikkelirikasteen bioliuotuksessa pH:n säädön 

avulla 

Diplomityö, 92 sivua 

Lokakuu 2012 

Pääaine: Ympäristöbiotekniikka 

Tarkastajat: professori Jaakko Puhakka ja maatalous- ja metsätieteiden tohtori Minna 

Peltola 

Avainsanat: Arseenin sitoutuminen, pH:n säätö, raudan saostuminen, bioliuotus, 

nikkelirikaste, lietetiheys 

Tunnetut maapallon metalliesiintymät ovat ehtymässä, eikä uusia, helposti 

hyödynnettävissä tai saatavissa olevia esiintymiä löydy riittävästi kattamaan alati 

kasvavaa maailmanlaajuista kysyntää. Biohydrometallurgisilla menetelmillä, so. 

bioliuotuksella ja -hapetuksella, arvometallipitoisuuksiltaan köyhiä tai vaikeasti 

hyödynnettäviä mineraaleja tai rikasteita voidaan prosessoida taloudellisesti 

kannattavalla tavalla. Biohydrometallurgisten menetelmien etuja perinteisiin 

pyrometallurgisiin menetelmiin, esimerkiksi pasutukseen tai sulatukseen, nähden ovat 

ympäristöystävällisyys sekä pienemmät kustannukset. Bioliuotukseksi kutsutaan 

prosessia, jossa usein sulfidimuotoinen liukenematon mineraali (esim. CuS2) muuttuu 

liukoiseen muotoon, usein sulfaatiksi (esim. CuSO4) mikro-organismien katalysoimana. 

Biohapetuksessa biokatalyyttinen liuotus ei kohdistu suoraan arvometallia sisältävään 

mineraaliin, vaan sitä ympäröivään mineraalimatriisiin. Esimerkkinä tästä on 

kultapitoisten malmien biohapetus, jonka seurauksena kultaa sisältävät partikkelit 

paljastuvat myöhempää syanidikäsittelyä varten, jossa kulta saatetaan liukoiseen 

muotoon. 

Sekä bioliuotuksessa että –hapetuksessa mikro-organismit tuottavat ferrirautaioneja 

[Fe(III)] hapettamalla ferrorautayhdisteitä. Ferrorautaa [Fe(II)] esiintyy runsaasti 

kaikkialla maapallolla esimerkiksi rikkikiisuna (FeS2). Mikro-organismien tuottamat 

ferrirautaionit puolestaan hapettavat kohdemineraalin ja bioliuotuksen ollessa kyseessä 

muuttavat arvometallin liukoiseen muotoon. Ferromuotoon pelkistyneet rautaionit ovat 

jälleen mikro-organismien hapetettavissa. Lisäksi sulfidimineraalien hapetuksessa 

vapautuu rikkiä, jonka rikinhapettajat voivat saattaa sulfaatiksi. Samalla vapautuu myös 

protoneja, mikä laskee ympäristön pH:ta. Bioliuotus- ja –hapetusolosuhteet ovat erittäin 

happamia (pH < 3). Bioliuotukseen ja -hapetukseen kykenevät mikro-organismit ovat 

ominaisuuksiltaan pääasiallisesti rautaa, rikkiä tai molempia hapettavia asidofiilisia 

kemolitotrofeja. Lisäksi useimmat lajit ovat aerobisia autotrofeja. Tutkituimpia 

bioliuotusbakteereja ovat rautaa ja rikkiä hapettava Acidithiobacillus ferrooxidans, 

rikkiä hapettava At. thiooxidans sekä rautaa hapettava Leptospirillum ferrooxidans. 

Suurin osa teollisista biohydrometallurgisista prosesseista toimii alle 40 °C 

lämpötilassa. Tästä johtuen suurin osa tutkituista bioliuottavista mikro-organismeista on 

bakteereja. Tulevaisuudessa prosessilämpötilat tulevat prosessien optimoinnin myötä 

III

todennäköisesti nousemaan jopa yli 60 °C:n, jolloin termofiiliset arkkieliöt kuten 

Sulfolobus metallicus tai Metallosphaera sedula ovat todennäköisiä bioliuottajia. 

Teollisen mittakaavan biohydrometallurgiset prosessit voidaan jakaa kahteen osaan: 

läjä- tai kasaprosesseihin sekä sekoitustankkiprosesseihin. Kasareaktori voi 

yksinkertaisimmillaan olla kasa käsittelemätöntä louhittua malmia, jota kastellaan päältä 

ja jonka alta kerätään arvometallipitoista valumavettä. Tehokkaimmilleen optimoitu 

kasareaktori on n. 6 ... 10 m korkea, tarkasti koottu kasa murskattua ja agglomeroitua 

malmia, jota ilmastetaan kasan alla olevan putkiston kautta ja jota kastellaan päältä 

happamoidulla ravinneliuoksella ja kierrätetyllä valumavedellä. Kasan pituus voi olla 

useita kilometrejä. Olosuhteet kasan sisällä eivät ole koskaan homogeeniset. 

Prosessiparametrit kuten pH, ilman ja veden saatavuus, lämpötila sekä näiden määräämä 

mikrobipopulaation koostumus voivat vaihdella paljonkin kasan sisällä. Sotkamossa 

sijaitsevan Talvivaaran kaivoksen liuotuskasat ovat esimerkki kasabioliuotuksesta. 

Täyssekoitteisen tankkireaktorin käyttö mahdollistaa olosuhteiden erinomaisen 

hallittavuuden ja reaktioiden huomattavan nopeuttamisen, mutta on huomattavasti 

kalliimpi menetelmä kasaliuotukseen verrattuna muun muassa pienen 

malminkäsittelykapasiteetin sekä happamien ja hapettavien olosuhteiden aiheuttamien 

suurten materiaalikustannusten takia. Tämän vuoksi sekoitustankkeja käytetään usein 

vain malmirikasteiden käsittelyyn. Tyypillinen esimerkki biohydrometallurgisesta 

tankkiliuotusprosessista on kultamalmin biohapetus ennen syanidikäsittelyä. 

Arseeni (As, atomiluku = 33) on useimmille eliöille toksinen, ihmiselle 

karsinogeeninen puolimetalli, jota esiintyy runsaasti maapallon ylemmässä kuoressa. 

Arseenilla on neljä hapetusastetta: +V (arsenaatti), +III (arseniitti), 0 (arseeni) sekä –III 

(arsiini). Näistä arsenaatti ja arseniitti ovat ainoat tavanomaisesti luonnossa esiintyvät 

hapetusmuodot. Kiinteässä muodossa arseeni esiintyy usein rikkiyhdisteinä, esimerkiksi 

auripigmenttinä (As2S3), realgaarina (As2S2 tai AsS) tai arsenopyriittinä (FeAsS). 

Arseenipitoisuus maaperässä on n. 0.1 ... 1000 ppm, josta suurin osa on arsenopyriittiä. 

Luonnonvesissä arseeni esiintyy arsenaatin ja arseniitin oksyanioneina (AsO3 3- , AsO4 3- ). 

Liuoksen redox-potentiaali ja pH määräävät pitkälti oksyanionien protonaatioasteen 

sekä arseenin hapetusluvun. Protonaatioaste pienenee ja hapetusluku kasvaa redoxpotentiaalin 

ja pH:n kasvaessa. Jotkin mikro-organismit sietävät arseenia, ja jotkin 

kykenevät jopa käyttämään arseenia energianlähteenä hapettamalla sitä. 

Kaivostoiminta on yksi suurimmista ihmisen toiminnasta aiheutuvien 

arseenipäästöjen lähteistä. Malmien sisältämä arseeni joutuu malmien käsittelyn 

seurauksena kaivosvesiin tai ilmaan, joiden mukana se kulkeutuu ympäristöön. 

Kaivosten lähivesistöissä arseenipitoisuus voi nousta useisiin milligrammoihin litraa 

kohti, kun maailman terveysjärjestö WHO:n juomavedelle asettama raja-arvo on 10 

µg/L. Arseenin hallinnalle on kaivosteollisuudessa suuri tarve, jotta toimintaa voidaan 

jatkaa vastuullisesti ympäristöä pilaamatta. Nykyään monet kaivokset poistavat arseenia 

vesistään lipeäkäsittelyllä, jonka seurauksena liuoksen pH nousee, ja liuoksessa oleva 

rauta saostuu erilaisina oksideina tai hydroksideina, esimerkiksi götiittinä (α-FeOOH), 

ferrihydriittinä (Fe2O3 · 0.5 H2O), hematiittina (α-Fe2O3) tai jarosiittina ([K, Na, 

NH4]Fe3(SO4)2(OH)6). Arseeni poistuu näiden rautasaostumien kanssa kerasaostumalla 

tai sorptiolla. 

Tässä työssä tutkittiin Mondo Minerals Oy:n Sotkamon kaivoksesta peräisin olevan 

nikkeli- ja kobolttipitoisen flotaatiorikasteen selektiivistä bioliuotusta, jossa arvokkaat 

nikkeli ja koboltti saadaan talteen samanaikaisesti minimoiden arseenin liukeneminen 

prosessissa. Työtä edeltävät panoskokeet olivat osoittaneet, että nikkeli ja koboltti 

saadaan bioliuotettua ja arseeni pidettyä tehokkaasti pois liuoksesta, kun pH pidetään 

noin arvossa 3,0. Tässä työssä koejärjestelyjen mittakaavaa kasvatettiin: 

IV

nikkelirikastetta bioliuotettiin ilmastetuissa täyssekoitusreaktoreissa 30 L 

panosprosessina ja 20 L puolijatkuvasyötteisenä prosessina noin 21 °C lämpötilassa 

pH:n ollessa 3,0. Panoskokeen tarkoituksena oli tutkia mittakaavan kasvattamisen ja 

nikkelirikasteen lietetiheyden (5, 10 ja 15 %) vaikutusta bioliuotustehokkuuteen. 

Panoskokeen tulosten perusteella tehtiin 20 L puolijatkuvasyötteinen bioliuotuskoe 5 % 

lietetiheydellä ja 10 päivän viipymäajalla. Kokeen tarkoituksena oli tunnistaa 

puolijatkuvasyötteisyyden vaikutuksia arvometallien liuotukseen ja arseenin 

pidättymiseen sekä tunnistaa tekijöitä ja parametrejä, jotka on otettava huomioon 

prosessin muuttamisessa jatkuvasyötteiseksi. Lisäksi tässä työssä tutkittiin pullokokeilla 

(100 mL) biohapettavien mikro-organismien läsnäolon ja liuoksen alku-pH:n roolia 

liukoisen arseenin pidättymisessä ja hapetusasteen määräytymisessä. 

30 L panoskokeessa nikkelirikastetta bioliuotettiin 94, 58 ja 79 päivää vastaavassa 

järjestyksessä lietetiheydellä 15, 10 ja 5 %. Nikkelisaannot olivat edellä mainitussa 

lietetiheysjärjestyksessä 73, 66 ja 76 %. Koboltin saannot olivat 78, 71 ja 100 %. 

Suurimmat hetkelliset liuotusnopeudet (250 mg-Ni/L/d and 18 mg-Co/L/d) saavutettiin 

15 % lietetiheydellä. Liukoisen arseenin määrä pysyi alhaisena (< 1 mg/L, < 0,4 % 

saanto) kaikilla tutkituilla lietetiheyksillä 21 liuotuspäivän jälkeen. Reaktorien pohjalta 

kerättyjen liuotusjäämänäytteiden röntgendiffraktioanalyysin perusteella suurin osa 

jäämistä oli rautasulfaattia sen massaosuuden ollessa 55 ja 70 % välillä kaikilla 

tutkituilla lietetiheyksillä. Arseenin adsorptio-ominaisuuksiltaan hyväksi todetun 

götiitin muodostuminen kasvoi lietetiheyden kasvaessa. Bakteeriyhteisöt panoskokeen 

lopussa koostuivat lähes yksinomaan yhdestä At. ferrooxidans-lajista. Poikkeuksena oli 

5 % lietetiheydellä toimineessa reaktorissa Alicyclobacillus sp. esiintyminen. 

Nikkelirikastetta bioliuotettiin 20 L puolijatkuvasyötteisessä reaktorissa 38 päivää. 

Liukoinen arseeni pysyi alle 0,4 %:n koko kokeen ajan. Nikkeli- ja kobolttisaannot 

saavuttivat huippunsa (12 ja 22 % vastaavassa järjestyksessä) kun liuotusta oli kestänyt 

14 päivää. Tämän jälkeen saannot laskivat aina kokeen loppuun asti. On todennäköistä, 

että käytetty 10 päivän viipymäaika oli liian lyhyt mikrobipopulaation mukautumiselle. 

Reaktorin puolijatkuvasyötteisyys toteutettiin päivittäisinä nopeina 2 L 

massanvaihdoilla. Tästä johtuen mikrobiyhteisö ei todennäköisesti ehtinyt kasvaa 

riittävään pitoisuuteen kokeen aikana. Liuotusjäämänäytteiden perusteella götiittiä 

muodostui puolijatkuvasyötteisessä reaktorissa noin 5 kertaa enemmän kuin 

panosreaktorissa vastaavalla lietetiheydellä [vrt. ~ 25 ja 5 % (w/w)]. Götiitin 

muodostumiseen vaikuttavia tekijöitä on tarpeen tutkia jatkossa tarkemmin. 

Pullokokeissa liuoksia, jotka sisälsivät 25,1 g/L rautasulfaattia (FeSO4 · 7 H2O), 20 

mg/L arseniittia ja 10 % (v/v) ymppiä panosreaktorista sekä vastaavia abioottisia 

kontrolliliuoksia inkuboitiin (150 rpm, 27 °C) 28 päivän ajan. Kokeissa käytettiin kahta 

aloitus-pH:ta, pH 1,7 ja 2,5. Bioottisten liuosten pH:t yhtyivät ja pysyivät samoina 6 

päivän jälkeen. Arseniitti hapettui arsenaatiksi kaikissa pulloissa yhdessä päivässä. 

Liukoisen arseenin pitoisuus bioottisessa liuoksessa, jonka aloitus-pH oli 2,5 laski 4 

mg/L:aan, josta suurin osa (75 %) oli arseniittia. Liuoksen arseniitti- ja 

arsenaattipitoisuudet pysyivät alhaisina kokeen loppuun saakka, jolloin pitoisuudet 

olivat 4 mg/L As(III) ja 1 mg/L As(V). Suurin osa (~81 %) liukoisesta arseenista oli 6 

päivän jälkeen arseniittia myös bioottisessa liuoksessa, jonka aloitus-pH oli 1,7, vaikka 

tässä liuoksessa kaikki arseeni oli pysynyt liukoisessa muodossa. Molemmilla 

hapetusasteilla esiintyvän arseenin konsentraatiot laskivat hitaasti kokeen loppua 

kohden, jolloin konsentraatiot olivat 12 mg/L As(III) ja 3 mg/L As(V). Tulosten 

perusteella liuoksen aloitus-pH:lla on merkittävä vaikutus arseenin pidättäytymiseen ja 

hapetusastejakaumaan. Mikro-organismien rooli näyttää olevan ratkaisevassa osassa 

arseenin pidättymisessä. Työssä ei määritetty, oliko arseenin pidättymisen 

V

päämekanismi sorptio biomassaan vai kerasaostuminen/sorptio mikro-organismien 

rautasulfaatista tuottamien rautaoksidien kanssa. 

Tämä diplomityö osoitti, että nikkelin ja koboltin selektiivinen bioliuottaminen ja 

samanaikainen arseenin liukenemisen minimointi nikkelirikasteesta on mahdollista 

pitämällä pH arvossa 3,0. Suurimmat nikkeli- ja kobolttisaannot saatiin suurimmalla 

tutkitulla lietetiheydellä (15 %), mikä kannustaa tutkimaan jatkossa yhä suurempien 

lietetiheyksien käyttöä. Tässä työssä korostui myös riittävän sekoituksen tärkeys: 

Lakeutunutta ja osittain bioliuottamatonta materiaalia esiintyi 15 ja 10 % 

lietetiheyksisien reaktorien pohjalta panoskokeen lopussa. Sedimentoitunutta ja osin 

bioliuottamatonta materiaalia löytyi myös puolijatkuvasyötteisen reaktorin pohjalta, 

vaikka tällöin lietetiheys oli vain 5 %. Mikro-organismien rooli arseenin pidättymisessä 

kaivosolosuhteissa osoitettiin tärkeäksi. Arseenin pidättymismekanismien suhde mikroorganismien 

läsnäoloon vaatii lisätutkimuksia. 

VI

PREFACE 

This Master of Science Thesis has been done at Tampere University of Technology, in 

the Department of Chemistry and Bioengineering as a part of the ARSENAL (Solutions 

for arsenic control in mining processes and extractive industry) project. The funding for 

the project provided by TEKES is highly appreciated. 

I wish to thank my supervisor and examiner, professor Jaakko Puhakka, for giving 

me the opportunity to work on this very interesting subject and for giving guidance, 

responsibility, inspiration and encouragement during the making of this thesis. I also 

want to thank my other supervisor and examiner, Dr. Minna Peltola, for all the help and 

encouragement she provided during this work and for the patience she showed and the 

fair feedback she gave during the writing process. And Minna, I especially want to 

thank you for just being there and listening when I wanted to clear my thoughts through 

those many incoherent speculative monologues about this and that in your office. You 

know what I’m talking about. I want to thank Raisa Neitola from GTK for the XRD 

analyses and all the friendly correspondence concerning them. Many thanks also to 

Taneli Mikkola from Mondo Minerals for providing me with plenty of NFC to last a 

lifetime and for welcoming all that nasty arsenic waste I produced from it. Aino-Maija 

Lakaniemi deserves a thank-you for those invaluable DGGE instructions. Sarita, thanks 

for the peer support, I truly appreciated it. To the whole staff in the Department of 

Chemistry and Biotechnology I wish to say my kindest thanks for a very enjoyable, 

open, supportive and occasionally even fun working atmosphere. 

Finally, I want to thank my friends for the support during my studies and this thesis 

work. But above all, my biggest, most gigantic, downright gargantuan thanks go to my 

parents who made it possible for me to be who and where I am now. Kiitos, isä ja äiti, 

tämä on omistettu teille! 

Tampere, October 24 th 2012 

Antti Venho 

VII

Kuulivat he virren kumman; 

käsi kädessä 

kulkivat kivistä tietä 

linnunlaulu-laaksoon, 

siinä lähde läikkyväinen, 

vesi vailla pohjaa. 

Katsoivat kuvastimehen; 

kuvan näkivät 

kumpainenkin itsestänsä, 

elon entismuistot 

otsillensa uurtunehet, 

surut sielunsilmäin. 

Koettivat kädellä vettä; 

iho värähti, 

valvahti pyhä vavistus 

sydänkammioissa, 

kulki verta kumpaisenkin 

kasvot kalventaen. 

Virkahteli valju impi: 

"Sydän minulla 

sykkyräksi sylkähtävi, 

kun ma tuumin tuonne 

syöksyä sylihin aallon 

syvän, pohjattoman." 

Vastasi vakava sulho: 

"Veri minulla 

niinkuin seinä seisahtavi 

auvon aavistusta, 

uskallusta onnen uuden; 

sylitysten syöstään!" 

Verhot vaatteiden valahti; 

hetevesihin 

heilahti heleät varret, 

painui alle pinnan, 

nousi kohta päätä kaksi 

virran vienon kalvoon. 

Päätä kaksi kaunokaista, 

hyvä-hymyistä, 

auvon nuoren autuasta, 

unho-onnellista; 

katsoi taivahan sinehen, 

sukeltausi jälleen. 

Eino Leino (1914): Metamorfosi 

VIII

ABBREVIATIONS 

A adenine 

AMD acid mine drainage 

Aox/Arx different arsenite oxidase types 

APHA American Public Health Association 

Arr dissimilatory arsenate reductase 

Ars(A)B arsenite efflux pump 

ArsC detoxifying arsenate reductase 

BSA bovine serum albumin 

C cytosine 

CSTR continuously-stirred tank reactor 

DI de-ionized 

DMA dimethylarsine 

DMAA dimethylarsinic acid 

Eh redox potential 

EPA U.S. Environmental Protection Agency 

EPS extracellular polymeric substance 

G guanine 

GlpF aquaglyceroporin F 

GSH glutathione 

GTK Geological Survey of Finland 

IC ion chromatography 

ILS intermediate leach solution 

LC-HG-AFS liquid chromatography with hydride generation – atomic fluorescence 

spectrometry 

MMA monomethylarsine/monomethylarsonic acid 

NFC nickel flotation concentrate 

PA polyamide 

p.d. pulp density 

Pit phosphate inorganic transport system 

PLS pregnant leach solution 

PP polypropylene 

Pst phosphate specific transport system 

T thymine 

TMA trimethylarsine 

TMAO trimethylarsine oxide 

Topt optimal growth temperature 

TTL Finnish Institution of Occupational Health 

WHO World Health Organization 

IX

XRD X-ray diffraction 

X

TABLE OF CONTENTS 

1 Introduction ........................................................................................................ 1 

2 Microbial Oxidation Mechanisms ...................................................................... 3 

2.1 Mineral Surface-Bacterium-Interaction.............................................................. 3 

2.2 Thiosulfate and Polysulfide Mechanisms ........................................................... 5 

3 Biomining Bacteria ............................................................................................ 8 

3.1 Physiology and Growth Environments ............................................................... 8 

3.2 Arsenic-oxidizers .............................................................................................. 11 

4 Biomining Processes ........................................................................................ 12 

4.1 Process Parameters ........................................................................................... 14 

4.1.1 Microbiological Parameters ...................................................................... 16 

4.1.2 Physicochemical Parameters ..................................................................... 18 

4.2 Irrigation-Type Processes ................................................................................. 20 

4.3 Stirred-Tank Processes ..................................................................................... 22 

5 Arsenic in Biomining ....................................................................................... 24 

5.1 General Characteristics ..................................................................................... 24 

5.2 Microbial Interaction and Resistance ............................................................... 26 

5.3 Removal of Arsenic from Mining Waters ........................................................ 28 

5.3.1 Mechanisms of Arsenic Immobilization with Iron Oxides ....................... 29 

5.3.2 Common Mining-related Iron Oxides ....................................................... 30 

5.3.3 Factors Affecting Arsenic Immobilization with Iron Oxides ................... 30 

6 Materials and Methods ..................................................................................... 32 

6.1 Experimental Outline ....................................................................................... 32 

6.2 Materials ........................................................................................................... 32 

6.2.1 Nickel Flotation Concentrate (NFC) ......................................................... 32 

6.2.2 Bioleaching Cultures and Growth Media .................................................. 33 

6.2.3 Analytical Solutions .................................................................................. 33 

6.3 Experimental Methods and Designs ................................................................. 34 

6.3.1 30 L Batch Experiment ............................................................................. 34 

6.3.2 20 L Semi-continuous Mode Experiment ................................................. 35 

XI

6.3.3 Shake-flask Experiment for the Assessment of the Role of Bacteria in 

Arsenic Speciation and Immobilization .................................................... 35 

6.4 Analytical Methods .......................................................................................... 36 

6.4.1 Redox Potential and pH ............................................................................ 36 

6.4.2 Ferrous Iron ............................................................................................... 36 

6.4.3 Total Iron, Nickel, Cobalt and Arsenic ..................................................... 36 

6.4.4 Arsenic Speciation .................................................................................... 36 

6.4.5 Sulfate ....................................................................................................... 37 

6.4.6 Leach Residue Mineral Composition ........................................................ 37 

6.4.7 Bacterial Communities .............................................................................. 37 

7 Results .............................................................................................................. 39 

7.1 30 L Batch Experiment ..................................................................................... 39 

7.1.1 Bioleaching Results ................................................................................... 39 

7.1.2 Operational Parameters ............................................................................. 43 

7.1.3 Bacterial Community ................................................................................ 49 

7.2 20 L Semi-continuous Mode Experiment ........................................................ 51 

7.2.1 Bioleaching Results ................................................................................... 51 

7.2.2 Operational Parameters ............................................................................. 55 

7.3 Shake-flask Experiment for the Assessment of the Role of Bacteria in Arsenic 

Speciation and Immobilization ........................................................................ 59 

8 Discussion ........................................................................................................ 63 

8.1 30 L Batch Bioleaching of Nickel Flotation Concentrate ................................ 63 

8.2 20 L Semi-continuous Mode Bioleaching of Nickel Flotation Concentrate .... 65 

8.3 The Role of Bacteria in Arsenic Speciation and Immobilization ..................... 66 

9 Conclusions ...................................................................................................... 68 

References ....................................................................................................................... 69 

XII

1 INTRODUCTION 

The modern society is highly dependent on minerals and metals. Like oil reserves, the 

Earth’s mineral deposits are limited and depleting. (Mason et al. 2011) High-grade ore 

deposits are becoming scarcer or increasingly harder to reach. Metal recovery from lowgrade 

ores is rarely economically viable by traditional metallurgical methods such as 

roasting or smelting. Compared to these traditional methods, more recent 

biohydrometallurgical approaches, commonly termed biomining, offer a more costeffective 

and potentially environmentally friendlier method for metal extraction from 

low-grade ores and concentrates. (Rawlings et al. 2003) 

Biomining utilizes the ability of iron- and sulfur-oxidizing microorganisms to 

solubilize minerals and heavy metal compounds on an industrial scale. The microbial 

phenomenon behind biomining is ancient and naturally occurring. It has simply been 

harnessed for mining purposes, and optimized to speed up the process and yield costeffectiveness. 

Currently, biomining methods are mostly used to recover metals such as 

copper, nickel, zinc, cobalt, uranium and gold from low-grade ores and concentrates, 

whose processing is not economically viable by conventional methods. (Pradhan et al. 

2008; Rawlings 2002) Compared to the traditional pyrometallurgical methods, 

biomining is not nearly as energy-intensive and the release of sulfur oxides and other 

gaseous emissions is much lower than that of traditional roasting or smelting processes. 

(Rawlings et al. 2003) 

The term ‘biomining’ comprises two distinct processes, i.e., bioleaching and 

biooxidation. In bioleaching, the role of the microorganisms is to directly oxidize 

insoluble metal sulfides into soluble form, usually sulfates. In biooxidation, the 

microbial oxidation is not aimed directly at the target metal compound, but rather at the 

refractory mineral matrix encapsulating the desired metal mineral. The goal is to reveal 

the metal mineral for further processing. Biooxidation is used mainly for pretreatment 

of gold and silver ores prior to cyanide treatment for metal recovery. (For reviews, see 

Rohwerder et al. 2003; Bosecker 1997) 

Although emissions from biomining processes are smaller than from the traditional 

pyrometallurgical processes, acid mine drainage (AMD) remains a significant 

environmental concern related to all mining activity. The uncontrolled microbial 

oxidation of mine tailings can lead to the release of highly acidic wastewater with 

elevated metal and organic contaminant concentrations into groundwater. (Johnson & 

Hallberg 2005; Bosecker 2001) However, compared to tailings from roasting or 

smelting, tailings from biomining operations are less biologically active, which renders 

them less susceptible to formation of AMD (Rawlings et al. 2003). A significant 

contaminant in mine waters is arsenic. It is a highly toxic metalloid abundant virtually 

1

everywhere in the Earth’s crust, making it a worldwide environmental threat largely 

originating from mining activity (Williams 2001). 

This thesis investigates selective bioleaching of nickel flotation concentrate (NFC) 

with the view to extract the high-value nickel and cobalt while immobilizing arsenic 

with elevated pH. NFC is a side-product of talc production relatively rich in nickel and 

cobalt but also arsenic. The NFC used in this study originated from the Mondo Minerals 

LLC mine site in Sotkamo, Finland. Based on earlier studies in shake-flasks (Wakeman 

et al. 2011), three 30 L continuously-stirred tank reactor (CSTR) experiments were 

conducted in batch-mode at pH 3.0 with 15, 10, and 5 % (w/v) pulp densities (p.d.) of 

NFC. As a subsequent preliminary study for a continuous-flow design, a 20 L CSTR 

with 5 % (w/v) p.d. was operated at pH 3.0 in semi-continuous mode with a retention 

time of 10 days. A shake-flask experiment without NFC was also conducted with a view 

to assess the role of chemical and microbial oxidation and immobilization of arsenic in 

the bioleaching process. 

2

2 MICROBIAL OXIDATION MECHANISMS 

The key role of the microorganisms taking part in biomining processes is oxidation. 

Iron-oxidizing microorganisms such as Acidithiobacillus ferrooxidans (Temple & 

Colmer 1951) and Leptospirillum ferrooxidans (Sand et al. 1992) oxidize ferrous iron 

[Fe(II)] found in abundance in the Earth’s crust (McKelvey 1960). At the same time, 

sulfur-oxidizing microbes like At. thiooxidans (Donati et al. 1996) and At. caldus 

(Rawlings et al. 1999a,b) create sulfate from reduced sulfur compounds in the ground or 

from sulfur intermediates generated by microbial oxidation of metal sulfides (e.g., 

chalcopyrite, CuFeS2 or zinc sulfide, ZnS) . The ferric iron created by iron-oxidizing 

microorganisms (Fowler et al. 1999) chemically oxidizes the insoluble metal sulfide and 

dissolves it. In the case of acid-soluble metal sulfides, also protons from the formation 

of sulfate by sulfur-oxidizers participate in the chemical attack on the metal sulfides 

(Chen & Lin 2004). 

2.1 Mineral Surface-Bacterium-Interaction 

Debate on the importance of bacterial attachment on the mineral surface and whether 

microorganisms can oxidize mineral sulfides directly via enzymatic attack lasted for 

decades (Porro et al. 1997; Bennett & Tributsch 1978; Silverman 1967), and numerous 

review articles have been written on the subject as an attempt to clarify the matter 

(Crundwell 2003; Tributsch 2001; Sand et al. 1999). During the past decade, however, 

it has been shown that the direct enzymatic attack does, in fact, not exist (Gehrke et al. 

1998; Sand et al. 1995). A division of the indirect attack mechanism into contact and 

non-contact mechanisms presented in Figure 2.1 has been suggested (Rawlings 2002; 

Tributsch 2001). Figure 2.1 illustrates microbial interaction with a sulfide mineral (here 

chalcopyrite, CuFeS2). Note that the microorganisms do not necessarily need to be 

attached to the mineral surface. 

3

Figure 2.1: Microbial interaction with sulfide minerals with chalcopyrite 

(CuFeS2) as an example. The indirect noncontact (A) and indirect contact (B) 

mechanisms. In (B), the microbe is in contact with the mineral surface via an 

extracellular polymeric substance (EPS) layer but mineral oxidation proceeds via 

ferric iron attack. (Watling 2006) 

In addition to the indirect contact and non-contact mechanisms presented in Figure 

2.1, there exists a third mechanism called cooperative leaching. In cooperative leaching, 

colloidal sulfur and sulfur globules, found in large amounts inside the bacterial 

extracellular polymeric substance (EPS) layer (Rojas et al. 1995), are released into the 

environment as a result of wasteful feeding of the microorganisms in contact with the 

mineral sulfide. The released energy-carrying colloids and globules can then be utilized 

by bacteria that do not have access to the mineral surface. The mechanism is suggested 

to be a survival mechanism that maintains bacterial life in the surrounding medium 

when sulfide surface area is limited. (Rojas-Chapana et al. 1998) 

4

The EPS layer 

Bacterial attachment to the mineral surface in bioleaching occurs via an EPS layer 

(Pogliani & Donati 1999). The EPS layer consists of sugars, lipids and free fatty acids 

and is a pivotal factor in biofilm formation for a wide variety of microorganisms 

(Gehrke et al. 1998). Not much attention had been paid to the EPSs excreted by 

biooxidative bacteria until approximately 15 years ago when Gehrke et al. (1998) made 

some pivotal discoveries concerning the biooxidation mechanisms and the respective 

role of the EPS layer. In addition to acting as a bacterial anchor to mineral surface, the 

EPS layer has been shown to contain ferric iron in large concentrations (Gehrke et al. 

1998). It also serves as a reaction space that enhances bioleaching of the mineral sulfide 

(Fowler et al. 1999). The study by Gehrke et al. (1998) also revealed that EPSs from At. 

ferrooxidans grown on sulfur, i.e., devoid of ferrous iron, did not attach to pyrite (FeS2), 

which suggested that ferric iron complexed by EPSs is a prerequisite to cell attachment 

on sulfide minerals. The results were supported by Pogliani and Donati (1999). The 

finding also shed some light on the debate concerning the existence of the direct 

mechanism. It is the ferric iron ions in the EPS layer and not bacterial enzymes that are 

responsible for sulfide oxidation when the bacterium is in contact with the mineral 

surface (Gehrke et al. 1998; Sand et al. 1995). 

2.2 Thiosulfate and Polysulfide Mechanisms 

The key mechanism in biomining is the chemical oxidation of metal sulfides by ferric 

iron created by ferrous iron-oxidizing microorganisms (For a review, see Rawlings 

2002). This mechanism is present regardless of whether the microorganisms are in 

contact with the mineral surface or not. It also applies to both acid-soluble and acidinsoluble 

metal sulfides. (Schippers & Sand 1999; Gehrke et al. 1998; Sand et al. 1995) 

In the case of acid-soluble metal sulfides, oxidation is carried out by a combination of 

ferric ions and protons (Schippers & Sand 1999). The reaction of microbial 

(re)oxidation of ferrous iron to ferric iron proceeds as follows (Breed & Hansford 

1999): 

4 Fe + O 2 + 4 H + → 4 Fe 3+ + 2 H 2O (2.1) 

Due to differences in crystal structure between different metal sulfides, ironoxidizing 

bacteria have developed different mechanisms to oxidize different sulfides 

(Tributsch 2001). Schippers and Sand (1999) discovered that sulfide oxidations can 

proceed via two different intermediates and thus via two different oxidation 

mechanisms (Figure 2.2): the thiosulfate mechanism in the case of acid-insoluble metal 

sulfides and the polysulfide mechanism for acid-soluble metal sulfides. 

5

Figure 2.2: Thiosulfate and polysulfide mechanisms in biooxidation of a metal 

sulfide (MS). Tf, Tt and Lf represent Acidihiobacillus ferrooxidans, At. 

thiooxidans and Leptospirillum ferrooxidans, respectively. Dashed lines indicate 

occurrence of intermediate sulfur compounds. (Schippers and Sand 1999) 

Using pyrite as an example, Schippers and Sand (1999) presented the following 

reactions that take place in the thiosulfate mechanism, with thiosulfate as an 

intermediate: 

FeS2 + 6 Fe3+ 2- 2+ + + 3 H2O → S2O3 + 7 Fe + 6 H (2.2) 

2- 3+ 2- 2+ + S2O3 + 8 Fe + 5 H2O → 2 SO4 + 8 Fe + 10 H (2.3) 

In the polysulfate mechanism, oxidation of the metal sulfide (MS) occurs as a 

combination of ferric ion and proton attack with polysulfate (H2Sn), and consequently 

sulfur (≥ 99% S8), as intermediates: 

MS + Fe 3+ + H + → M 2+ + 0.5 H 2S

for further solubilization of metal sulfides. The polysulfide mechanism also provides an 

explanation to why sulfur-oxidizing microorganisms can dissolve certain metal sulfides 

but not others. (Schippers & Sand 1999) 

7

3 BIOMINING BACTERIA 

In all bioprocesses, including biomining, the process conditions largely determine the 

composition of the microbial community. As commercial biomining processes are 

usually operated with indigenous microbial consortia under different mineral 

compositions and climates or process conditions, the diversity of microorganisms in the 

biomining environment is large. However, only a small fraction of the contributing 

species have been successfully isolated and studied. (Hallberg & Johnson 2003) 

3.1 Physiology and Growth Environments 

Current commercial biomining processes operate mostly under acidic conditions (pH < 

3) and at or below 40 °C (for a review, see Rawlings 2002). For this reason, the 

microorganisms studied the most are (extremely) acidophilic or acid-tolerant 

mesophiles or moderate thermophiles (for a review, see Watling 2006). Nevertheless, 

many future biomining processes will most likely operate in temperatures above 60 °C 

(for a review, see Rawlings 2002). The reason for this is obvious as many ores dissolve 

more rapidly in high temperatures (for a review, see Watling 2006). In addition, 

chalcopyrite (CuFeS2) leaching, for example, requires a temperature above 60 °C for 

the reaction rate to reach an economically viable level (Rodrı́guez et al. 2003). 

Currently considered to be the key microorganisms capable of iron and sulfur oxidation 

at high temperatures are those of the domain Archae. The most extensively studied 

species belong to the genera Sulfolobus and Metallosphaera, e.g., S. metallicus (Huber 

& Stetter 1991) and M. sedula (Huber et al. 1989), but also other genera of significance 

have been reported (Mikkelsen et al. 2006). However, as the thermophilic 

microorganisms are out of the scope of this thesis work, only the biomining bacteria 

will be discussed in more detail here. 

Characteristics of some of the most commonly reported microorganisms associated 

with biomining processes are listed in Tables 3.1a and 3.1b. As described in Chapter 2, 

the iron-oxidizers play the key role in biomining. Acidithiobacillus ferrooxidans 

(Halinen et al. 2012; Schrenk et al. 1998; Devasia et al. 1993; Pronk et al. 1992; Sugio 

et al. 1985) and Leptospirillum ferrooxidans (Rawlings et al. 1999b; Pizarro et al. 1996; 

Goebel & Stackebrandt 1994) are the most widely reported iron-oxidizing species in 

processes operating under 40 °C. A considerable portion of L. ferrooxidans reports must 

actually be those of its more recently distinguished close relative, L. ferriphilum (Coram 

& Rawlings 2002). It is noteworthy that not all microorganisms listed are capable of 

iron oxidation (e.g. At. caldus or At. thiooxidans). These species are, however, capable 

of sulfur oxidation. As described in Chapter 2, most valuable metals in the ground exist 

as sulfides. Oxidation of the reduced sulfur from these sulfides produces sulfuric acid 

(protons and sulfate), which facilitates the growth of acidophilic microorganisms as well 

as leaching of acid-soluble ores. 

8

FAnaer facultatively anaerobic 

a Highest pH tested; A autotrophic; FA facultatively autotrophic; M mixotrophic; Aer aerobic; FAer facultatively aerobic; 

Alicyclobacillus tolerans Fe, S < 20 ... 55 1.5 ... 5.0 M/FAer Schippers 2007; Karavaiko et al. 2005 

(37 ... 42) (2.5 ... 2.7) 

Alicyclobacillus disulfidooxidans Fe, S 4 ... 40 0.5 ... 6.0 FA/FAnaer Schippers 2007; Karavaiko et al. 2005; 

(35) (1.5 ... 2.5) Dufresne et al. 1996 

Acidithiobacillus thiooxidans S 2 ... 37 0.5 ... 5.5 A/Aer Schippers 2007; Karavaiko et al. 2006; 

(28 ... 30) (2.0 ... 3.0) Hallberg et al. 2010; Kelly & Wood 2000 

Acidithiobacillus ferrooxidans Fe, S 2 ... 37 1.3 ... 4.5 A/FAer Schippers 2007; Karavaiko et al. 2006; 

(30 ... 35) (2.5) Hallberg et al. 2010; Kelly & Wood 2000; 

Temple & Colmer 1951 

(27 ... 32) (2.5) 

Acidithiobacillus ferrivorans Fe, S < 5 ... 37 1.9 ... 3.4 a 

FA/FAnaer Hallberg et al. 2010 

Acidithiobacillus caldus S 32 ... 52 1.0 ... 3.5 FA/Aer Schippers 2007; Karavaiko et al. 2006; 

(45) (2.0 ... 2.5) Kelly & Wood 2000; Pizarro et al. 1996 

Acidithiobacillus albertensis S 25 ... 30 2.0 ... 4.5 A/Aer Schippers 2007; Karavaiko et al. 2006; 

(25 ... 30) (3.5 ... 4.0) Bryant et al. 1983; Kelly & Wood 2000 

9 

Table 3.1a : Characteristics of common microorganisms relevant to biomining processes 

Microorganism Fe/S T (°C) pH Growth Reference 

oxidation growth range growth range metabolism/ 

(optimal) (optimal) respiration 

Acidianus brierleyi (Archaea) Fe, S 45 ... 75 1.0 ... 6.0 FA/FAer Schippers 2007; Karavaiko et al. 2006; 

(~70) (1.5 ... 2.0) Segerer et al. 1986

Thiomonas delicata S 20 ... 45 1.5 ... 7.2 FA/Aer Schippers 2007; 

(30 ... 36) (3.5 ... 4.0) Battaglia-Brunet et al. 2011; Huber & Stetter 

1990 

A autotrophic; FA facultatively autotrophic; M mixotrophic; Aer aerobic; FAer facultatively aerobic 

Sulfolobus metallicus (Archaea) Fe, S 50 ... 75 1.0 ... 4.5 A/Aer Schippers 2007; Karavaiko et al. 2006; 

(65) (2.0 ... 3.0) Melamud et al. 2003 

Sulfobacillus thermotolerans Fe, S 20 ... 60 1.2 ... 5.0 M/Aer Schippers 2007; Karavaiko et al. 2006; 

(40) (2.0 ... 2.5) 

Sulfobacillus thermosulfidooxidans Fe, S 20 ... 60 1.5 ... 5.5 M/Aer Schippers 2007; Karavaiko et al. 2006; 

(45 ... 48) (~2) Zhuravleva et al. 2008 

Sulfobacillus sibiricus Fe, S 17 ... 60 1.1 ... 3.5 M/Aer Schippers 2007; Karavaiko et al. 2006; 

(55) (2.2 ... 2.5) Melamud et al. 2003 

Sulfobacillus acidophilus Fe, S < 30 ... 55 N/A M/FAer Schippers 2007; Norris et al. 1996 

(45 ... 50) (~2) 

Metallosphaera sedula (Archaea) S 50 ... 80 1.0 ... 4.5 FA/Aer Schippers 2007; Karavaiko et al. 2006; 

(75) (2.0 ... 3.0) Huber et al. 1989 

Leptospirillum ferrooxidans Fe 2 ... 37 1.1 ... 4.0 A/Aer Schippers 2007; Karavaiko et al. 2006; 

(28 ... 30) (1.5 ... 3.0) Rawlings 2008 

10 

Table 3.1b : Characteristics of common microorganisms relevant to biomining processes 

Microorganism Fe/S T (°C) pH Growth Reference 

oxidizer growth range growth range metabolism/ 

(optimal) (optimal) respiration 

Leptospirillum ferriphilum Fe 30 ... 37 N/A A/Aer Schippers 2007; Coram & Rawlings 2002; 

(N/A ... 45) (1.3 ... 1.8) Ojumu & Petersen 2011

Therefore also the microorganisms that do not directly solubilize valuable metals play 

an important role in a well-functioning biomining process. (Schippers & Sand 1999) 

The fact that most of the key microbial species are autotrophs makes the 

maintenance of the biomining culture simple (Rawlings 2007). A typical biomining 

process at best only requires acidic water as a growth medium and an evenly distributed 

supply of air for CO2 and oxygen (Rawlings 2002). Energy and some nutrients are 

provided by the minerals (Rawlings & Johnson 2007) but if necessary, nutrients can 

also be added as inexpensive fertilizer chemicals (Rawlings 2007). 

3.2 Arsenic-oxidizers 

Only few reports have been published of microorganisms capable of arsenite [As(III)] 

oxidation in mining environments. Members of the main biomining genera 

Acidithiobacillus and Leptospirillum are not able to oxidize arsenite (Battaglia-Brunet et 

al. 2002a). Some strains of the genus Thiomonas such as T. delicata (Battaglia-Brunet 

et al. 2011) have been shown to oxidize arsenite under chemoautotrophic and 

moderately acidophilic conditions. However, the growth optima for the strains are in the 

pH range of 4 … 7.5, suggesting they would not persist in the low pH environment 

typical of biomining processes. (Battaglia-Brunet et al. 2006) Santini et al. (2000) 

isolated from an Australian gold mine the Agrobacterium/Rhizobium sp. strain NT-26, a 

chemolithoautotrophic bacterium strain capable of arsenite oxidation. However, the 

bacterium prefers growth in circumneutral pH and thus will most likely not be directly 

associated to actual acidic biomining process environments. 

The biooxidation of arsenic-bearing minerals such as arsenopyrite releases 

considerable amounts of arsenic into the bioleaching solution. Therefore, although not 

able to oxidize arsenite, the main biomining microorganisms have developed a high 

resistance to arsenic. The most famous example of this is the resistance generated by L. 

ferriphilum and At. caldus in the CSTRs of the Fairview Mine, South Africa. The 

species developed a tolerance to arsenic from approximately 1 g/L to 13 g/L total-As 

over the course of a few years. (Dew et al. 1993) Tuffin et al. (2006) isolated 

transposons containing the arsenic resistance genes from both strains. The transposons 

were shown to be related to each other but not closely enough for the resistance gene to 

have been transferred from one strain to the other, i.e., the strains developed their 

resistance independently. Also the isolate HGM 8 of At. ferrooxidans obtained by Dave 

et al. (2008) from the Hutti gold mine in India exhibited resistance to arsenite up to 14.7 

g/L. The ability of some biomining organisms to slowly adapt to high arsenic 

concentrations, as well as other toxic elements, e.g., heavy metals is an important factor 

when considering the biomining consortium as discussed later in Chapter 4.1.1. 

11

4 BIOMINING PROCESSES 

The role of acidophilic microorganisms in oxidizing metal sulfides in mining 

environments was recognized in the late 1940’s (Colmer & Hinkle 1947). One of the 

first utilizations of biooxidation mechanisms was reported by Zimmerley et al. (1958; 

U.S. Patent 2,829,964). Since then, mining processes utilizing biohydrometallurgical 

methods for metal extraction have commenced all over the world. Table 4.1 lists some 

locations with on-going full-scale biomining processes. 

Generally the processes can be divided into irrigation type and stirred-tank type 

processes. The process and design depend on economical preferences and production 

capacity needs. For example, it may be economically viable to utilize high-throughput 

tank reactors for the extraction of high-value metals such as gold whereas slower but 

also cheaper methods such as heap or dump leaching are a more economical option for 

the extraction of lower-value metals. Irrigation type processes, subdivided into dump, 

heap and in situ processes, offer a more economical solution with a large ore processing 

capacity. However, reaction rates and metal recovery percentage are higher when 

applying stirred-tank processes in carefully adjusted and controlled continuous-flow 

bioreactors. Drawbacks of the stirred-tank processes are high construction and operation 

costs and limited processing volume. For this reason, stirred-tank processes are mainly 

used for low-grade ores and concentrates. (For reviews, see Watling 2006; Olson et al. 

2003; Rawlings 2002) The metals extracted by far the most is copper by means of 

bioleaching and gold by biooxidation (Rawlings 2002) but commercial scale plants also 

for the extraction of nickel, zinc and cobalt (Riekkola-Vanhanen 2010) as well as 

uranium (Talvivaara 2012b; for a review, see Muñoz et al. 1995) have been introduced. 

12

Table 4.1: Examples of biomining plants currently in operation 

Location 

Fairview, South 

Africa 

Saõ Bento, 

Brazil 

Harbour Lights, 

Australia 

Wiluna, 

Australia 

Metals 

Extracted Process Type 

Au BIOX TM tank 

leaching 


leaching 


leaching 


leaching 

Sansu, Ghana Au BIOX TM tank 

leaching 

Tamboraque, 

Peru 

Fosterville, 

Australia 

Suzdal, 

Kazakhstan 

Gold Quarry, 

Nevada, USA 

Agnes, South 

Africa 

Talvivaara, 

Finland 

Cerro Colorado, 

Chile 

Quebrada 

Blanca, Chile 


leaching 


leaching 


leaching 

Au BIOPRO TM 

heap leaching 

Au GEOCOAT® 

heap 

biooxidation 

Ni, Cu, Zn, 

Co 

Year of 

Commissioning Reference 

13 

1986 van Aswegen et al. (2007) 








1999 Logan et al. (2007) 

2003 Harvey & Bath (2007) 

Heap leaching 2009 Talvivaara (2012a) 

Cu Heap leaching 1993 Domic (2007) 


Zaldívar, Chile Cu Heap leaching 1995 Domic (2007) 

Ivan, Chile Cu Heap leaching 1994 Domic (2007) 

Chuquicamata, 

Chile 

Collahuasi, 

Chile 

Dos Amigos, 

Chile 

La Escondida, 

Chile 

Spence Mine, 

Chile 

Cu Dump leaching 1994 Domic (2007) 



Cu Heap leaching 1990, heap 

leaching since 

1999 

Domic (2007) 

Cu Heap leaching 2007 Domic (2007)

4.1 Process Parameters 

As with any bioprocess, careful consideration of process conditions is necessary for 

successful operation. Current commercial biomining processes that operate below 40 °C 

are rather forgiving in terms of biological conditions as biooxidative microorganisms 

are ubiquitous, i.e., they occur naturally in most environments. As such, no inoculation 

or upkeep of a defined culture is required. However, only by careful planning and 

control of the process conditions is it possible to reach high ore processing rates and 

improved economical viability. (For reviews, see Watling 2006; Rawlings et al. 2003) 

Brandl (2001) lists the most important factors affecting biomining process operation 

(Table 4.2). 

14

Table 4.2: Factors affecting biomining processes. Adapted from (Brandl 2001). 

Factor Parameter 

Physicochemical parameters of a bioleaching temperature 

environment pH 

redox potential 

water potential 

oxygen content and availability 

carbon dioxide content 

mass transfer 

nutrient availability 

iron(III) concentration 

light 

pressure 

surface tension 

presence of inhibitors 

Microbiological parameters of a bioleaching microbial diversity 

environment population density 

microbial activities 

spatial distribution of microorganisms 

metal tolerance 

adaptation abilities of microorganisms 

Properties of the mineral to be leached mineral type 

mineral composition 

mineral dissemination 

grain size 

surface area 

porosity 

hydrophobicity 

galvanic interactions 

formation of secondary minerals 

Processing leaching mode (in situ, dump, heap or tank) 

pulp density 

stirring rate (in case of tank operations) 

heap geometry (in case of heap leaching) 

A selected set of parameters the most relevant with respect to this thesis work are 

reviewed in this chapter. An overview on the principles of different processing methods, 

i.e., in situ, dump, heap and tank bioleaching/biooxidation is also given. 

15

4.1.1 Microbiological Parameters 

In terms of biomining processes that operate in the mesophilic temperature range, 

mineral sulfide oxidizing microorganisms are ubiquitous, i.e., the biooxidative consortia 

will naturally form in the process without inoculation. Also, the harsh acidic process 

conditions keep the growth environment rather clear of any unwanted competing 

populations. (Rawlings 2002) In fact, many current commercial scale bioleaching 

operations rely on indigenous bioleaching microorganisms as inoculation is usually not 

necessary for an effective bioleaching consortium to develop in the process (Riekkola- 

Vanhanen 2010; He et al. 2008; Lavalle et al. 2008). Therefore, manual optimization of 

microbiological parameters in biomining processes has not been of highest priority. 

However, studies have been done on the role of different microorganisms and their 

interaction in different consortia. Also the use of an indigenous inoculum has been 

challenged with studies on carefully selected inocula and consortia and different 

inoculation principles. (For a review, see Rawlings & Johnson 2007) 

Inoculation strategies: top-down vs. bottom-up 

Generally there are two different approaches to inoculation of biomining processes. 

Rawlings and Johnson (2007) refer to them as top-down and bottom-up approaches. The 

top-down approach stems from the idea of inoculating the process with a wide selection 

of different species and relying on the strains with the best adaptability to the process 

conditions to dominate and form the optimal microbial consortium. The bottom-up 

approach relies more on logical selection of the microbial species, i.e., inoculating the 

process with hand-picked species that would be considered beneficial to process 

efficiency. 

The top-down approach requires a vast and diverse enough inoculant consortium 

where as many different species as possible capable of contributing to sulfide mineral 

dissolution are represented (Rawlings & Johnson 2007). These species include iron- and 

sulfur-oxidizing microorganisms such as Acidithiobacillus spp. and Leptospirillum spp. 

as well as possible heterotrophic species that may contribute to the process by 

metabolizing organic substances that might otherwise inhibit the growth of the 

lithotrophic iron- and sulfur-oxidizers (Okibe & Johnson 2004). Inocula used for 

inoculation of processes with elevated temperatures should also contain 

thermoacidophilic archae (Logan et al. 2007; Plumb et al. 2007; Mikkelsen et al. 2006). 

The species can be obtained either by enriching from natural sources or from preexisting 

biomining processes (Rawlings & Johnson 2007). 

The bottom-up approach is based on careful pre-selection of individual species. The 

idea is to include only the species that are necessary for the most efficient oxidation of 

minerals. Limiting the microbial diversity in the process prevents the enrichment of 

harmful or unbeneficial species that might, e.g., produce excess amounts of acid. 

(Rawlings & Johnson 2007) As a rule of thumb, a biomining process should contain at 

least one species capable of iron-oxidation and one species capable of sulfur-oxidation 

(Schippers & Sand 1999). L. ferriphilum, for example, is a very effective iron-oxidizer 

16

at temperatures around 40 °C but it is unable to oxidize sulfur. It is therefore often 

accompanied by At. caldus, a moderately thermophilic sulfur-oxidizer incapable of ironoxidation. 

L. ferriphilum then oxidizes the iron and At. caldus keeps up the acidic 

conditions by oxidizing the sulfur to sulfuric acid. (Rawlings et al. 1999b) In 

commercial bioleaching processes, however, inoculation with selected species is rarely 

necessary and an indigenous consortium can outcompete a designed consortium in 

terms of both bioleaching efficiency and adaptability (Bryan et al. 2011). 

Effect of reactor type on the microbial population 

With many bioprocesses, the idea of constructing the microbial community species-byspecies 

is impossible, but not with continuous-flow stirred-tank biomining processes. In 

rather extreme acidic conditions and continuous flow of material, it is usually only a 

very limited number of species that dominate the microbial consortium (Battaglia- 

Brunet et al. 2002b; Rawlings et al. 1999b). Further in steady-state conditions in a 

continuous-flow reactor, the microbial population keeps evolving and adapting to the 

conditions for years instead of weeks (Rawlings & Silver 1995). Thus species that adapt 

slower but might eventually reach a higher rate of sulfide oxidation than the initially 

faster-adapting species might be washed out too soon. This makes identifying the 

ultimately optimal population a very challenging task and requires long-term research. 

Another thing to consider is the fact that the processed ore is not sterile. Thus, it is 

impossible to prevent the contamination of the inoculated consortium with other species 

(Rawlings & Johnson 2007). Therefore, if a designed consortium were to be used to 

inoculate a bioleaching process, the top-down approach would be most useful with 

processes operated in very extreme conditions, e.g., very low pH (Yahya & Johnson 

2002) and/or high temperature (Mikkelsen et al. 2006), where the microbial growth 

niches are scarce. 

Monitoring the microbial population in continuous-flow stirred-tank processes is 

relatively straightforward as the process environment is totally homogenous due to 

constant mixing. Additionally, the constant flow and otherwise demanding environment 

make for conditions where only a couple of species dominate. (Rawlings et al. 1999b) 

Heaps, on the other hand, are not homogenous and thus contain varying environments 

for microbial growth. For example, temperature and the amount of air may vary 

considerably within a heap. (Casas et al. 1998) Thus compiling a prescribed inoculum 

with hand-picked species for a heap reactor is neither possible nor something to pursue. 

Instead, biomining heaps are usually inoculated either in or just prior to the stacking of 

the ore in heaps in order to ensure an even distribution of microorganisms throughout 

the heap from the start. Just recirculating the raffinate solution (i.e., the heap effluent) 

and waiting for the indigenous microorganisms to enrich in the heap is enough. (Plumb 

et al. 2007) Microbial diversity inside the heap is not a well studied subject as most 

microbiological samples are taken from the effluent solution and not the ore itself 

(Plumb et al. 2007; Lizama 2001). 

In the future, more research will definitely be done on the microbial adaptation to 

stirred-tank process environments. For example, it has been shown that after years of 

17

operation in a stirred-tank for processing of arsenopyrite-containing refractory goldbearing 

ore, the dominating species L. ferriphilum and At. caldus had developed a 

strong arsenic-resistance. (Tuffin et al. 2006) The mechanism through which the 

bacteria acquired the resistance genes from the horizontal gene pool is still unknown. 

Understanding of the mechanism would make it a lot easier to, e.g., enrich 

microorganisms that have high resistance to heavy metals or other harmful substances. 

Also the efficiency of different consortia inside heap reactors should be studied and 

engineering approaches that maximize the proportion of the most efficient 

microorganisms and growth-supporting conditions should be developed. (For a more 

comprehensive discussion on the subject, see the review from Rawlings & Johnson 

2007) 

4.1.2 Physicochemical Parameters 

Although microbial dissolution of sulfide minerals is a naturally occurring phenomenon 

exhibited by ubiquitous microorganisms, industrial applications of the phenomenon 

need to be optimized for maximized economical viability and minimized environmental 

effects. This requires understanding of the influence of physical and chemical factors on 

the process efficiency. 

Temperature 

Temperature is an obvious factor to affect the mineral sulfide dissolution process. For 

example, different microorganisms have different optimal growth temperatures, and 

thus the selection of the consortium with the fastest mineral solubilization rate is 

strongly temperature-dependent. (Plumb et al. 2007) Most commercial biomining 

processes operate mostly at or below 40 °C but many future biomining processes will 

most likely operate at elevated temperatures above 60 °C. (for a review, see Rawlings 

2002) Process temperature plays an important role especially in the leaching of 

chalcopyrite (CuFeS2) ores or concentrates, as an elevated temperature is crucial for 

economically viable leaching rates (Watling 2006). Elevated temperature also works to 

prevent the precipitation of iron-hydroxy sulfates such as jarosites on mineral surfaces. 

Jarosite precipitation on mineral surface blocks access from microorganisms to the 

mineral surface, effectively inhibiting biooxidative reactions. (Tshilombo et al. 2002) 

Mineral surface passivation by jarosite precipitation is a major concern especially 

related to chalcopyrite dissolution (Watling 2006). 

For the processing of other ores and concentrates, the temperature is not quite as 

important, but as the biooxidation reactions are exothermic, i.e., heat-generating, the 

process temperature tends to rise especially in the industrial-scale (Plumb et al. 2007). 

The exothermic nature of the reactions makes it possible, e.g., to operate bioheap 

reactors even in cold climate regions such as the Talvivaara deposit in Sotkamo, 

Finland, without external heating (Riekkola-Vanhanen 2010). On the other hand, some 

processes, especially those of the stirred-tank kind, may need to be cooled down in 

order to maintain process conditions supporting the optimal microbial consortium and 

metal extraction rate [Schaffner et al. 2000 (U.S. Patent 6,096,113)]. 

18

pH 

Most, if not all, industrial biomining processes are operated in a very acidic 

environment (pH ≈ 1…2) (Watling 2006; Rawlings 2002). As described in Chapter 2.2, 

the biooxidation of sulfur results in the formation of sulfuric acid, which lowers the pH 

of the biomining environment. The acidic conditions favor acidophilic microorganisms 

that are efficient in oxidation of inorganic iron- and sulfur-compounds. Low pH also 

promotes chemical dissolution of acid-soluble sulfide minerals. (Schippers & Sand 

1999) Yahya and Johnson (2002) point out that extremely low pH (< 1.5) hinders the 

formation of mineral-passivating iron sulfate (jarosite) precipitation on the mineral 

surface. In contrast, Halinen et al. (2009) recommend the heap leaching of black schist 

ore to be conducted in pH 2.0 instead of the lower pH of 1.5. Thus, the concept of 

optimal process pH is ambiguous as it depends on other process parameters, most 

importantly the mineral composition of the ore. 

Ferric ion concentration 

Redox potential in biomining solutions is proportional to the ferric to ferrous iron ratio, 

i.e., the more ferric iron in the solution compared to ferrous iron, the more positive the 

redox potential (May et al. 1997). At. ferrooxidans and L. ferrooxidans, for example, 

usually generate redox potentials in the range of +600 mV to up to +900 mV (Rawlings 

et al. 1999a). Redox potential thus acts as an indicator of microbial oxidation activity. 

However, Tshilombo et al. (2002) demonstrated on chalcopyrite that mineral surface 

passivation by iron-hydroxy compounds such as jarosites is increased with increasing 

redox potential. Yahya and Johnson (2002) reported an extremely acidophilic 

Sulfobacillus sp. capable of efficient pyrite oxidation in pH below 1 with a redox 

potential of < +650 mV, effectively reducing jarosite formation. 

Mineral composition and gangue minerals 

As biomining processes are mostly used for the treatment of low-grade ores, the 

proportion of gangue minerals in the ores is usually rather high. Most gangue materials 

such as quartz, mica, chlorite, potassium- and calcium-feldspar are acid-consuming. The 

sulfuric acid production from sulfur-oxidation reactions may not be enough to cover the 

acid consumption by gangue minerals. As a result, acid has to be added to the process in 

order to maintain acidic conditions, which can be a major source of costs. (Rawlings et 

al. 2003) Dissolution of gangue silicates also results in the release of structural elements 

such as Al, Fe, K, Mg and Si and trace elements, which can inhibit the sulfide mineral 

dissolution. Dopson et al. (2008) reported the inhibition of microbial growth by 

dissolved fluorine and leachate coagulation by dissolved silicates. 

Also the inhibitory effects of the desired metals have to be taken into account. 

Nickel tolerance, for example, varies not only between species but between individual 

strains as well. The reported resistance of different strains of At. ferrooxidans to nickel 

varies between 6 mg/L and as much as 50 g/L (Watling 2008). Microorganisms also 

tend to adapt to the process conditions, which includes generating a stronger resistance 

to metals and other inhibitory substances over time. For example, At. caldus and L. 

ferriphilum found in biooxidation tanks in Fairview, South Africa, have been reported to 

19

generate a strong resistance to arsenic (up to 13 g/L) in the course of several years. 

(Rawlings 2008) 

4.2 Irrigation-Type Processes 

Irrigation-type processes can be divided into three groups: heap, dump and in situ 

processes. They are mostly used for bioleaching of large amounts of ore. In heap 

leaching, the ore is crushed, agglomerated and stacked into carefully designed heaps 

2…10 m high. Prior to stacking, the crushed ore is cured (i.e., acidified) in order to 

create an acidic environment for the acidophilic microorganisms. Also, if inorganic acid 

is used, the ore can be inoculated with the leaching microorganisms before stacking to 

ensure an even distribution of microorganisms throughout the heap right from the 

beginning of the leaching process. The heaps are irrigated from the top through a 

network of drip lines or sprinklers with recycled acidic and bioleaching microorganismcontaining 

raffinate solution. The metal-impregnated solution is collected from the 

bottom of the heaps (Figure 4.1) and the metals are recovered. For the recovery of, e.g., 

nickel, zinc and cobalt, hydrogen sulfide can be used for precipitation (Riekkola- 

Vanhanen 2010). In the case of copper, metal recovery is done by solvent extractionelectrowinning 

(SX-EW) (Bartos 2002). The metal-free raffinate is then recycled back 

into the process. (For reviews, see Watling 2006; Rawlings 2002) Inorganic nutrients 

such as (NH4)2SO4 and KH2PO4 are usually added to the raffinate solution before 

irrigation (Rawlings 2002). Aeration pipes can be installed at the bottom of the heap to 

provide the bioleaching microorganisms with more oxygen and CO2 to further enhance 

the leaching process (Riekkola-Vanhanen 2010; Lizama 2001). 

20

Figure 4.1: Schematic of an aerated heap bioleach process. Partly impregnated 

intermediate leach solution (ILS) from the primary (upper) heap is used to 

irrigate the secondary (lower) heap. Pregnant leach solution (PLS) from the 

secondary heap goes through solvent extraction and the metal-free raffinate is 

recycled back into the process. (Watling 2006) 

The basic principle behind the dump process is similar to that of the heap process, 

the difference being that it is much more robust and less controlled and several 

magnitudes larger in scale. In dump leaching, millions of tons of run-of-mine ore can be 

piled up in stacks even as high as 350 m. It is obvious that with such large amounts of 

ore, economical distribution of sufficient amount of air and water would be challenging. 

That is why dump leaching is mostly used when dealing with vast amounts of ore and 

when efficient metal extraction of metals from these ores is not a priority. Thus, only 

the minimal amount of treatment is performed on the ore and the dump. The dump is 

irrigated with recycled raffinate and the pregnant leachate is collected from the bottom 

like in the heap process but aeration, agglomeration, ore curing etc. are not performed in 

order to minimize the costs. This way, the leach cycle in dump processes can take years. 

(For reviews, see Watling 2006; Olson et al. 2003; Rawlings 2002; Brierley 1982) 

In situ leaching takes the minimizing of costs even further as the metals are 

extracted from the ore underground, without haulage (Sand et al. 1993). The geology of 

the site, however, sets limitations to the implementation of in situ processing as the ore 

body has to be sufficiently permeable and the gangue rock sufficiently impermeable for 

efficient metal recovery (Bosecker 1997). For example, at San Manuel near Tucson, 

Texas, copper is leached underground by injecting acidified leaching solution into the 

mineral deposits via injection wells. The solution then percolates through the ore and 

collects to abandoned mine workings or production wells from which the pregnant 

solution is pumped to the surface for metal extraction. (Schnell 1997) Underground in 

situ leaching can make metal extraction economically viable from ores that could not be 

21

leached in an economical way using the conventional methods (Rawlings 2002). 

However, as evidence of microbial involvement in underground leaching is scarce, 

Watling (2006) questions the importance of microorganisms in the process, based on the 

fact that deep underground, oxygen is very limited, and as a result microbial activity 

would be expected to slow down considerably. 

4.3 Stirred-Tank Processes 

The use of highly aerated continuous-flow stirred-tank reactors boosts the efficiency 

and rate of biomining processes. However, construction and operation of such reactors 

is costly compared to the more robust irrigation-type processes described above. Stirredtank 

processes are used mainly for biooxidation of high-value gold-bearing refractory 

ores and concentrates (Langhans et al. 1995) and bioleaching of cobalt (d'Hugues et al. 

1997). 

In biooxidation, the goal of microbially catalyzed oxidative attack is not to leach the 

metal to be extracted, but rather to dissolve the sulfide matrix enclosing the metal and 

expose it for leaching by other means (for a review, see Rawlings 2002). In the case of 

gold, most industrial processes use biooxidation in tank reactors for the pretreatment of 

recalcitrant arsenopyrite concentrates (Deng et al. 2000). The leaching of gold is then 

done with cyanide, although a microbiological substitute has been studied (Feng & van 

Deventer 2006; Groudev et al. 1996). A typical biooxidation process comprises highly 

aerated and vigorously stirred continuous-flow biooxidation tanks arranged in series 

(Figure 4.2), with a continuous feed of the mineral concentrate and nutrients in the first 

Figure 4.2: Flow diagram for continuous stirred-tank reactor process of gold 

recovery from refractory arsenopyrite concentrate. Primary aeration tanks are 

arranged in parallel in order to increase retention time and thus provide microbes 

with enough time to duplicate so that washout does not occur. (Rawlings 2002) 

tank. Tank sizes can range from 1000 to 2000 m 3 (Rawlings 2002). Temperature and pH 

can be adjusted individually for each tank for optimal process yield (Rawlings et al. 

1999b). For example, using the BIOX ® process, the most extensively used commercial 

22

iooxidation method, the mineral decomposition typically only takes some days, 

whereas the reactions in heap reactors can take from months to even years (for reviews, 

see Watling 2006; Rawlings et al. 2003). 

23

5 ARSENIC IN BIOMINING 

5.1 General Characteristics 

Geological distribution 

Arsenic (As, atomic number = 33) is a metalloid found abundantly in the Earth’s upper 

crust. It exists in four oxidation states: +V (arsenate), +III (arsenite), 0 (arsenic), and – 

III (arsine). (Sharma & Sohn 2009) In nature, only arsenite and arsenate oxidation forms 

are clearly distinguishable. In natural waters, arsenic occurs as oxyanions of arsenite 

and arsenate, i.e., AsO3 3- and AsO4 3- , respectively. Solution pH and redox potential 

largely determine the degree of hydration as well as arsenic state of oxidation. Figure 

5.1 presents a redox potential (Eh) – pH diagram for an acid mine drainage-type As-Fe- 

O-H-S system. 

Figure 5.1: Eh–pH diagram for an acid mine drainage-type system (adapted from 

Bednar et al. 2005) 

24

In the solid phase, arsenic is rarely found in its elemental form. Instead, it is often 

found combined with sulfur, e.g., in orpiment (As2S3), realgar (As2S2 or AsS), or 

arsenopyrite (FeAsS). In addition to sulfur, other common combinations are with 

tellurium as in As2Te, and with selenium as in As2Se3. Heavy metal combinations 

include those with iron (loellingite, FeAs2), copper (domeykite, Cu3As), nickel (nicolite, 

NiAs), and cobalt (Co2As). Other mineral forms include arsenolite (or claudetite, 

As2O3), erythrite (Co3(AsO4)2 · 8 H2O, scorodite (FeAsO4 · 2H2O), and olivenite 

(Cu2(AsO4)(OH)). Arsenic concentration in soil ranges from 0.1 to 1000 ppm, and is 

mainly in the form of arsenopyrite. (Ehrlich & Newman 2009) 

Toxicity 

Arsenic is toxic to most living organisms. To humans, arsenic is a carcinogen and, due 

to its vast natural abundance, arsenic is regarded as one of the most important sources of 

chemical poisoning in the world, especially in the developing countries. Tailings from 

metal-mining processes along with the combustion of fossil fuels are the major sources 

of anthropogenic arsenic contamination of the environment (ICPS 2001). In many 

mining areas, the arsenic concentration in nearby waters is in the range of 0.1 … 5.0 

mg/L (Williams 2001). The drinking water standard set by the World Health 

Organization (WHO) is 10 µg/L (WHO 2011). However, there is great variance in the 

toxicity between different oxidation states as well as organic and inorganic arsenic 

compounds. Generally, arsenite and inorganic arsenic are more toxic than arsenate and 

organic arsenic, respectively. (Ng 2005) 

Arsenite has a very high affinity for protein sulfhydryl groups of biomolecules, e.g., 

glutathione (GSH). The formation of As(III)-S bonds inhibits enzymes such as 

glutathione reductase, glutathione peroxidases, thioredoxin reductase, and thioredoxin 

peroxidase. (For a review, see Sharma & Sohn 2009) Arsenate, on the other hand, 

uncouples oxidative phosphorylation thus inhibiting ATP synthesis (Ehrlich & Newman 

2009). 

25

5.2 Microbial Interaction and Resistance 

Even though arsenic is toxic to most organisms, its abundance and natural occurrence 

have made it possible for microorganisms to develop a wide variety of resistance 

mechanisms. Some strains are even able to oxidize or reduce arsenic for metabolic 

purposes. Figure 5.2 illustrates different mechanisms with which microorganisms 

interact with arsenic. 

Figure 5.2: Schematic representation of the different processes evolved by 

prokaryotes to cope with arsenic: (1) arsenic (As) enters the cell through the 

phosphate transporters [as arsenate, As(V)] or through the aquaglyceroporin [as 

arsenite, As(III)]; (2) arsenic is immobilized in the environment by extracellular 

precipitation; (3) As(V) inside the cell is reduced to As(III) by an arsenate 

reductase (ArsC). As(III) inside the cell is pumped out of the cell through a 

membrane protein [Ars(A)B]; (4) inorganic arsenic can be transformed into 

volatile or non-volatile organic species via a series of methylation steps; (5) As(V) 

is used as an electron acceptor during respiration by the dissimilatory arsenate 

reductase ArrAB; (6) As(III) can serve as an electron donor via the As(III) 

oxidase AoxAB or ArxAB. Pit, phosphate (PO4 3- ) inorganic transport system; Pst, 

phosphate specific transport system; GlpF, aquaglyceroporin GlpF; ArsC, 

detoxifying arsenate reductase; Ars(A)B, arsenite efflux pump; MMA, 

monomethylarsonic acid; DMAA, dimethylarsinic acid; DMA, dimethylarsine; 

TMAO, trimethylarsine oxide; Aox/Arx, different arsenite oxidase types; Arr, 

dissimilatory arsenate reductase. (Slyemi & Bonnefoy 2011) 

26

At pH < 9.3 arsenite exists in the form As(OH)3, which is similar to glycerol. Hence 

the uptake of arsenite proceeds mainly via GlpF, the transport system for glycerol. 

(Meng et al. 2004) Arsenate, on the other hand, is analogous to phosphate and thus 

enters the cell through phosphate transport systems Pit and Pst (Rosenberg et al. 1977). 

Active arsenite extrusion via the Ars system, encoded in the ars operon, is the most 

extensively studied arsenic resistance mechanism in prokaryotes. The ars operon can 

exist either on a plasmid or on a chromosome. ArsB is an integral membrane protein 

that pumps arsenite out of the cell. When present, ArsA functions as an ATPase that 

provides energy to ArsB from the hydrolysis of ATP, making the extrusion of arsenite 

more efficient. No arsenate extrusion mechanism has been reported to date. Instead, 

ArsC, a cytoplasmic arsenate reductase, reduces arsenate to arsenite which is then 

pumped out. (For a review, see Slyemi & Bonnefoy 2011) 

Arsenic methylation is a widespread arsenic resistance mechanism. Methylation is a 

multistep process that starts with the reduction of arsenate to arsenite and proceeds with 

a step-wise introduction of methyl groups to arsenite. (For a review, see Bentley & 

Chasteen 2002). Although arsenic methylation by fungi and other eukaryotes has been 

extensively studied, the respective mechanisms for prokaryotes are still somewhat 

unclear. The pathway proposition by Challenger (1945), based on studies on the fungus 

Scopulariopsis brevicaulis, is still widely in use (Páez-Espino et al. 2009; Dombrowski 

et al. 2005; Bentley & Chasteen 2002) The main prokaryotic methylation products 

consist of volatile arsenic compounds such as mono-, di- and trimethyl arsine (MMA, 

DMA and TMA) (Bentley & Chasteen 2002). 

Arsenate reduction usually occurs in one of two contexts. Arsenate can undergo 

detoxifying reduction by the arsenate reductase ArsC as mentioned above, but some 

microorganisms such as the bacteria Sulfurospirillum barnesii and S. arsenophilum 

(Stolz et al. 1999) can also use arsenate as a terminal electron acceptor in anaerobic 

respiration. This dissimilatory reduction is performed by a heterodimeric terminal 

reductase ArrAB complex, where arsenate acts as an electron acceptor of the microbial 

electron transfer chain. ArrAB consists of two subunits, ArrA (~100 kDa) and the 

smaller ArrB (~30kDa). ArrB functions as an electron conductor from the respiratory 

chain to ArrA where the actual arsenate reduction takes place. (For a review, see Slyemi 

& Bonnefoy 2011) The respiratory arsenate reductases of only three species, i.e., 

Chrysiogenes arsenates (Krafft & Macy 1998), Bacillus selenitireducens (Afkar et al. 

2003) and Shewanella trabarsenatis (Malasarn et al. 2008) have been characterized. 

Arsenate reduction is a cause of environmental problems, as arsenite is much more 

mobile and toxic than arsenate. The reverse, arsenite oxidation, on the other hand, is an 

important mechanism for immobilization of arsenic and a widely reported property of 

various microorganisms. (For reviews, see Slyemi & Bonnefoy 2011; Ehrlich & 

Newman 2009; Páez-Espino et al. 2009) Arsenite oxidation can take place aerobically 

or anaerobically and, unlike what was believed earlier, not all arsenic-oxidizing 

microorganisms oxidize arsenite only as a means of detoxification; some can actually 

draw energy from the reaction (Ehrlich & Newman 2009). These microorganisms 

include such species as Alkalilimnicola ehrlichii (Hoeft et al. 2007) or the strain 

27

elonging to the Agrobacterium/Rhizobium branch of the α-Proteobacteria found from 

a gold mine in Australia (Santini et al. 2000). A widely recognized enzyme responsible 

for the oxidation of arsenite in most arsenite oxidizing microorganisms is the arsenite 

oxidase AoxAB (also known as AsoBA or AroBA), consisting of two subunits: the 

larger AoxB (~90 kDa) and the smaller AoxA (~14 kDa) (Slyemi & Bonnefoy 2011). 

The electrons are transferred from the large subunit to the smaller and on to the 

respiratory chain, namely periplasmic azurin and cytochrome c in the case of 

Alcaligenes faecalis (Ellis et al. 2001). Recently, another arsenite oxidase, ArxAB, has 

been reported that is phylogenetically much closer to the dissimilatory arsenate 

reductase ArrAB that to AoxAB. However, ArxAB has been detected so far only in the 

family Ectothiorhodospiraceae. (Slyemi & Bonnefoy 2011) 

5.3 Removal of Arsenic from Mining Waters 

The microbial oxidation of arsenic-bearing minerals releases considerable amounts of 

arsenic into the bioleaching solution. Preventing mine tailings and with them dissolved 

arsenic from reaching the environment is of utmost importance. (ICPS 2001) Existing 

water treatment technologies for arsenic removal are many. They employ mechanisms 

such as ion exchange, sorption/(co-)precipitation, coagulation, and different membranebased 

techniques such as ultra-, micro-, and nanofiltration or reverse osmosis (for an 

extensive overview on commonly used methods for arsenic removal from drinking 

water, see (EPA 2000)). However, mining related wastewaters generally exhibit 

extreme and challenging characteristics in terms of wastewater treatment. The biggest 

challenges are caused by a very acidic pH (< 3) and high concentrations of sulfate and 

various soluble metals, especially iron. (King 1995) Such demanding conditions make 

many traditional wastewater treatment methods too challenging or expensive to use 

(EPA 2000). 

Arsenic removal methods utilizing various iron compounds are widely used and 

studied (Tresintsi et al. 2012; Zeng et al. 2008; Nikolaidis et al. 2003; Zouboulis et al. 

2002; Joshi & Chauduri 1996) Thus, given the usually high soluble iron content in 

mining waters, it is not surprising that the majority of studies on arsenic removal from 

mining waters concentrate on the ability of various iron oxides, including 

oxyhydroxides and hydroxides, to efficiently remove soluble arsenic (Dong et al. 2011; 

Doušová et al. 2005; Kim & Davis 2003; Carlson et al. 2002). The high iron content of 

mining waters specifically presents for the mining industry a more lucrative and costeffective 

option than the use of other compounds or methods for arsenic control (EPA 

2000). Currently arsenic is removed from wastewater streams of the mineral processing 

industry by lime neutralization with co-precipitation of arsenic with iron oxides (Esper 

et al. 2012; Twidwell & McCloskey 2011). 

Apart from the more traditional and engineered approach for arsenic removal using 

iron oxides, constructed wetlands are used for a more black-box type of treatment 

method. Constructed wetlands are widely utilized for immobilization or transformation 

of heavy metal and sulfate (Esper et al. 2012; Nyquist & Greger 2009) as well as 

28

manganese (Hallberg & Johnson 2005) in acid mine drainage. However, literature on 

the utilization of constructed wetlands for arsenic removal specifically is scarce, save a 

recent attempt by Lizama et al. (2011) to spotlight the topic. Despite the lack of 

published in-depth studies, constructed wetlands have been shown to efficiently remove 

arsenic from wastewater (Ye et al. 2003) and are also used at mine sites for arsenic 

retention (Esper et al. 2012). As the mechanisms of arsenic removal are as of yet largely 

unstudied, Lizama et al. (2011) drew the conclusion that the dominant mechanisms are 

(co-)precipitation and sorption, much like in the use of iron oxides for arsenic removal 

as described below. The hypothesis was supported by the findings of Ye et al. (2003) 

stating that immobilization in the sediment accounted for the most of the arsenic 

removal by a constructed wetland for coal gasification plant wastewater treatment. As 

constructed wetlands for arsenic removal are such a poorly covered subject in the 

literature and irrelevant in the light of the present study, this chapter will only 

concentrate on arsenic removal based on the use of iron oxides. 

5.3.1 Mechanisms of Arsenic Immobilization with Iron Oxides 

The mechanisms behind arsenic removal with iron oxides are predominantly 

sorption/ion exchange and (co-)precipitation (De Klerk et al. 2012; Langmuir et al. 

2006; Zhang & Itoh 2005). Pre-formed iron oxides can be added to the process for 

arsenic removal, e.g., as a packed bed of iron oxide coated sand through which the 

contaminated solution is passed (Thirunavukkarasu et al. 2003). Using pre-formed iron 

oxides, the removal mechanism is mainly sorption/ion exchange (Zeng et al. 2008; Gu 

et al. 2005; Zhang & Itoh 2005). Alternatively, arsenic removal with iron oxides can be 

performed in situ by precipitating iron as oxides from the contaminated solution itself, 

provided that the water contains soluble iron. In simultaneous arsenic and iron removal 

arsenic removal proceeds via both sorption/ion exchange and (co-)precipitation. (De 

Klerk et al. 2012) Arsenic can precipitate with iron as amorphous ferric arsenate 

(FeAsO4), crystalline scorodite (FeAsO4 · 2 H2O) (Langmuir et al. 2006) or as an 

amorphous phase of variable composition such as pitticite (Fe(III) – SO4 – As2O5 – 

H2O) (Dunn 1982). For example, lime softening, a routine method in water treatment, 

has been shown to effectively remove soluble iron and arsenic as various precipitates 

due to increase in pH. The precipitates then further enhance the recovery of soluble 

arsenic from the solution through sorption. (EPA 2000) 

29

5.3.2 Common Mining-related Iron Oxides 

Widely studied naturally occurring iron (oxyhydr)oxides that all have a strong affinity 

for arsenic include amorphous hydrous iron oxide (FeOOH), goethite (α-FeOOH), and 

hematite (α-Fe2O3) along with more complex iron oxide based compounds such as 

schwertmannite (Fe8O8(OH)5.5(SO4)1.25) or jarosite ([K, Na, NH4]Fe3(SO4)2(OH)6), 

(Asta et al. 2009; Mamindy-Pajany et al. 2009; Gräfe et al. 2008; Zhang et al. 2004; 

Carlson et al. 2002; Jain et al. 1999). Natural iron oxides from mine drainage, e.g., 

magnetite (Fe 2+ Fe 3+ 2O4), ferrihydrite (Fe2O3 · 0.5 H2O) and goethite have even been 

proposed as low-cost substitutes for some expensive high-grade adsorbent and catalyst 

compounds (Flores et al. 2012). 

Schwertmannite, (K-)jarosite and goethite are readily formed in the mining 

environment (Asta et al. 2009) and have been studied extensively for arsenic adsorption 

(Paikaray et al. 2011; Tang et al. 2010; Savage et al. 2005). Schwertmannite is 

amorphous and rather unstable and it can transform to the more crystalline jarosite or 

goethite. The crystallization can lead to the release of sorbed arsenic into the solution. 

(Courtin-Nomade et al. 2003) Jarosite and goethite, on the other hand, are stable 

minerals. However, there remains ambiguity concerning the relative arsenic retention 

capabilities between the two minerals, especially in acidic conditions. (Asta et al. 2009) 

The arsenic sorption mechanisms on goethite and jarosite are thought to be different. 

According to Gräfe et al. (2008), the difference in sorption mechanisms is most clearly 

witnessed in the ability of the structural sulfate groups in jarosite to be replaced by and 

competed with arsenate, which does not occur with goethite. 

5.3.3 Factors Affecting Arsenic Immobilization with Iron Oxides 

Understanding arsenic chemistry is vital for successful arsenic removal. The arsenic 

oxidation state is an important factor as arsenate generally exhibits better sorptive 

abilities than arsenite and is considerably less toxic to most organisms. In many studies, 

arsenate has been shown to adsorb better than arsenite on various adsorbents (Zouboulis 

et al. 2002; Lin & Wu 2001; Suzuki et al. 2000), including schwertmannite, jarosite and 

goethite that are readily formed in mining water environments (Asta et al. 2009). Figure 

5.1 shows that under very oxidizing conditions (Eh ≥ 600 mV), arsenate is the dominant 

form of arsenic virtually regardless of pH. Generally such high redox potentials are rare 

but a typical biomining environment is characterized by a high microbially generated 

ferric to ferrous iron ratio which generates a high redox potential (Bednar et al. 2005). 

Sedelnikova et al. (1999) determined the minimum Fe:As molar ratio for stable 

iron-arsenic precipitate formation to be 4:1 with the precipitate stability increasing with 

an increasing Fe:As ratio. Thus, a sufficient amount of ferric iron must be present in the 

solution to retain as much arsenic as possible. 

The level of protonation caused by solution pH is important as it directly affects the 

electrical charge of the arsenic species (e.g., H2AsO3 - and H3AsO3) as well as the 

surface charge of the sorbent. Thus pH affects the efficiency of removal methods based 

30

on ionic bonding. (Stollenwerk 2003) As iron precipitation is also pH dependent 

(Watling 2006), the challenge in arsenic control in biomining processes boils down to 

the appropriate selection and control of process pH. 

Asta et al. (2009) noted that arsenic sorption on jarosite was much less pH 

dependent compared to that on goethite. They attributed that to the finding by Gräfe et 

al. (2008) that sulfate groups are replaced by arsenate in the jarosite sorption 

mechanism. This exchange mechanism makes arsenic sorption on jarosite sensitive to 

sulfate concentration as stated by Asta et al. (2008): an addition of sulfate to the 

solution only in the mmol/L range was enough to diminish arsenate adsorption on 

jarosite from 38 % to less than 5 %. Arsenate adsorption on goethite was much less 

affected by sulfate concentration, namely 25 % with 20 mmol/L sulfate compared to the 

1 % adsorption by jarosite at the same sulfate concentration. 

31

6 MATERIALS AND METHODS 

6.1 Experimental Outline 

This thesis comprises three individual experiments. Based on previous shake-flask 

studies by Wakeman et al. (2011), a 30 L batch experiment was conducted to assess the 

simultaneous bioleaching amenability of nickel flotation concentrate (NFC) and 

immobilization of arsenic at constant, elevated pH of 3.0. As a follow-up to the 30 L 

batch experiment, a 20 L semi-continuous mode experiment was conducted in semicontinuous 

mode at pH 3.0 to assess parameters required for developing a continuous 

process. Finally, a shake-flask experiment without NFC was conducted in order to 

assess the role of bacteria in arsenic speciation and immobilization in the presence of 

iron, arsenic and sulfate. 

6.2 Materials 

6.2.1 Nickel Flotation Concentrate (NFC) 

The nickel flotation concentrate studied in this thesis was provided by Mondo Minerals 

LLC. NFC is a flotation side-product of talc production yet relatively rich in nickel and 

cobalt. The elemental and mineral compositions of NFC are presented in Table 6.1 and 

Table 6.2, respectively. The elemental distribution was provided by Mondo Minerals 

LLC and the mineralogical composition by the Geological Survey of Finland (GTK). 

Table 6.1: Elemental distribution of nickel flotation concentrate (courtesy of Mondo 

Minerals LLC) 

Element 

Fe 43.86 

Ni 9.19 

Co 0.40 

As 1.46 

Mg 2.07 

C n.d. 

S n.d. 

n.d.: not determined 

Nickel flotation 

concentrate (%) 

32

Table 6.2: Mineralogy of nickel flotation concentrate (courtesy of GTK) 

Mineral Chemical formula 

Pyrrhotite Fe(1-x)S (x = 0 ... 0.2) 59.5 

Pentlandite (Fe, Ni)9S8 22.1 

Talc Mg3Si4O10(OH)2 5.5 

Pyrite FeS2 4.8 

Magnesite MgCO3 4.3 

Gersdorffite NiAsS 2.2 

Dolomite (CaMg)(CO3)2 0.7 

Chlorite (Mg,Fe)3(Si,Al)4O10(OH)2·(Mg,Fe)3(OH)6 0.4 

Chalcopyrite CuFeS2 0.1 

Chromite (Fe, Mg)Cr2O4 0.1 

Unclassified 0.1 

6.2.2 Bioleaching Cultures and Growth Media 

Nickel flotation 

concentrate (%) 

The bioleaching cultures used in the experiments were subcultures of enrichments from 

a tailings pond in the Mondo Minerals LLC mine site as described in Wakeman et al. 

(2011). The subcultures were incubated in shake-flasks of various volumes at 150 rpm 

and 27 °C. The microbial growth was supported with 1 % (w/v) NFC, 10 % (v/v) 

Mineral Salt Medium (MSM), 1 % (v/v) Trace Element Solution (TES), and 79 % (v/v) 

de-ionized (DI) H2O. Inoculation of new subcultures was performed by adding a 10 % 

(v/v) inoculant sample from a previous subculture. 

Both MSM and TES were prepared in DI H2O. MSM composition was as follows: 

37.5 g/L (NH4)2SO4, 18.75 g/L Na2SO4 · 10 H2O, 1.25 g/L KCl, 0.625 g/L K2HPO4, 

6.25 MgSO4 · 7 H2O, 0.175 g/L Ca(NO3)2 · 4 H2O. The MSM stock solution pH was 

adjusted to 1.8 with concentrated H2SO4. TES composition was as follows: 1.375 g/L 

FeCl3 · 6 H2O, 0.0625 g/L CuSO4 · 5 H2O, 0.25 g/L H3BO3, 0.319 g/L MnSO4 · 4 H2O, 

0.1 g/L Na2MoO4 · 2 H2O, 0.075 g/L CoCl2 · 6 H2O, 0.1125 g/L ZnSO4 · 7 H2O, 0.1125 

g/L Na2SeO4. The TES components were dissolved in DI H2O adjusted to pH 1.5 with 

concentrated H2SO4. All chemicals used were reagent grade and provided either by 

Merck (Germany), J.T. Baker (the Netherlands) or Sigma-Aldrich (Germany). 

6.2.3 Analytical Solutions 

For ferrous iron analysis, ammonium acetate buffer was prepared by dissolving 250 g 

NH4C2H3O2 in 180 mL DI H2O followed by addition of 700 mL glacial acetic acid 

(C2H3O4). The phenanthroline solution was prepared by dissolving 5 g 1,10phenanthroline 

monohydrate (C12H8N2 · H2O) in 450 mL DI H2O. A few drops of 

concentrated HCl were added to enhance the dissolution. After complete dissolution the 

solution was diluted to 500 mL. All solutions were prepared in acid washed containers. 

33

The eluent for ion chromatography was 0.5 M NaHCO3 and 0.5 M Na2CO3 in DI 

H2O. Both NaHCO3 and Na2CO3 stock solutions were filtered to 0.45 µm through GF/A 

glass microfiber filters (Whatman, UK). 

The EDTA wash solution for washing of atomic absorption spectrometry (AAS) 

standard flasks was 1 % (w/v) EDTA (C10H14N2Na2O8 · 2 H2O) and 5 % (v/v) Suma 

Nova Pur-Eco L6 machine dishwashing detergent (Johnson Diversey, USA) in DI H2O. 

TRIS buffer was prepared by dissolving 6.1 g C4H11NO3 in 500 mL DI H2O and 

adjusting the pH to 8.0 with concentrated HCl. 

6.3 Experimental Methods and Designs 

6.3.1 30 L Batch Experiment 

NFC was bioleached in three 30 L reactors with pulp densities of 15, 10, and 5 % (w/v) 

at room temperature (~21 °C) for 94, 58, and 79 days, respectively. The 10 % p.d. 

reactor was shut down prematurely due to sludge overflow resulting from a failure of a 

metallic air tube connector. 

The reactors contained initially 3 L MSM, 300 mL TES, 3 L inoculant, 23.7 L DI 

H2O and 4.5, 3.0, and 1.5 kg NFC corresponding to pulp densities 15, 10, and 5 % 

(w/v), respectively. The reactors were lidded cylindrical 45 L polypropylene (PP) tanks 

(P/N 365, Orthex, Finland). The lids were not tightly sealed. The reactors were mixed at 

250 rpm with RZR 2052 overhead mixers fitted with C-shaped Teflon impellers 

(Heidolph, Germany). The reactors were aerated from the bottom with air at 8 L/min 

through two ToppTube TM PA12 polyamide (PA) tubes (Toppi, Finland) placed on 

opposite sides of the reactor (Figure 6.1). Air flow was adjusted with NP-G24 flow 

meters (Kytola Instruments, Finland). 

Figure 6.1: Reactor schematic for the 30 L batch and 20 L semi-continuous mode 

experiments. 

In the beginning of the experiment, solution pH was maintained at 3.0 by daily 

manual adjustment with 2 M H2SO4 and 5 M NaOH. After 21, 15, and 6 days of 

34

operation of 15, 10, and 5 % p.d. reactors, respectively, constant pH of 3.0 was 

maintained by Titrino 719 S automatic titrators (Metrohm, Switzerland) with 5 M 

NaOH until the end of the experiment. Water evaporation was compensated for by 

measuring the solution level at regular intervals and adding an appropriate amount of DI 

H2O. 

The reactors were sampled weekly for total iron, nickel, cobalt, and arsenic and 

sulfate. Ferrous iron concentration, redox potential and pH were measured daily. 

Samples for leach residue analysis were collected after 27, 37, 69, 77, 80, 86, and 90 

days of operation from the 15 % p.d. reactor, after 21 and 31 days of operation from the 

10 % p.d. reactor, and after 12, 22, 54, 62, 65, 71, and 75 days of operation from the 5 

% p.d. reactor. 

6.3.2 20 L Semi-continuous Mode Experiment 

The reactor configuration of the 20 L semi-continuous mode experiment was similar to 

that of the 30 L batch experiment. Only one reactor was operated with NFC pulp density 

of 5 % (w/v) at room temperature (~21 °C) and pH 3.0. A Titrino 719 S automatic 

titrator (Metrohm, Switzerland) was used for pH control throughout the experiment. The 

reactor was operated for 38 days with a retention time of 10 d, i.e., 2 L of solution from 

the reactor was replaced with 2 L of fresh solution [10 % (v/v) MSM and 1 % (v/v) TES 

in DI H2O] on a daily basis. No mass exchange was performed on weekends. Instead, a 

4 L exchange was performed on Fridays and Mondays. The reactor was sampled for 

total iron, nickel, cobalt, and arsenic prior to a mass exchange. Redox potential and pH 

were measured daily. 

6.3.3 Shake-flask Experiment for the Assessment of the Role of Bacteria 

in Arsenic Speciation and Immobilization 

Shake-flasks with a total solution volume of 100 mL were incubated at 150 rpm and 27 

°C for 28 days. Two starting pH values were used, i.e., pH 1.7 and 2.5. Two parallels 

and one abiotic control were prepared per each starting pH. No pH control was carried 

out during the experiment. The starting pH was adjusted prior to inoculation, or to the 

addition of DI H2O into the abiotic controls. The solutions consisted of 10 % (v/v) 

MSM, 1 % (v/v) TES, 25.1 g/L FeSO4 · 7 H2O, 20 mg/L As(III) as As2O3 in 2 % nitric 

acid (Arsenic Standard for AAS, Sigma-Aldrich, USA) and 10 % (v/v) inoculant. To 

remove any sorbed metals or arsenic from the cell surfaces the inoculant was washed 

with DI H2O at pH 2.0 (adjusted with concentrated H2SO4) and centrifuged at 5,000x g 

for 10 min at 20 °C. The inoculant was replaced by a similar amount of DI H2O at pH 

2.0 (adjusted with concentrated H2SO4) in the abiotic controls. Samples were collected 

weekly for pH and redox potential measurements and for ferrous iron, sulfate, and 

arsenic speciation analyses. 

35

6.4 Analytical Methods 

6.4.1 Redox Potential and pH 

Redox potential was measured with a BlueLine 31 Rx redox potential electrode (Schott 

Instruments, Germany) connected to a pH 340 multimeter (WTW, Germany). Shakeflask 

sample pH was measured with a Sentix 41 pH electrode (WTW, Germany) 

connected to a pH 330 pH meter (WTW, Germany). Reactor pH was continuously 

measured by a Titrino 719 S automatic titrator (Metrohm, Switzerland). 

6.4.2 Ferrous Iron 

Ferrous iron was measured using the ortho-phenanthroline method according to the 

standard 3500-Fe (APHA 1992). Samples were filtered (0.45 µm) using 

polysulfonemembrane filters (Whatman, UK) and diluted with 0.07 M HNO3 to fit the 

external standard curve range (0.05…20 mg/L ferrous iron). Absorbance was measured 

at 510 nm with a UV 1601 spectrofotometer (Shimadzu, Japan). 

6.4.3 Total Iron, Nickel, Cobalt and Arsenic 

Total concentrations of iron, nickel, cobalt and arsenic were measured with an atomic 

absorption spectrometer (AAS; AAnalyst 400, Perkin Elmer, USA). Iron, nickel, and 

cobalt were measured using a multielement hollow cathode lamp (Co-Cr-Cu-Fe-Mn-Ni, 

Perkin Elmer, Singapore). Arsenic was measured with an electrodeless discharge lamp 

(Perkin Elmer, Shelton CT, USA). Analyses were conducted according to the standards 

SFS 3044 (1980) and SFS 3047 (1980). All samples were filtered to 0.45 µm using 

polysulfonemembrane filters (Whatman, UK) and diluted with 0.07 M HNO3 to fit the 

external standard range. External standard curves were prepared from 1000 ± 4 mg/L 

stock solutions for AAS (Sigma-Aldrich, USA). Standard ranges for different analytes 

were as follows: 0.5…3.0 mg/L Fe, 1.0…4.0 mg/L Ni, 0.2…1.0 Co, and 1.0…20.0 

mg/L As. 

6.4.4 Arsenic Speciation 

Arsenic speciation samples were analyzed by the Finnish Institution of Occupational 

Health (TTL) by liquid chromatography with hydride generation – atomic fluorescence 

spectrometry (LC-HG-AFS). The analysis method used by TTL was accredited for the 

analysis of organic and inorganic arsenic compounds from urine. Samples were 

prepared by filtration to 0.45 µm using polysulfomembrane filters (Whatman, UK) and 

dilution with 0.07 M by a dilution factor of 10. 

36

6.4.5 Sulfate 

Sulfate concentration was measured by ion chromatography (IC). The ion 

chromatograph DX-120 (Dionex, USA) was equipped with an IonPac AG23 (4 x 50 

mm) guard column, an IonPac AS23 (4 x 250 mm) and an AS40 autosampler (all: 

Dionex, USA). Samples were prepared according to the standard SFS-EN ISO 10304-2 

(1997). Samples were filtered to 0.45 µm with polysulfomembrane filters (Whatman, 

UK) and diluted with DI H2O to fit the external standard range (10…250 mg/L sulfate). 

The eluent solution (0.5 M NaHCO3, 0.5 M Na2CO3 in DI H2O, filtered to 0.45 µm) was 

degassed prior to the analysis with helium and kept under helium pressure during the 

analysis in order to prevent dissolution of gasses into the eluent. 

6.4.6 Leach Residue Mineral Composition 

Mineral composition of the leach residue samples was analyzed by the Geological 

Survey of Finland (GTK) with X-ray diffraction (XRD). The collected samples were 

prepared by combining two parallel 50 mL samples by separating the solid material 

from the liquid phase by centrifugation (7,000x g, 20 °C for 10 min). 

6.4.7 Bacterial Communities 

Bacterial community analysis was done via polymerase chain reaction coupled with 

denaturing gradient gel electrophoresis (PCR-DGGE) followed by partial sequencing of 

the 16S rRNA gene. 

DNA samples were prepared by suction filtering 20 mL of leach solution through a 

polysulfone membrane filter (0.2 µm; Whatman, UK). The cells remained on the filter 

were washed by first passing 15 mL DI H2O (acidified to pH 2.0 with concentrated 

H2SO4 to prevent premature lysis of the cells of acidophilic bacteria) and then 15 mL 

TRIS-buffer at pH 8.0. Filters were stored at -20 °C for DNA extraction. 

DNA extraction was performed using the Soil DNA Extraction Kit (MoBio 

Laboratories, USA) according to the manufacturer’s instructions for maximum yield. 

The partial 16S rRNA gene in the extracted DNA was amplified by PCR. The primers 

for 16S rRNA amplification were Ba357F-GC [5’-CGC CCG CCG CGC GCG GCG 

GGC GGG GCG GGG GCA CGG GGG GCC TAC GGG AGG CAG CAG-3’; Muyzer 

et al. (1993)] and 907R [5’- CCG TCA ATT CMT TTG AGT TT -3’; Lane (1991)]. 

The PCR mix prepared in nuclease-free H2O contained 1x DreamTaq TM buffer (Thermo 

Fisher Scientific), 0.8 U DreamTaq TM DNA polymerase (Thermo Fisher Scientific), 100 

µM dNTP, 500 nM primers, and 400 ng/µL bovine serum albumin (BSA). PCR was 

performed with a T 3000 Thermocycler (Biometra, Germany) using the following 

program: 95 °C for 5 min, 31 cycles of 95.0 °C for 30 s, 54.6 °C for 30 s, and 72.0 °C 

for 1 min 30 s followed by final extension at 72.0 °C 10 min. 

DGGE was performed with an INGENYphorU 2 x 2 –system (Ingeny, the 

Netherlands) using 8 % polyacrylamide gel with a denaturing gradient from 30 to 70 % 

37

(Auvinen et al. 2009). Gel was run at 60 °C in 1 x TAE with 100 V for 22 h and stained 

with SYBR ® Gold (Molecular Probes, Inc., USA). Visible bands were cut from the gel 

with sterile surgical blades and eluted overnight in sterile H2O. A partial 16S rRNA 

gene was re-amplified for sequencing with PCR as described above except BSA was not 

added and there was no GC clamp in the 5’ end of the forward primer. Amplified 

samples were sent to Macrogen, Inc. (South Korea), for sequencing. Sequence data were 

analysed with the trace data editor, Trev (version 1.9-r2852M), of the Staden Package 

(http://staden.sourceforge.net/) and compared with sequences in GenBank 

(http://www.ncbi.nlm.nih.gov/blast/). 

38

7 RESULTS 

7.1 30 L Batch Experiment 

7.1.1 Bioleaching Results 

Nickel and cobalt 

NFC was bioleached in batch mode using different pulp densities. Nickel and cobalt 

dissolution was as presented in Figure 7.1 A and B, respectively. Maximum nickel 

concentrations of 10,070; 6060, and 3480 mg/L were reached at pulp densities of 15, 10, 

and 5 %, respectively. The concentrations correspond to yield percentages of 73, 66, 

and 76 %, respectively. Maximum cobalt concentrations of 470, 280, and 220 mg/L 

were reached at pulp densities of 15, 10, and 5 % p.d., respectively. The concentrations 

correspond to yield percentages of 78, 71, and 100 %, respectively. Maximum leaching 

rates of both nickel and cobalt were reached at pulp density of 15 %, i.e., 250 mg-Ni/L/d 

and 18 mg-Co/L/d. 

39

Figure 7.1: Nickel (A) and cobalt (B) concentrations and final yields in the 30 L 

batch experiment with 5, 10, 15 % pulp densities (p.d.) of nickel flotation 

concentrate. 

40

Arsenic 

Arsenic concentrations during bioleaching of NFC were as shown in Figure 7.2. After 

the activation of the microbial consortium and resulting rise in redox potential, the level 

of soluble arsenic remained very low: Overall, the amount of arsenic remained below 6 

mg/L and 0.4 % after 21 days of operation at all pulp densities studied. Occasionally the 

amount of arsenic was below the detection limit (1 mg/L) of the atomic absorption 

spectrometer. 

Figure 7.2: Arsenic concentration during the 30 L batch experiment with 5, 10, 

15 % pulp densities (p.d.) of nickel flotation concentrate. 

Leach residue mineralogy 

Table 7.1 presents the proportions of main constituents of leach residue mineral 

composition of bioleached NFC. Table 7.1 shows that not all of the nickel-bearing 

pentlandite was solubilized by the end of the experiment. Iron sulfate formation was 

dominant at all pulp densities. The amount of goethite (α-FeOOH) increased with 

increasing pulp density. 

41

Others 3.8 2.2 2.0 1.4 1.4 1.5 0.8 0.9 2.3 3.3 2.4 5.1 1.6 1.2 1.5 0.9 0.5 

Talc Mg3Si4O10(OH) 2 5.5 1.0 0.7 0.6 0.6 1.0 0.9 0.9 0.9 1.1 0.5 0.9 1.1 1.3 0.7 0.7 0.9 

Magnesite MgCO3 4.3 2.0 1.5 1.4 1.7 2.1 1.2 1.1 1.6 2.2 1.9 3.0 2.6 3.7 2.9 1.4 2.5 

Pyrite 4.8 3.5 3.0 1.6 2.4 4.2 2.4 2.5 2.9 3.8 2.9 4.0 3.1 5.6 5.8 2.6 2.6 

Fe2S Pyrite, secondary 0.0 4.8 2.6 2.0 3.2 2.7 2.0 3.0 3.0 3.4 2.3 2.8 2.0 1.8 1.2 0.9 1.1 

Goethite α-FeO(OH) 0.0 2.6 8.9 3.9 2.8 5.7 5.5 4.8 5.4 7.7 10.5 3.2 6.4 10.7 12.1 12.1 10.5 

Fe-sulfate 0.0 53.5 58.6 65.3 64.3 62.6 70.2 68.6 61.0 53.5 57.9 55.7 57.6 49.6 57.0 66.5 66.7 

Fe-sulfate(Ni) 0.0 1.5 2.3 1.8 1.6 1.6 1.3 1.2 1.7 2.2 1.3 1.9 2.8 3.7 2.2 1.9 1.9 

5 % p.d. 10 % p.d. 15 % p.d. 

Day 

0 12 22 54 62 65 71 75 21 31 27 37 69 77 80 86 90 

Mineral 

w/w % 

Pentlandite (Fe, Ni) 9S8 22.1 12.0 10.4 5.5 5.3 7.6 4.9 4.1 15.0 14.8 12.0 15.6 9.9 12.1 9.6 6.9 6.7 

Pyrrhotite 59.5 6.8 2.0 1.4 1.2 1.0 0.5 0.4 0.8 0.7 2.0 0.9 2.0 1.1 0.8 0.4 0.3 

Fe (1-x)S (x = 0 ... 0.2) 

Pyrrhotite, oxidized 0.0 10.1 7.9 15.1 15.5 10.0 10.3 12.6 5.6 7.4 6.3 6.8 11.1 9.3 6.3 5.7 6.2 

Table 7.1 : Mineral composition of leach residues during the 30 L batch experiment with 5, 10, 15 % pulp densities (p.d.) of nickel flotation 

concentrate. Values of special interest are highlighted in bold. 

42

7.1.2 Operational Parameters 

Redox potential 

Redox potential during bioleaching of NFC in batch mode is presented in Figure 7.3. 

Redox potential remained around +600 mV after the initial lag-phase at all pulp 

densities. The duration of the initial lag-phase was related to pulp density. As a result of 

the sludge overflow and shutdown of the 5 % p.d. reactor due to air tube connector 

failure, the air tube connectors of the 15 and 5 % p.d. reactors were replaced, 

corresponding to 64 and 49 days of operation, respectively. The maintenance 

procedures caused a notable release of unleached sediment from the bottom of the 15 % 

p.d. reactor, which was seen as a drop in redox potential after 62 days of operation. 

Figure 7.3: Redox potential (Eh) during the 30 L batch experiment with 5, 10, 15 

% pulp densities (p.d.) of nickel flotation concentrate. 

43

pH 

Figure 7.4 illustrates pH of the batch NFC bioleaching experiments. Process pH was 

maintained accurately at 3.0 with automated titrators after 16, 10, and 5 days of 

operation with 15, 10, and 5 % p.d., respectively. The initial high variation in pH was 

caused by acid consumption of gangue minerals and the manual adjustment of pH 

performed only once a day. A slight peak in the pH curve of the 15 % p.d. after 62 days 

of operation was caused by the release of unleached material from the reactor bottom as 

described above. 

Figure 7.4: pH level during the 30 L batch experiment with 5, 10, 15 % pulp 

densities (p.d.) of nickel flotation concentrate. 

44

Acid balance 

Acid (2 M H2SO4) and base (5 M NaOH) consumptions during NFC bioleaching was as 

presented in Figure 7.5 A and B, respectively. By the end of the experiment, H2SO4 

consumption reached 20, 7, and 8 g-H2SO4/kg-NFC with 15, 10, and 5 % p.d., 

respectively. NaOH consumption at the end of the experiment was 90, 50, and 20 g- 

NaOH/kg-NFC with 15, 10, and 5 % p.d., respectively. Most of the acid consumption is 

likely attributed to gangue minerals. NaOH consumption corresponds to acid production 

from minerals and thus to the rate of bioleaching. The release of fresh, unleached 

material from the bottom sediment is seen as a sudden increase in NaOH consumption 

in the 15 and 5 % p.d. reactors after 62 and 54 days of operation, respectively. 

Figure 7.5: H2SO4 (A) and NaOH (B) consumption during the 30 L batch 

experiment with 5, 10, 15 % pulp densities (p.d.) of nickel flotation concentrate. 

45

Table 7.3 presents maximum molar acid consumption and acid production as well as 

their ratio during bioleaching of NFC in batch mode. The maximum acid production 

rate versus the maximum acid consumption rate increased with pulp density. 

consumption production production / 

Pulp density mmol H max. acid consumption 

+ Table 7.3 : Acid consumption and acid production properties of 

bioleaching of nickel flotation concentrate (NFC) with 5, 10, 15 % pulp 

densities (p.d.). 

Maximum acid 

/d/kg-NFC 

15 % 115.6 889.8 7.7 

10 % 46.7 196.1 4.2 

5 % 73.3 142.4 1.9 

Total iron and ferrous iron 

Total iron and ferrous iron concentrations during bioleaching of NFC in batch mode 

were as presented in Figure 7.6. Especially the iron concentrations of the 15 % p.d. 

reactor reached high levels (> 10,000 mg/L) during the initial lag-phase. However, the 

ferrous iron levels dropped to around 100 mg/L in each reactor with the onset of 

microbial activity. 

Figure 7.6: Ferrous and total iron concentrations during the 30 L batch 

experiment with 5, 10, 15 % pulp densities (p.d.) of nickel flotation concentrate. 

46

Sulfate 

Sulfate concentrations during batch bioleaching of NFC were as illustrated in Figure 

7.7. With 15 % p.d., the amount of sulfate increased at a rather constant rate from 15 to 

64 g/L over the course of 67 days. With 10 % p.d., the increase in sulfate concentration 

was from 4.9 to 22 g/L in 49 days, and with 5 % p.d. from 6 to 19 g/L in 79 days. The 

sulfate concentrations increased with pulp density. 

Figure 7.7: Sulfate concentration during the 30 L batch experiment with 5, 10, 15 

% pulp densities (p.d.) of nickel flotation concentrate. Error bars represent a 95 

% confidence interval (± 2σ; n = 2). 

47

Figure 7.8 illustrates the sulfate-to-total-arsenic mass ratio in the present study. The 

amount of arsenic was calculated as the theoretical maximum amount of arsenic that 

would have been in solution if all arsenic in the concentrate had been solubilized. The 

sulfate-to-arsenic mass ratio grew steadily with all pulp densities studied (from 68 to 

290 in 87 days with 15 % p.d., from 33 to 150 in 49 days with 10 % p.d., and from 82 to 

260 in 79 days with 5 % p.d.). The overall level of the ratio decreased with increasing 

pulp density. The sulfate-to-total-arsenic mass ratio plays a role in the arsenic sorption 

capacity of jarosites and will be discussed in more detail in Chapter 8. 

Figure 7.8: Theoretical sulfate-to-arsenic ratio during the 30 L batch experiment 

with 5, 10, 15 % pulp densities (p.d.) of nickel flotation concentrate. Arsenic 

calculated as the theoretical maximum amount released from the nickel flotation 

concentrate. Error bars represent a 95 % confidence interval (± 2σ; n = 2) 

omitting variations in the arsenic content of nickel flotation concentrate. 

48

7.1.3 Bacterial Community 

Figure 7.9 presents the DGGE gel from samples collected at the end of batch 

bioleaching of NFC. The numbered bands were cut out and sequenced in terms of the 

partial 16S rRNA gene after re-amplification by PCR. Table 7.4 presents the sequenced 

bands and the corresponding bacterial species. All reactors by the end of the experiment 

were dominated by the Acidithiobacillus ferrooxidans strain ATCC23270 with the sole 

exception of the Alicyclobacillus sp. represented by band 8. The result is not entirely 

unexpected as the selective factors are many in the long-operational batch systems, 

including arsenic, a notable level of heavy metals concentration and an elevated pH. 

Figure 7.9: DGGE gel used for separation of partial 

bacterial 16S rRNA genes from samples taken at the 

end of the experiments from the 30 L batch reactors 

with 5, 10, 15 % pulp densities (p.d.) of nickel flotation 

concentrate. Sequenced bands are numbered. 

49

Table 7.4 : Sequence data from bacterial samples taken from the 30 L batch reactors at 

the end of the experiments. 

Band Accession no. Description Identity (%) 

1 NR_041888.1 Acidithiobacillus ferrooxidans strain ATCC23270 98 







8 NR_041475.1 Alicyclobacillus sp. N/A 




N/A not available 

50

7.2 20 L Semi-continuous Mode Experiment 

7.2.1 Bioleaching Results 

Nickel 

Figure 7.10 presents the nickel dissolution and the corresponding percent yield during 

semi-continuous mode bioleaching of NFC. The highest nickel leaching yield (12.4 %) 

was measured after 14 days of operation. The maximum nickel leaching rate (265 mg- 

Ni/L/d) was measured between 8 and 9 days of operation. After 14 days of operation, as 

the increase in nickel concentration, i.e., leaching rate started to slow down, the leaching 

yield correspondingly began to decline until the end of the experiment. In other words, 

the rate of nickel bioleaching was lower than the addition of nickel as concentrate, 

indicating that the 10 d retention time in the experiment was too short. 

Figure 7.10: Nickel concentration and yield during the 20 L semi-continuous 

mode bioleaching of nickel flotation concentrate. 

The ratio between the maximum leaching rate and the average nickel addition rate 

as concentrate was only 0.6, which means that nickel leaching was never more rapid 

than nickel addition as concentrate. Ideally, the ratio would be 1, meaning that leaching 

takes place at the same rate as the addition of the mineral to be leached. The inefficient 

bioleaching rate in the experiment was probably caused by too short a retention time 

that resulted in higher wash-out of bacteria compared to the growth rate. 

During the latter half of the experiment, the cyclic, non-constant nature of the mass 

exchange is evidenced by the saw-like profile of the concentration curve. The leaching 

rate appears to be the highest when there is a short break in the mass exchange. The 

following daily mass exchange then seems to result in a slight dip in the leaching rate. 

This supports the hypothesis of the too short a retention time. 

51

Cobalt 

The leaching of cobalt and the corresponding cobalt percent yield during semicontinuous 

mode bioleaching of NFC were as presented in Figure 7.11. The cobalt 

leaching followed a trend rather similar to that of nickel as presented in Figure 7.10. 

The maximum cobalt leaching yield (22 %) was obtained after 14 days of operation. 

The maximum leaching rate of cobalt (10 mg-Co/L/d) was measured between 11 and 14 

days of operation. After 14 days, the leaching rate started to slow down and the yield 

decline until the end of the experiment. The steep rise in cobalt concentration during the 

final days of operation could be attributed to some sudden release of cobalt-bearing 

material from the settled material but it could also be related to some analytical failures. 

Figure 7.11: Cobalt concentration and yield during the 20 L semi-continuous 


The cobalt concentration profile was similar to that of nickel. This corresponds to a 

maximum leaching rate to average cobalt addition rate ratio of 0.9, i.e., cobalt leaching 

at its most intensive was not as fast as addition of cobalt into the reactor as concentrate. 

52

Arsenic 

Figure 7.12 presents the arsenic concentration in the leachate and the arsenic percent 

yield during semi-continuous mode bioleaching of NFC. Arsenic concentration 

increased throughout the experiment, reaching a final concentration of 6.0 mg/L, which 

was considerably higher than in the 30 L batch experiment (< 1 mg/L with all pulp 

densities). This suggests that the retention time was too short for complete arsenic 

immobilization. However, the yield of arsenic remained at around 0.2 % during the 

latter part of the experiment, indicating effective immobilization of arsenic. 

Figure 7.12: Arsenic concentration and yield during the 20 L semi-continuous 


The release of immobilized arsenic may be considerable in a semi-continuous mode 

process where chemical conditions and thus mineralogy change abruptly upon addition 

of large amounts of minerals and chemicals. It might be possible to minimize the 

amount of soluble arsenic with a longer retention time and implementation of a 

continuous-mode process that allows the mineralogy-dictating conditions to be more 

constant. 

53

Leach residue mineralogy 

Table 7.5 presents the proportions of main constituents of leach residue mineral 

composition of bioleached NFC in semi-continuous mode. At best, between 15 and 27 

days, about 9 % (w/w) of the leach residue consisted of nickel-bearing pentlandite. The 

proportional amount of pentlandite in the leach residue was below that of the raw 

concentrate [22.1 % (w/w)] during the whole experiment, indicating that pentlandite did 

not accumulate in the reactor. Goethite and iron sulfate both accounted for 

approximately 25 % (w/w) of the leach residue after 7 days of operation, whereas 

jarosite accounted for 8.8 % (w/w) at best after 23 days. 

N/A not available 

Pentlandite (Fe, Ni) 9S8 22.1 17.2 10.4 9.1 9.2 13.6 11.2 

Pyrrhotite 59.5 24 7 4 3.9 2.4 3.6 

Fe (1-x)S (x = 0 … 0.2) 

Pyrrhotite, oxidized 0 8.8 4.7 4.4 4.4 5.3 5.1 

Fe-sulfate 0 13.5 34.1 25.9 27.9 25.7 26.1 

Fe-sulfate(Ni) 0 0.9 1 1.3 1.5 2.2 2.3 

Fe-sulfate, jarosite N/A 2.7 8 8.5 8.8 5.1 7.9 

Fe-oxide N/A 3.4 2.2 3.2 2.6 4.8 4.2 

Goethite α-FeO(OH) 0 15.7 21.7 31.9 28.1 24.1 27.3 

Pyrite 4.8 5.1 3.6 4.3 4.9 7.2 4.8 

Fe2S Pyrite, secondary 0 0.8 1.7 2.4 3.5 4.5 3.6 

Talc Mg3Si4O10(OH) 2 5.5 1 0.8 0.8 0.7 0.9 0.7 

Magnesite MgCO3 4.3 2.8 1.4 1.5 1.2 2 1.5 

Others 3.8 4.1 3.4 2.7 3.3 2.2 1.7 

Mineral 

Day 

0 2 7 15 23 29 37 

w/w % 

Table 7.5 : Mineral composition of leach residues of nickel flotation concentrate 

bioleaching during the 20 L semi-continuous experiment. 

54

7.2.2 Operational Parameters 

Redox potential and pH 

The redox potential and pH during semi-continuous mode bioleaching of NFC were as 

presented in Figure 7.13. Similarly to the 30 L batch experiment, the redox potential 

reached the +600 mV level in approximately 7 days and remained around +600 mV 

until the end of the experiment. Maintaining the pH at 3.0 was successful with 5 M 

NaOH after the start of acid production by biooxidative reactions. 

Figure 7.13: Redox potential (Eh) and pH during the 20 L semi-continuous mode 

bioleaching of nickel flotation concentrate. 

55

Acid balance 

Figure 7.14 presents the cumulative base (5 M NaOH) consumption during semicontinuous 

mode bioleaching of NFC. After 10 days until the end of the experiment 

NaOH consumption remained constant at around 50 g-NaOH/kg-NFC. Base 

consumption corresponds to acid production from mineral oxidation, i.e., it remained 

quite constant over the course of the experiment. There was no need for pH control with 

acid. 

g-NaOH/kg-NFC 

60 

50 

40 

30 

20 

10 

0 

0 5 10 15 20 25 30 35 40 

Time (days) 

Figure 7.14: 5 M NaOH consumption during the 20 L semi-continuous mode 

bioleaching of nickel flotation concentrate. 

56

Nickel, cobalt, and arsenic mass balance 

The minerals were added into the reactor in constant amounts and the metal 

concentrations in the removed leachate increased constantly. Therefore, it is to be 

expected that at some point the removal and addition of the metals become equal, i.e., 

the process reaches a steady state. Figure 7.15 presents the daily mass balance for 

nickel, cobalt, and arsenic for each mass exchange performed during the experiment. 

Figure 7.15 presents the difference between the amount of an element added into the 

solution as NFC and the amount of the respective element removed from the reactor in 

soluble form. The removal of solid material and the elements therein was not taken into 

account. 

Figure 7.15: Daily mass balance of nickel, cobalt, and arsenic in the logarithmic 

scale during the 20 L semi-continuous mode bioleaching of nickel flotation 

concentrate. A positive value represents net accumulation of the respective 

element in the reactor. Data points are connected with a line for clarity. 

The process did not reach a steady-state with respect to nickel, cobalt, or arsenic. 

This can be seen in Figure 7.15 as a constantly positive addition of each of the three 

elements. Even the build-up of cobalt, which had the lowest net addition level, was at 

least 147 mg with each mass exchange. This rules out the possibility of mass balance 

being the reason for the decrease in nickel and cobalt leaching rates; more nickel and 

cobalt was introduced than what was leached and removed. 

57

Sulfate 

Sulfate concentration during semi-continuous mode bioleaching of NFC was as 

presented in Figure 7.16. The sulfate concentration in the 20 L semi-continuous mode 

experiment shows a logarithmic trend similar to the nickel and cobalt concentrations 

described in Chapter 7.2.1. In comparison, the amount of sulfate during the 30 L batch 

experiment increased continuously. As redox potential retained a high level (~+600 

mV) throughout the experiment and the acid production remained rather constant, it 

indicates that biooxidationalso remained rather constant until the end of the experiment. 

Figure 7.16: Sulfate concentration in the 20 L semi-continuous mode bioleaching 

of nickel flotation concentrate. Error bars represent a 95 % confidence interval (± 

2σ; n = 2). 

58

Total and ferrous iron 

Total and ferrous iron concentrations during semi-continuous mode bioleaching of NFC 

were as presented in Figure 7.17. As in the 30 L batch experiment, the level of total and 

ferrous iron concentrations was low (< 300 mg-Fe(tot)/L; < 100 mg-Fe(II)/L) once the 

redox potential reached a high level and the pH stabilized at 3.0. 

Figure 7.17: Ferrous and total iron concentrations in the 20 L semi-continuous 

mode experiment. 

A distinct increase in total iron concentration occurred between 24 and 36 days of 

operation. Concurrently, the amount of ferrous iron in the solution remained rather 

constant, suggesting the increase in total iron be mainly composed of ferric iron. The 

reason for this transient increase in total iron concentration may lie in the release of 

ferric iron due to changes in mineral composition inside the reactor. This change might 

be triggered, e.g., by a certain iron-to-sulfate ratio. The transient increase in total iron 

occurred when sulfate concentration (Figure 7.16) is above approximately 13 g/L and 

drops again when sulfate concentration declined to around 13 g/L. 

7.3 Shake-flask Experiment for the Assessment of the Role of 

Bacteria in Arsenic Speciation and Immobilization 

Solutions containing ferrous sulfate, arsenite and inoculate from the 30 L batch reactors 

as well as abiotic controls were incubated in shake-flasks (100 mL, 150 rpm, 27 °C) for 

28 days. Two starting pH values were used, i.e., pH 1.7 and 2.5. The aim of the 

experiment was to assess the role of bioleaching microorganisms in arsenic speciation 

and immobilization. 

Redox potential, pH, and ferrous iron concentration for the shake-flask experiment 

were as presented in Figure 7.18 A, B, and C, respectively. The role of microorganisms 

is seen as a rapid increase in redox potential and gradual decrease in pH as well as 

59

extremely low ferrous iron concentration. The abiotic control of the starting pH 2.5 

batch was contaminated between the last two measurement points, which can be seen as 

a quick increase in redox potential and decrease in pH and ferrous iron concentration. 

60

Figure 7.18: Development of redox potential (Eh) (A), pH (B), and ferrous iron 

concentration (C) during the shake-flask experiment with different starting pH 

values (pH 1.7 and 2.5). Error bars of the biotic flask data points represent a 95 

% confidence interval (± 2σ; n = 2). 

61

Arsenic speciation in the experiment for starting pH values of 2.5 and 1.7 was as 

presented in Figure 7.19 A and B, respectively. With both starting pH values as well as 

with biotic and abiotic solutions all soluble arsenic was almost immediately oxidized 

from arsenite to arsenate. This is most likely due to chemical oxidation rather than 

microbiological as the rapid oxidation took place also in the abiotic control solutions. 

After 6 days of incubation most of the arsenic in the biotic solution with the starting pH 

2.5 was in insoluble form and most of the soluble arsenic was present as arsenite. 

Arsenite was also the dominant species in the biotic solution with the starting pH 1.7, 

although there all arsenic was in solution. The occurrence of arsenite was consistent 

with the tendency of many microorganisms to extrude arsenic as arsenite as described 

earlier in Chapter 5.2. Most of the arsenic in the abiotic control solutions remained in 

solution and as arsenate throughout the experiment. 

Figure 7.19: Arsenic speciation in the shake-flask experiment with starting pH 

values of 2.5 (A), and 1.7 (B). 

62

8 DISCUSSION 

8.1 30 L Batch Bioleaching of Nickel Flotation Concentrate 

The 30 L batch experiment showed that selective bioleaching of nickel and cobalt and 

simultaneous immobilization of arsenic in bioleaching of NFC is possible. The elevated 

pH of 3.0 did not inhibit the bioleaching microorganisms and also did not result in 

significant passivation of mineral surfaces by, e.g., the formation of jarosites. Instead, it 

is likely that the elevated pH together with the high redox potential generated by 

biooxidative processes made for conditions that promote the occurrence of arsenic as 

the pentavalent form (arsenate) instead of the trivalent form (arsenite). Arsenate has 

been shown to be less toxic to most organisms (Ng 2005) and also exhibit better 

sorptive features (Asta et al. 2009). Elevated pH also promotes the precipitation of iron, 

e.g., as iron sulfates. Therefore, it can be hypothesized that the arsenic immobilized in 

the process was mainly by sorption on iron precipitates. These include jarosites, 

goethite, and schwertmannite, but the respective role of each mineral in arsenic 

adsorption can only be speculated on in light of this experiment. Also, the possible role 

of microorganisms and their cell surface as an arsenic adsorbent or nucleation site for 

arsenic precipitation should not be overlooked, as suggested by the results of the shakeflask 

experiment discussed later on. 

The fact that the microbial community at the end of the experiment in all reactors 

comprised virtually a single strain of Acidithiobacillus ferrooxidans is definitely 

something to take a closer look at in future studies. Duquesne et al. (2003) reported 

already in 2003 an A. ferrooxidans strain capable of arsenic immobilization. An 

interesting fact pointed out by them was that the soluble arsenic was never oxidized 

from arsenite to arsenate. This is contra to the hypothesis of arsenate adsorption on iron 

sulfate precipitates being the main arsenic immobilization mechanism. As the speciation 

of soluble arsenic in the 30 L batch experiment was not studied, it cannot be stated 

whether the soluble arsenic in this study in association with the dominant A. 

ferrooxidans strain exhibited similar speciation. Further research is required for distinct 

identification of the pathways of arsenic immobilization in the described conditions. 

Also the stability of arsenic immobilized in this way is a matter yet to be evaluated. 

The study on different pulp densities of NFC showed, first and foremost, that 

adequate mixing is a major challenge as the pulp density is increased. In order to 

minimize reactor size and make the bioleaching process overall more economical, pulp 

density is desired to be as high as possible, and, therefore, much consideration should be 

put on different mixing solutions. In the 30 L batch experiment, mixing was enough to 

keep all the particles in suspension in the 5 % p.d. reactor but less so in the 10 % p.d. 

reactor, and definitely not enough in the 15 % p.d. reactor. Lots of settled, for most parts 

unleached mineral material was present on the bottom of the 15 % p.d. reactor. This 

resulted in lower final yields of the valuable metals with respect to, e.g., the 5 % p.d. 

63

eactor, where all of the cobalt was successfully brought into solution. For this reason, 

bioleaching efficiency per se was not significantly lower with 15 % p.d.: it is merely a 

question of sufficient mixing. In fact, the highest nickel and cobalt leaching rates (250 

and 18 mg/L/d, respectively) were achieved with 15 % p.d. In future studies, both the 

use of even higher pulp densities of NFC and corresponding application of more 

efficient mixing should be assessed. 

The high pyrite and pyrrhotite content of NFC proved to eliminate the acid addition 

costs, which can be significant in bioleaching of ores and concentrates of different 

compositions. On the contrary, the acid production from bioleaching of pyrite and 

pyrrhotite in NFC required the addition of base (up to 90 g-NaOH/kg-NFC with 15 % 

p.d. NFC), especially when pH was to be maintained at a certain value. If the pyrite and 

pyrrhotite content of NFC could be reduced by alterations in the ore processing phase, 

also the costs of base addition could be cut down. 

Leach residue analysis showed that goethite formation increased with increasing 

pulp density. Goethite formation is expected to have a positive effect on arsenic 

immobilization as the arsenate adsorption capacity of goethite compared to jarosite is 

much less affected by sulfate concentration (Asta et al. 2009). The increased formation 

of goethite in the 15 % p.d. reactor might have to do with the slightly higher ferrous iron 

concentration and less efficient oxygen transport in the 15 % p.d. system: ferrous iron 

has been shown to catalyze the transformation of schwertmannite [Fe8O8(OH)6SO4] 

into goethite (α-FeOOH) in anoxic aquatic environments (Burton et al. 2008). 

Sulfate has been shown to markedly inhibit the adsorption of arsenic on jarosite but 

less so on goethite. In a study by Asta et al. (2009), a mass ratio of approximately 1.3 

mg-SO4 2- / mg-As(V) was enough to diminish the arsenic adsorption percentage on 

jarosite from the initial percentage of 38 % acquired without sulfate to 4 %. A mass 

ratio as high as approximately 360 mg-SO4 2- / mg-As(V) was required to have a 

significant effect on adsorption on goethite (from 32 to 0 %). In the present study, the 

initial sulfate-to-total-arsenic mass ratio was around 70 … 80 mg-SO4 2- / mg-As with 

each pulp density but never reached the 360 mg-SO4 2- / mg-As level. However, the ratio 

with 15 % p.d. did end up high at 290 mg-SO4 2- / mg-As at the end of the experiment 

after 94 days of operation. Additionally, the growth of the ratio remained somewhat 

stable throughout the experiment with each pulp density, without any sign of decline in 

growth rate. Therefore, based on these results alone, it cannot be stated whether the 

sulfate production from NFC bioleaching could inhibit arsenic immobilization on 

goethite or other iron oxides with similar properties in a full-scale, continuous-mode 

system. However, it can be hypothesized that goethite was a key sorbent for arsenic in 

the 30 L batch experiment. The results showed that the overall level of the SO4 2- /As 

mass ratio was lower with higher pulp density, indicating decreasing relative sulfate 

production with increasing pulp density. This in turn could suggest that arsenic 

immobilization with goethite be more effective the higher the pulp density. 

64

8.2 20 L Semi-continuous Mode Bioleaching of Nickel 

Flotation Concentrate 

The semi-continuous mode experiment was performed in order to shed light on the 

kinetic requirements and restrictions of the NFC bioleaching process. Arsenic yield 

stabilized by the end of the experiment at around 0.2 %, indicating very efficient arsenic 

immobilization. Leach residue analysis revealed that goethite accounted for 20 … 30 % 

(w/w) of the leach residue after one week of operation until the end of the experiment. 

Compared to the 30 L batch experiment [max. 9 % (w/w) goethite with 5 % p.d.], the 

increase in relative goethite content caused by the semi-continuous mode is 

considerable and could easily explain the very low arsenic yield in this experiment. 

Further studies on the conditions of goethite formation during bioleaching of NFC are 

required. 

Nickel and cobalt leaching rates in the course of the experiment showed a saturating 

trend, which was projected as a declining trend in yield after reaching the maximum 

values of 12 and 22 % after 14 days of operation, respectively. In other words, 

unleached nickel and cobalt were added to the system a lot more than what was 

removed in solubilized form throughout the experiment. An explanation to this might be 

that the applied 10 d retention time was too short. The valuable metals were not brought 

into solution at a rate high enough to measure up to the rate of metals addition into the 

reactor as concentrate. As the retention time is desired to be as short as possible and 

thus the bioleaching as fast as possible, one possible direction to take in following 

studies could be to assess the effect of temperature on both bioleaching rate (Rodríguez 

et al. 2003) and arsenic immobilization. The temperature in this experiment was very 

moderate, approximately 21 °C, which is rarely seen in industrial-scale processes, 

where the exothermic nature of the bioleaching reactions results in significant increase 

in temperature (Riekkola-Vanhanen 2010). The change in temperature naturally changes 

the composition of the bioleaching consortium and is therefore almost a completely 

different matter to the present study but is still a justified step to take as most full-scale 

bioleaching processes take place in above ambient temperatures (for a review, see 

Rawlings & Johnson 2007). 

The abruptly changing conditions within the reactor as a result of the addition of 

fresh NFC and growth media in single large quantities at once might provide an 

explanation to this. It could be possible that a species better capable of coping with the 

abrupt additions of large quantities of fresh solution and concentrate out-competed the 

species mainly responsible for the oxidation of the nickel and cobalt -bearing minerals. 

However, as described in Chapter 7.2.2, the redox potential remained at around +600 

mV throughout the experiment, which suggests that microbial oxidative activity was 

high throughout the experiment. Therefore, it is possible that the challenging conditions 

in the reactor drove the microorganisms towards oxidation of gangue minerals like 

pyrite instead of, e.g., the nickel-bearing pentlandite (Mason & Rice 2002). As the 

microbial population was not characterized in this experiment, it is certainly a matter to 

be considered for further research. Also, changing the process more towards the 

65

continuous mode would relieve the stress on the microbial population and might result 

in better leaching rates. A continuous-mode system would also provide for more 

constant mineralogy-defining conditions, which might improve arsenic immobilization. 

In the present experiment, abruptly changing chemical and mineralogical conditions 

caused by the addition of NFC and growth media might result in transformation of 

possibly arsenic-carrying mineral precipitates to another form with different sorptive 

properties, which could result in the release of already once immobilized arsenic as 

reported by Courtin-Nomade et al. (2003). 

8.3 The Role of Bacteria in Arsenic Speciation and 

Immobilization 

The shake-flask experiment with arsenite, ferrous sulfate, MSM and TES was 

performed with the view to assess the fate of arsenic and its speciation as well as the 

role of microorganisms in the immobilization of arsenic in bioleaching. Two different 

starting pH values (1.7 and 2.5) were studied with no pH control during the experiment. 

The results indicated that as no microorganisms were present, most of the arsenic added 

as arsenite was chemically oxidized to arsenate but remained in solution. This is 

unsurprising as the hypothesized mechanism of arsenic immobilization is adsorption on 

iron precipitates. As there was no bioleaching activity, also no iron (sulfate) precipitates 

formed, which left the arsenate without an adsorbent. 

In the flasks inoculated with bioleaching microorganisms, the amount of arsenic in 

solution was markedly decreased in the flask with the starting pH 2.5 and less so in the 

flask with the starting pH 1.7. In both cases, however, most of the arsenic found in 

solution was present as arsenite. This is probably the result of microbial extrusion of intaken 

arsenate as arsenite (for a review, see Slyemi & Bonnefoy 2011). The 

immobilized arsenic was likely adsorbed. The adsorbent could have been either the iron 

(sulfate) precipitates formed as a result of bioleaching or the microbial cells themselves 

(Yan et al. 2010), or both. Arsenic levels dropped dramatically in the abiotic solution 

with the starting pH of 2.5 after its microbial contamination. It is not distinguishable, 

whether the reason for the sudden drop in arsenic concentration was the introduction of 

biomass or iron sulfate formation due to biooxidative activity. Whether or not the 

microbial cell surface can act effectively as an adsorbent or a nucleation site for arsenic 

precipitation, it is clear that the presence and bioleaching activity by these 

microorganisms is elemental to the immobilization of arsenic. 

Arguably the most interesting result received from the experiment was the effect of 

starting pH. The amount of arsenic in solution was significantly lower in the flask with 

the starting pH of 2.5 than in the flask with the starting pH 1.7. After 6 days of 

incubation, probably even sooner, the pH in all biotic flasks was more or less equal, 

meaning that the effect of starting pH had to be rather swift. As the difference between 

the initial pH values is quite small, its effect on the selection of the microbial population 

is not likely to have made a significant difference. However, as the microbial 

populations were not characterized, nothing can be said for certain. One possibility 

66

explaining the different arsenic immobilization results could be that the initial pH might 

have a marked effect on mineral formation. A higher pH could promote the formation of 

iron precipitates of different structure (Ramesh Reddy & DeLaune 2008) with better 

arsenic adsorption capacity and stability (Asta et al. 2009). Yet, no mineralogical 

characterization of the precipitates was performed, which leaves matters open for 

speculation. 

67

9 CONCLUSIONS 

- Selective bioleaching of nickel flotation concentrate (NFC) for nickel and cobalt 

and effective immobilization of arsenic can be maintained at pH 3.0. 

- The high pyrite and pyrrhotite content of NFC results in high acid production 

and thus base addition costs when pH is controlled. The maximum base 

consumption obtained was 90 g-NaOH/kg-NFC with 15 % pulp density. The 

possibility of partial pyrite and pyrrhotite removal from NFC at NFC production 

stage should be assessed and its costs compared to the base addition costs in 

order to improve economical viability. 

- The highest pulp density studied, i.e., 15 % p.d., showed the highest maximum 

nickel and cobalt leaching rates. The use of even higher pulp densities should be 

assessed in order to minimize costs related to reactor volume. 

- Sufficient mixing to keep all NFC particles in suspension must be provided in 

future experiments and reactor designs for efficient extraction of precious 

metals. 

- A retention time of 10 d was too short for a 20 L semi-continuous mode reactor 

where mass exchange was performed once a day. However, with continuous and 

stable material flow, an effective retention time possibly even shorter than 10 d 

could be achieved. The possibilities of continuous-mode reactor configurations 

should be studied in the future. 

- Operating a bioleaching reactor in semi-continuous mode compared to batch 

more resulted in a significant increase in goethite formation, which promotes 

more efficient arsenic immobilization. 

- Biotic and abiotic shake-flask experiments with arsenite and ferrous sulfate 

demonstrated that bioleaching microorganisms play an important role in the 

immobilization of arsenic. The exact mechanism(s) of arsenic immobilization in 

the studied conditions remain, however, yet to be identified. 

68

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