HUB RESEARCH PAPER

ANC, BNC and mobilization of Cr From polluted sediments in Function of PH changes A. Shtiza, R. Swennen, V. Cappuyns and A. Tashko Centrum voor Duurzaam Ondernemen (CEDON) HUB RESEARCH PAPER 2008/26. JULI 2008

Hogeschool-Universiteit Brussel University College Brussels Stormstraat 2, 1000 Brussel, Belgium T: +32 2 608 14 41 T: +32 2 210 12 11 F: + 32 2 217 64 64

ORIGINAL ARTICLE

ANC, BNC and mobilization of Cr from polluted sedimentsin function of pH changes

A. Shtiza Æ R. Swennen Æ V. Cappuyns ÆA. Tashko

Received: 5 December 2007 / Accepted: 24 February 2008

� Springer-Verlag 2008

Abstract During the manufacturing of chromate salts

(1972–1992) large quantities of Chromite Ore Processing

Residue (COPR) were released into a decantation pond east

of the former chemical plant of Porto-Romano (Durres,

Albania), giving rise to yellow colored pond sediments.

These Cr(VI) bearing sediments were deposited upon

Quaternary silty-clay lagoonal sediments rich in iron oxi-

des and organic matter. The pH values in these lagoonal

sediments vary around 6.6, while in the pond sediments, it

is mainly acidic (due to the presence of the sulfur stock

piles in the area and the release of the H2SO4 from the

activity of the former chemical plant), varying between 1.4

and 3.8. Continuous leaching of the COPR waste resulted

in yellow-colored surface water runoff. The prediction of

pH changes in the different types of sediments based upon

acid/base neutralizing capacity (ANC/BNC) jointly with

the quantitative data on release of heavy metals and

especially Cr is considered an important advantage of the

pHstat leaching test if compared to conventional leaching

procedures. Thus, factors controlling the leaching of

Cr(VI), Cr(III), Ca, Al, Fe, Mg from the COPR were

investigated by means of pHstat batch leaching tests and

mineralogical analysis. Moreover, mathematical and geo-

chemical modeling complemented the study. The COPR in

the area contain very high concentrations of chromium

24,409 mg/kg, which mainly occurs as Cr(III) (75–90%) as

well as Cr(VI) (25–10%). The leaching of Cr(VI) occurs in

all the range (2–10) of the tested pH values, however, it

decreases under acidic conditions. Beside some reduction

of Cr(VI) to Cr(III), the Cr(VI) content of the leachtes

remains relatively high in the acidic environment, while the

limning of Cr(VI) pond sediments will increase the release

of the latter specie. The leaching of the Cr(III) occurs

strictly under acidic conditions, whereby limning of these

sediments will give rise to the lower solubility of Cr(III).

The key mineral phases responsible for the fast release of

the Cr(VI) are: the chromate salts (i.e. sodium chromate

and sodium dichromate), while sparingly soluble chroma-

tite (CaCrO4) and hashemite (BaCrO4) release Cr(VI) very

slowly. Thus, pH and mineral solubility have been identi-

fied as key factors in the retention and the release of the

hexavalent CrO42- and Cr2O7

- from the COPR-rich pond

sediments.

Keywords COPR � Cr speciation � Cr(VI) � Cr(III) �pHstat � ANC � BNC � Leaching � Solubility

Introduction

Cr in Albania: occurrences and uses

Albania was ranked (until 1991) among the most important

producers (the third) of chromium ore in the world (Perron

1995), due to the major podiform chromite deposits

occurring mainly in the eastern part of the country, in the

so-called ophiolite complex in the geological zone of

Mirdita (Robertson and Shallo 2000). The mining of large

amounts of chromium ores was expanded with additional

industrial sectors that widely used Cr as raw material.

Production of ferrochromium occurred in smelters near the

A. Shtiza (&) � R. Swennen � V. Cappuyns

Geology Department, Faculty of Sciences,

Katholieke Universiteit Leuven, Celestijnenlaan 200E,

3001 Heverlee, Belgium

e-mail: aurela_shtiza@yahoo.com

A. Tashko

Faculty of Geology and Mining, Geochemistry Department,

Polytechnic University Tirana, Rruga Elbasanit, Tirana, Albania

123

Environ Geol

DOI 10.1007/s00254-008-1263-7

cities of Elbasan and Burrel (from 1979 to date) while

production of the sodium (di)chromate took place in the

former chemical plant of Porto-Romano (which was active

in the period 1972–1992). In Albania, due to absence of

any environmental legislation until 1990 as well as due to

economical restrictions with respect to waste storage, the

areas around the mining and industrial sites became

severely contaminated. In order to study the degree of

contamination, different investigations were undertaken by

Shallari et al. (1998), Dhimo et al. (1999), Sallaku et al.

(1999) and Shtiza et al. (2005a, b, 2008) especially in

Albanian sites related to the treatment of chromium ores.

Cr chemistry

Speciation of an element according to Templeton et al.

(2000) and Hursthouse (2001) is defined as ‘‘the distribu-

tion of an element amongst defined species in a system’’

e.g. Cr(III) and Cr(VI).

Cr(III) of geogenic (natural) origin in soils and rocks

mostly occurs as the mineral chromite (Cr2O3) which is

extremely insoluble (Bartlett and James 1988; Becquer

et al. 2003), while Cr(VI) rarely occurs under natural

conditions (Bartlett and Kimble 1976), apart from the rare

mineral crocoite (PbCrO4) as reported in Zayed and Terry

(2003). Moreover, Cr(III) is considered to be an essential

trace element for the functioning of living organisms

(known as glucose tolerance factor (CrGTF)), while Cr(VI)

is toxic to humans via inhalation (James 1994, 1996;

Whalley et al. 1999) and skin contact can induce allergies

(Yassi and Niober 1988). Cr(VI) is estimated to be about

100 times more toxic and soluble than Cr(III) (Cheung and

Gu 2003). As reported by Turner and Rust (1971) con-

centrations of Cr(VI), as low as 0.5 mg/kg in solution and

5 mg/kg in soils can be toxic to plants.

An increase in the solubility of Cr(III) is observed in

acid conditions especially at pH B 4 according to Fendorf

(1995). Chromate salts (i.e. rich in variable amounts of

Cr(VI)) are mainly of anthropogenic (man-made) origin

and are easily soluble at all the pH ranges predominantly

forming (hydro)oxyanions including hydrogenchromate

(HCrO4-), dichromate (Cr2O7

2-) and chromate (CrO42-)

(Fendorf 1995; McLeod 2001). Moreover, chromate is

weakly sorbed at high pH values and reducing hexavalent

Cr to the less mobile and less toxic trivalent state is also

very difficult at alkaline environments (Deakin et al. 2001).

This reduction might occur due to the presence of the iron

oxides, organic matter as well as reduced sulfur, while the

only known reaction which might oxidize Cr (III) to Cr(VI)

occurs in the presence of manganese oxides (Bartlett and

James 1979; Palmer and Wittbrodt 1991; Fendorf and

Zasoski 1992; Buerge and Hug 1997; Bolan et al. 2003).

Therefore, it is important to know the redox reactions

occurring in a site, since they control the valence state and

the mobility of Cr.

Knowledge of these soil chemical transformations

allows predicting the mobility of Cr in the soil/sediment-

water systems and assisting in the decision to remediate

Cr(VI) enriched soils/sediments via reduction processes.

Therefore, besides, total element concentrations and spe-

ciation, some general sediment features such as grain size,

pH, organic matter and CEC were determined since they

reflect key physico-chemical properties like governing the

distribution, speciation, mobility and availability of chro-

mium and chromates in the study area. Mineralogical

analyses were conducted to provide direct information on

solid-phase speciation of heavy metals and especially of

chromium.

Overall industrial setting and sampling

The so-called high lime process responsible for the gener-

ation of the chromite ore processing residue (COPR) waste

was largely abandoned in the Western World in the late

1960s, but it is still being used in countries such as China,

Russia, India, Pakistan (Geelhoed et al. 2003) and Porto-

Romano (Albania) from 1972 to 1992 (Zelfo 1987). The

industrial COPR waste generated by the manufacturing of

chromate salts (Deakin et al. 2001) contains variable

amounts of Cr(III) and Cr(VI) due to the incomplete

leaching of Cr(VI). In the world millions of tons of COPR

have been deposited in the past in urban areas [Hudson

County, New Jersey (USA) and Glasgow (Scotland)] and

continue to leach Cr(VI) at very high concentrations (Burke

et al. 1991; Farmer et al. 1999; Geelhoed et al. 2003).

This study focuses on the Porto-Romano site (Longi-

tude: E 19.42754; Latitude: N 41.36968) where the sodium

chromate and dichromate salts were intensively produced

during a period of 20 years. During this period more than

100 tonnes of COPR wastes were released as suspended

particles in waste waters or as dissolved species (Zelfo

1987), in the nearby decantation pond, east of the Porto-

Romano chemical plant. The majority of chromium in

COPR is present as Cr(III), partly as unreacted chromite

ore, but 10 to 25% occurs as Cr(VI) as reported for pond

sediments of Porto-Romano in Shtiza et al. (2008). More-

over, yellow crystals (known as chromate blooms, Burke

et al. 1991) could be easily observed on the site especially

after the rainy season. Due, to the high solubility of the

COPR-constituents, a yellow color marks the surface

waters and concentrations up to 168 mg/L could be mea-

sured (Shtiza et al. 2008) which are 3,000 times higher than

Environmental Quality Standards. Similar values have also

been reported from Farmer et al. (2002) for the COPR

wastes occurring in Glasgow.

Environ Geol

123

The subsurface of the polluted site is composed of

quaternary lagoonal swamps rich in silty-clays, which

forms a nearly impermeable layer for the further infiltration

of pollutants in depth. These sediments are rich in iron

oxides and organic matter, while the COPR-rich pond

sediments are mainly sand dominated and poor in iron

oxides and organic matter. The pH of the silty-clays is

neutral to slight alkaline while the COPR-sandy sediments

often are acidic. In Porto-Romano, due to the complexity of

the pollutants (i.e. sodium (di)chromate salts and presence

of sulfur) the acidic and slight alkaline conditions are

dominating in the polluted COPR-pond sediments (Shtiza

et al. 2008), while in other sites (i.e. Glasgow) the pH of

COPR-wastes is rather alkaline (pH [ 11) as reported in

Geelhoed et al. (2002) and Farmer et al. (1999, 2006).

The aim of this study is to understand the factors which

control speciation and solubility of different chromium

species with regard to pH changes based on the investi-

gation of COPR polluted pond sediments and lagoonal

sediments (from Porto-Romano). For this purpose, one

lagoonal sediment and three COPR-pond sediment samples

with different general physical–chemical, Cr-speciation,

geochemical, mineralogical signatures as well as the dif-

ferent acid/base buffering behavior were selected for the

pHstat tests. The lagoonal sediment (sample B1) was

selected since natural attenuation conditions were reported

to occur and no forms of soluble Cr were detected. From

the pond sediment samples, one posses a high solubility of

Cr(VI) (sample B3), and two others are characterized by a

high solubility of Cr(III) (samples F3 and F4).

Methodology

General sample characteristics

In order to differentiate between the pond and the lagoonal

sediments a detailed description was based on field

observations and laboratory measurements.

Determination of the sediment colors was made with the

Munsell (1997) catalog under laboratory conditions, while

granulometry (Gee and Bauder 1986) was determined by

laser diffraction (Malvern Mastersizer S long bed; Malvern,

Worcestershire, UK). Soil-pH (after standard calibration at

pH 4 and 7 of the instrument) was measured in a suspension

solution of 10 g of sediment in 25 mL water, after magnetic

stirring for 30 min and 5 min sediment settling. The pH was

measured in the sediment suspension with TitroWico

Multititrator� (Wittefield and Cornelius, Germany). The

organic matter was determined according to the Walkley-

Black method (Allison 1965). Cation exchange capacity

(CEC) was measured applying the ‘‘silver thiourea method’’

(van Reeuwijk 1992). In Table 1, the general characteristics

of the selected samples are summarized.

Geochemistry

The bulk samples collected for the pHstat batch tests

weighted about 3 kg. The sediments were dried, disag-

gregated gently in an agate mortar, homogenized, sieved,

and the\2 mm fraction subdivided in four parts and one of

them was crushed to fine powder. After homogenization,

finally 0.1 g was digested; 4-acid digestion was applied

(2.5 mL HNO3conc, 5 mL HFconc, 1.5 mL HClO4conc,

2.5 mL HClconc) for total digestion of the pond samples

(Shtiza et al. 2005a). The term total concentrations will be

used for the geochemical results obtained from the 4-acid

digestion. Total concentrations of Cr, Ni, Co, Zn, Cu, Mn,

Fe, Ca, Mg, K and Al were measured with an Flame

Atomic Absorption Spectrophotometry (FAAS) Varian

AA6 and Varian AA-1475. The detection limit for most of

the elements varied around 1 mg/kg. All used reagents

were of analytical grade. The results of total concentrations

of the selected sediments are summarized in Table 2.

International soil standards (i.e. SO-1 Regosolic and

GBW07411) were added to series analyzed with the 4-acid

method. Values for three replicates of the SO-1 standard (at

95% confidence level) were (in mg/kg) for Cr 159 ± 12.5

(certified 160 ± 15); Cu 63 ± 7.5 (certified 61 ± 2.9); Ni

89.2 ± 10.2 (certified 87 ± 6); Zn 147 ± 6 (certified

146 ± 5). In two replicates of the GBW07411 soil stan-

dard, the concentrations were (in mg/kg): Cr 57.2 ± 3

(certified 59.6 ± 5.0); Co 9.5 ± 3 (certified 11.6 ± 1.4);

Cu 63 ± 4.1 (certified 65.4 ± 4.7); Ni 23.6 ± 2.5

(certified 24.2 ± 2.1) and Zn 3,700 ± 100 (certified

3,800 ± 100).

Table 1 General characteristics of the selected pond and lagoonal sediments from the former chemical plant of Porto-Romano (Durres, Albania)

Bulk samples Depth (cm) Sediment type Munsell classification color Clay % Silt % Sand % Org. C % pH H2O CEC (cmol/kg)

B1 45–105 Lagoonal sediment 5Y 5/3 Grayish Olive 15 83 2 1.8 6.6 34

B3 0–30 Pond sediment 5Y 6/3 Olive Yellow 9 55 36 0.3 3.8 11

F3 3–17 Pond sediment 10YR 4/4 Brown 15 73 12 1.9 3.3 29

F4 0–3 Pond sediment 5Y 6/3 Olive Yellow 9 63 28 2.6 1.4 6

Environ Geol

123

Cascade leaching test (CLT)

A test designed to assess the maximum leachable amount

of metals, which becomes available for leaching under

normal conditions was adopted from NEN 7349 (1995).

Sample material with a grain size \2 mm is used for the

cascade leaching test (CLT). A measure of 1.5 g of sedi-

ment was suspended in a 30 mL of deionized water

(pH = 4) in an acid-rinsed extraction tube and shaked

using a mechanical shaker for 12 h. After centrifugation

(10 min, 3,000 rpm), the eluate was filtered through a

\0.45 lm Millipore filter. The leaching procedure is

subsequently repeated five times resulting in (L/S) ratios of

20, 40, 60, 80 and 100 L/kg dry material. The water in final

leached eluate had a transparent color, in contrast to the

first eluates which were yellow colored, due to the disso-

lution of chromate salts. The results presented in Table 3

show the total amounts of the phases, which become sol-

ubilized in the five subsequent extractions.

The residue remaining after the five steps of the CLT

was analyzed with X-ray powder diffraction (XRD), which

proved that within the detection limit of the XRD no

Cr(VI) phases were present any more, which suggested that

the remaining Cr was occurring as insoluble.

pHstat batch leaching tests

pHstat leaching experiments are batch leaching tests in

which automatic titrators are used to keep pre-defined pH

values constant over time. Such experiments can be

efficient to assess the long-term effects of pH variation

with regard to the leaching characteristics of the heavy

metals (Forstner and Hasse 1998). Different pre-defined

values (acid or alkaline) can be used during the leaching

experiment in order to study the potential release of the

heavy metals under these conditions. The quantity of acid

or base added during the pHstat experiment to keep the pre-

defined pH value constant during the experiment is known

as acid neutralizing capacity (ANC) or base neutralizing

capacity (BNC), respectively. It should be stressed that

obvious differences exist between the pHstat leaching tests

and the ‘‘in situ’’ pore water situation (van Herreweghe

et al. 2002). These differences relate to the natural leaching

conditions, temperature, acid rain precipitations as well as

the biological activity. Since the main investigation will

focus on the effect of the pH in Cr solubility and specia-

tion, the other factors are considered as of minor impact.

pHstat batch experiments (Fig. 1) were carried out with

an automatic multi-titration system (Titro-Wico Multiti-

trator�, Wittefield and Cornelius, Bochum, Germany). An

amount of 80 g (\2 mm) fraction was put in each Erlen-

meyer flask with 800 mL of distilled water and placed on a

horizontal-shaking device. A pH-electrode and an auto-

matic titration dispenser were attached to each flask. From

earlier experiments performed by Paschke et al. (1999),

24 h pHstat tests were insufficient to determine accurately

the heavy metal release, as well as the buffering capacity

of the system. Batch tests carried out by Deakin et al.

(2001) on samples rich in COPR have shown that the Cr

release occurs rapidly in the first hours, but after 96 h no

Table 2 Summary of the total concentrations in the selected samples

Bulk samples Ca

(g/kg)

K

(g/kg)

Mg

(g/kg)

Al

(g/kg)

Na

(g/kg)

Fe

(g/kg)

Ni

(mg/kg)

Zn

(mg/kg)

Cu

(mg/kg)

Co

(mg/kg)

Mn

(mg/kg)

Cr

(mg/kg)

B1 1.81 2.97 2.39 6.42 7.84 6.04 319 169 71 72 460 1,710

B3 5.27 0.09 1.72 4.49 9.02 1.66 153 45 36 53 201 3,452

F3 4.16 1.92 1.95 6.45 0.60 4.75 230 155 110 41 237 24,409

F4 11.38 0.52 0.39 1.09 0.39 1.16 34 19 73 45 146 2,814

The differences between the lagoonal sediment (B1) and the pond sediments (B3, F3 and F4) are clear if major elements (i.e. Ca, K, Mg, Na and

Fe), and Cr are compared

Table 3 Summary of the cumulative soluble concentrations of the selected samples after five consequent steps of cascade leaching test (CLT)

Bulk

samples

Ca

(g/kg)

K

(mg/kg)

Mg

(mg/kg)

Al

(mg/kg)

Na

(g/kg)

Fe

(mg/kg)

Ni

(mg/kg)

Zn

(mg/kg)

Cu

(mg/kg)

Co

(mg/kg)

Mn

(mg/kg)

Cr(VI)

(mg/kg)

Cr(III)

(mg/kg)

SO42-

(g/kg)

B1 0.46 835 956 D.L. 0.25 D.L. D.L. D.L. D.L. D.L. N.A. \0.1 \0.1 3.27

B3 1.77 13 295 D.L. 0.42 D.L. D.L. D.L. D.L. D.L. N.A. 422 504 8.63

F3 2.77 299 1,901 258 0.18 D.L. D.L. D.L. D.L. D.L. 9 \0.1 88 8.34

F4 5.03 235 1,724 685 0.21 679 D.L. D.L. D.L. D.L. 10 \0.1 574 12.42

D.L. Below detection limit; N.A. Not analyzed

Environ Geol

123

further chromium is released in solution, thus, Cr con-

centrations are likely to remain constant towards the end of

the experiment. Hence, 96 h pHstat batch tests appear to

give a good representation of the release of the heavy

metals as well as data with regard to ANC and BNC

evolution for the selected samples. The suspensions were

first shaken for 30 min without the addition of any

chemicals. Previous experiments performed in our labo-

ratory (van Herreweghe et al. 2002; Cappuyns et al. 2004)

demonstrated that the rapid addition of the acid/base to the

sediment suspension could lead to the over passing of the

pre-defined pH. Therefore, the concentration of the titra-

tion solution was adapted to the set-point pH as

summarized in Table 4. The choice for the concentration

of the titration was based on the preliminary tests. Thus,

the interval of pH acquisition was set to 200 s. By doing

so, after adding acid/base to the suspension the system has

enough time to react and eventually neutralize the acid/

base before more acid or base is added. At regular time

intervals (0, 0.5, 1, 3, 6, 12, 24, 48, 72 and 96 h) a sample

of the suspension was taken over a filter (0.45 lm Acro-

disc, Pall; East Hills, NY) by means of a syringe attached

to a flexible tube (Fig. 1).

For a better understanding of the reactions responsible

for pH buffering, slow potentiometric titrations were car-

ried out on the four samples (B1, B3, F3 and F4). During

the pHstat experiments, different buffers were added to the

respective sediment/water suspensions in order to maintain

constant the pH values. The pH was continuously moni-

tored along all the duration of the experiment. Moreover,

ANC/BNC and the leaching behavior of the Cr species (i.e.

III and VI) and other elements are investigated. The results

from the pHstat tests allow inferring the mobility of the

contaminants, simulating changes that might occur in the

natural environment with respect to pH. Moreover, a blank

pHstat test (in which no acid or base was added) was carried

out for 96 h on each of the samples to investigate the

behavior of the pH as well as the solubility of the different

elements under natural conditions.

Immediately after sampling, the electrical conductivity,

redox potential as well as Cr(VI) (see below), sulfate

(Vogel 1961) and chlorite (Vogel 1968) concentrations

were measured. Moreover, some samples were analyzed

for total organic carbon (TOC) content with a Therm-

aloxTM TOC (Analytical Sciences Ltd, England). Finally,

the samples were acidified with a drop of concentrated

HNO3 (ultra pure) to bring the pH to *2. Subsequently,

the sample was kept in refrigerator prior to additional

analysis. Concentrations of Ca, Mg, Al, Fe, K, Na, Mn, Ni,

Cr, Zn, Cu, Co, Pb and Ba were measured by FAAS for the

major elements and inductively coupled plasma-mass

spectroscopy (ICP-MS Hewlett Packard 4500 series; Palo

Alto, CA) for the trace elements.

Geochemical modeling

The quantities of soluble species measured by the pHstat

leaching procedure and the respective pH values (Table 5),

were used as input data to calculate the main species dis-

tribution by means of the Visual MINTEQ version 2.40

(Gustafsson 2006) thermodynamic geochemical speciation

model. The database from Visual MINTEQ version 2.40

was used to assess the species distribution.

Moreover, the data derived from the XRD analysis were

combined with the data derived from the geochemical

modeling in order to compare the main mineralogical find-

ings and estimate the solid state sources which contributed to

the main solubilized species. Due to the rather low detection

limit of the XRD analysis and the difficulty to trace minor

mineralogical phases, the combination of the geochemical

modeling was considered important in order to be able to

confirm even the presence of these mineralogical phases.

Mathematical modeling

ANC and leaching of metals as a function of time was

described mathematically based on the acquired pHstat data

by the use of MATLAB (The mathworks; Natick, MA) and

Fig. 1 Schematic layout of the pHstat batch system. a Erlenmeyer, bhorizontal shaker, c soil suspension, d pH electrode connected with a

computer for continuous pH registration, e automatic computer

controlled titration dispenser, f flexible tube, g filter, h syringe to

sample the eluate from the suspension

Environ Geol

123

EXCEL (Microsoft; Redmont, WA) softwares. ANC and

BNC curves obtained by the pHstat tests with a continuous

set-point titration were described according to Schwartz

et al. (1999). The proton buffering capacity (BC) of soils

during the pHstat experiments can be described as the sum

of two independent first-order reactions:

Hb tð Þ ¼ BC1 1� exp �kitð Þð Þ þ BC2 1� exp �kitð Þð Þ ð1Þ

Hb(t) (mmol/kg), corresponds to the buffered protons at a

time t; BC1 (mmol/kg) is the buffering capacity of the

system 1, ki (per h) is the rate coefficient of the buffer

system i and t (h) is the time after starting the titration. The

BC is defined as the total amount of H+ that can be neu-

tralized by a buffer system, while buffered protons are those

which are neutralized from reactions with the soil matrix.

In analogy with the earlier reaction, the release capacity

of an element (m) at a time (t) can be described by the

following reaction:

RLm ¼ RC1 1� exp �r1tð Þð Þ þ RC2 1� exp �r2tð Þð Þ ð2Þ

RC1 is the release capacity of buffer system 1 (mg/kg), r1 is

the rate coefficient of the system 1 (per h), and t is the time

after starting the titration (h). The release capacities of the

two buffer systems (RC1 and RC2) can be considered as two

dominant sinks for heavy metals from which the elements

are released with different rates. Important to be stressed is

the fact that the release capacities and the rate coefficients

of this empirical fit can not be assigned to well-defined

buffer substances (Schwarz et al. 1999), thus the selection

of only ‘two buffer systems’ or ‘pools’ is operationally

defined. In theory, it is possible to consider more than just

two components since it is known that soils consist of dif-

ferent components like clays, organic matter and Fe, Mn, Al

hydroxides, etc. To make an accurate description of the

mechanisms, much more complex reaction descriptions are

needed, however, making the understanding of the reaction

mechanisms complex and difficult.

Cr speciation

Speciation analysis is a term used to identify and/or mea-

sure the quantities of one or more individual species in a

sample (Templeton et al. 2000). The determination of

Cr(VI) in the eluates (e.g. CLT and pHstat) and water

samples was carried out applying the colorimetric diphe-

nylcarbazide (DCB) method (USEPA 1995) versus

calibration standards in the range 0.5–2.0 lg/L. In order to

achieve accurate results with regard to the chromium spe-

ciation, analysis were carried out within the day the eluates

were collected.

In presence of Cr(VI), the reaction with DCB gives a

red-violet color to the solution. After 5 min reaction, the

absorbance of the sample is measured with an UV–VIS

Spectrophotometer 635, at a wavelength of 540 nm. To

check the accuracy of the Cr(VI) colorimetric method, the

eluate samples were also measured for soluble Cr by AAS.

If no red-violet color is formed in presence of DCB, but Cr

is measurable by AAS, this was considered to be Cr(III).

Mineralogy

Quantitative analysis by XRD is a difficult task, especially

when the samples contain complex mixtures of phases, as

is often the case with samples from highly contaminated

sites, or from waste materials (Hillier et al. 2003). The

complexity of many such mixtures and the resulting XRD

patterns mean that quantitative methods based on the

measurement of single/several resolved peaks are often not

feasible. However, a combination of geochemistry and

XRD were applied to deduce the main mineralogical

composition of the different solid sampling media (i.e.

insoluble residue after 4-acid digestion as well as in sedi-

ments from the CLT and pHstat).

XRD analysis were carried out using a Philips� powder

diffractometer with Co radiation tube, (k-1.7902 A). The

samples were scanned from 5-75� 2h, in steps of 0.02� at

1 s/step. All mineralogical determinations were carried out

in duplicates.

Results

Major and trace element mobility during

the pHstat batch tests

The results for the final eluates of the pHstat tests are given

in Table 5. During the pHstat leaching test of the selected

Table 4 Summary of the operational parameters for the pHstat

leaching test

Operational parameters pHstat leaching

test

Blank pHstat leaching

test

Solid/liquid ratio 1:10 1:10

pH Pre-defined Natural

Concentration of

HNO3 solution (M) 2 2.5

4 1

6 0.25

NaOH solution (M) 6 0.25

8 1

10 2.5

Titration speed Variable (pH dependent)

Shaking frequency (per min) 175 175

PH measurement interval (s) 200 200

Duration (h) 96 96

Environ Geol

123

Ta

ble

5T

he

resu

lts

fro

mth

ep

Hst

at

leac

hin

gte

sts

for

the

fou

rse

lect

edsa

mp

les

mg

/kg

dry

wei

gh

t

Sam

ple

pH

Fin

al

AN

C/B

NC

(mm

ol/

kg

)

Eh

(mV

)

DO

C

(mg

/L)

EC

(mS

/cm

)

Ca

Mg

Al

Fe

SO

42-

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KN

aM

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r(II

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83

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1,2

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42

N.A

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50

31

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11

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41

21

74

41

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33

8

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61

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N.A

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37

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N.A

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n7

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N.A

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3N

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14

61

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B1

81

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-8

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N.A

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06

92

97

26

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32

N.A

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35

B1

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–-

64

N.A

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.A.

17

50

84

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.

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99

36

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.A.

N.A

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89

N.A

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.A.

N.A

.0

\0

.13

83

91

10

23

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.1N

.A.

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n4

.2–

N.A

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.A.

N.A

.6

,27

8N

.A.

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21

34

,22

1N

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53

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.L.

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40

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16

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60

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77

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75

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83

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05

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48

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28

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22

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0.1

56

61

53

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n4

.0–

38

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.A.

2.5

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,58

91

,58

81

51

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35

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15

51

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.

F3

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24

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92

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N.A

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82

F3

88

28

N.A

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.L.

8.7

44

,71

78

96

11

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3

Th

efi

nal

pH

,ac

ido

rb

ase

neu

tral

izin

gca

pac

ity

(AN

C,

BN

C),

red

ox

-po

ten

tial

(Eh

),d

isso

lved

org

anic

carb

on

(DO

C),

elec

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nd

uct

ivit

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ach

ing

of

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ents

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vy

met

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ral

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bel

ow

det

ecti

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lim

it

Environ Geol

123

samples without pH adjustment, a continuous increase in

the pH was observed. The pH values measured at the end of

these tests are 1.1, 0.4, 0.6 and 1.3 units higher for samples

B1, B3, F3 and F4, respectively, during the 96 h of the

experiment as can be deduced if pH values from Tables 1

and 5 are compared.

In the following text, the percentages given always refer

to the percentage released if compared to the total con-

centration measured in the samples (Table 2). The focus,

mainly, will be on the release of Cr, since the released

amounts of the other heavy metals were often below the

detection limit.

In sample B1, the most important leaching of heavy

metals occurs at pH 2. Leached amounts in percentage for

different metals are given between brackets Mn (with

17.9%), Cr (with 8.2%), Cu (with 7.7%), Co (with 5%), Zn

(with 4.7%) and Ni (with 4%). Moreover, the leaching of

the heavy metals in the pH range 4–8 is negligible. Only at

pH 10, important leaching occurs for Zn, Cu and Cr,

respectively 5.4, 4.4 and 0.1%. No Cr(VI) was detected in

the eluates collected from the pHstat leaching tests of sample

B1. This indicates that Cr which is leached at pH = 2 and

10 is considered to be entirely Cr(III). With regard to the

major elements the highest leacheable amount is recorded

for Ca, for which the release decreases from 44.1 (at pH 2)

to 0.8% (at pH 10). The leaching of Fe and Al is highest at

pH 2 (respectively 2.3 and 1.9%). The leaching behavior of

sulfates appears to be independent from the pH.

In sample B3, the leaching of heavy metals is highly

variable. Mn, Ni, Cu release at pH = 2 is, respectively,

only 0.2, 1.9 and 5.5%, while under alkaline conditions

these elements are not released. The release of Zn at pH

values of 8 and 2, respectively, was between 9.3 and 22%

and for Co it was leached 6.4% (pH = 2) and in unde-

tectable amounts in the other pH values. The most

important mobilization of Cr occurs under alkaline condi-

tions (e.g. pH 10 with 19%) while under natural pH-soil

conditions (pH = 4.2) the release of Cr is about four times

lower (only 5%). The speciation tests showed that leached

Cr from this sample is dominantly Cr(VI), while under

extreme alkaline and acidic conditions also important

quantities of Cr(III) are released. The release of Ca is

important with maximal values of 11.9 and 15.8% recorded

at pH values 3.8 and 2, respectively.

Due to the acidic pH (3.97) of sample F3, the most

important leaching occurs at this pH (Table 5). The per-

centage release of measured elements (given in brackets) in

decreasing order is: Ca (23.7%) [ Mg (8.2%) [ Zn

(4.5%)[ Mn (4%)[ K (0.8%) [ Cu (0.4%) [ Al (0.2%)[Cr (0.2%). Important releases mainly for the major

elements are also recorded for pH 6 and 8 in the following

order Ca (11.4%), Mg (4.61%), K (0.47%), Fe (0.1%) and

Al (0.02%). No Cr(VI) was detected in the alkaline eluates

from this sample, while the solubility of sulfates increases

in alkaline conditions (pH = 8).

Sample F4, displays a similar leaching behavior as

recorded for sample F3, although the mobilized fractions

differ in their order of magnitude. As the natural soil pH of

the sample is 2.7, the majority of the elements are easily

released, i.e. Mg (42.5%), Zn (41.7 %), Cu (28.4%), Cr

(19.1%), Ca (8.2%), Mn (6.9 %), Fe and Al (5.9 %) and

finally, K (3.5%). Lowest leaching for Mg (33.7%) and Ca

(4.5%) occurs under alkaline (pH 8) conditions. With regard

to the speciation of Cr, no hexavalent Cr was detectable

from the pHstat eluates of this sample. The mobilization of

sulfates records its highest concentrations at pH = 2.7.

ANC/BNC in the samples

pH is one of the most important parameters which deter-

mines the mobility of the major and trace elements. The

quantity of the acid or base added for 96 h, to the soil/water

suspension to keep constant the pre-defined pH value

versus the buffering capacity of the system is reported as

ANC or BNC (acid neutralizing capacity or base neutral-

izing capacity) of the system. In order to estimate the ANC/

BNC and mobility of pollutants in the selected samples a

differentiated approach consisting in: the use of pHstat,

mathematical modeling, geochemical modeling and min-

eralogy was applied.

Estimation of ANC/BNC by pHstat

The ANC/BNC highly depends on the natural pH of the

sample, pre-defined pH in the pHstat experiment, titration

speed and whether the maximal value has been reached at the

end of the experiment. During the pHstat leaching tests, the

ANC/BNC are calculated as the amount (mL) of acid/base

added, multiplied by their respective molarities and divided

by the weight of the sample in kg. In Table 5, the ANC,

BNC, DOC are reported next to the release of the major

elements and heavy metals during the pHstat experiments.

The link between ANC/BNC and release of Cr is well

expressed in samples B3 and B1 as shown in Fig. 2, but

absent in samples F3 and F4. The mobility of Cr(III) spe-

cies in the samples B1, F3 and F4 in acidic environments is

high, consequently the liming of these sediments, up to

neutral pH values will reduce considerably the mobility of

the Cr(III) species which in the sample F4, at soil pH = 2.7

has released up to 538 mg/kg.

As representative examples the ANC and BNC curves for

the respective pH values the samples B1 and B3 are given in

Fig. 3. From Fig. 3, it is clear that the ANC and BNC curves

display an asymptotic behavior as a function of time. In the

beginning of the experiment the quantities of the acid or

base added are high but during the experiment the ANC/

Environ Geol

123

BNC curves gradually become saturated and towards the

end of the experiment an almost horizontal pattern is

observed. However, as can be seen in the pattern at pH 10

for sample B3 the plateau level of the BNC is not entirely

reached even after 96 h of the pHstat experiment. Moreover,

it is noteworthy to stress the fact that the high ANC

(pH = 2) of the lagoonal clays (sample B1; 818.2 mmol/

kg) contrasts with the lower ANC value of the pond sedi-

ments (sample B3; 398.5 mmol/kg; Fig. 3). The contrasting

ANC values illustrate the capacity of the lagoonal sediments

(B1) in neutralizing large amounts of the acid effluents.

The release pattern of Cr in samples B1 and B3 at pH 2

is selected due to the important mobilization of Cr under

acidic conditions. In sample B1 (Fig. 4a), trivalent Cr is

released starting from 0.5 h after the starting of the pHstat

experiment (12 mg/L) while towards the end of the

experiment the Cr(III) increases gradually until reaching

252 mg/L as shown at 96 h. In sample B3, Cr(VI) was the

only species released during the first 24 h, but latter on

until towards the end of the experiment (24–96 h), also

Cr(III) species were contributing to the total amount of Cr

released (Fig. 4b). This contribution can be assessed from

the difference between total chromium and Cr(VI). The

amount of Cr(III) varies between 1 mg/L at 24 h, 21 mg/L

at 48 h and up to 38 mg/L at 96 h of the pHstat experiment.

Mathematical modeling of ANC/BNC

In an attempt to infer whether ANC/BNC were fully

exhausted after the entire duration of the experiment, i.e.

96 h, the ANClong term and BNClong term were calculated as

the sum of their buffering capacities.

ANC=BNClong term ¼ BC1 þ BC2: ð3Þ

The estimation for the long term ANC and BNC was made

based on the modeling of the measured ANC and BNC

Fig. 2 Release of Cr in the four

selected samples in association

with the ANC and BNC values

in the respective samples. In

sample B3, the released Cr

species were Cr(VI) and Cr(III),

while from the other samples

only Cr(III) was released,

mainly under extreme acid

conditions (pH 2). The

rectangle indicates the leaching

of the Cr during the blank pHstat

Fig. 3 Example of ANC and BNC at different pH values of the pHstat in the samples B3 and B1, respectively

Environ Geol

123

(Cappuyns et al. 2004). Buffering capacities (BC), rate

coefficients (k), and measured and modeled ANC/BNC values

are given in Table 6. Measured and modeled values of ANC

and BNC of all the samples display relatively good fitting

(Table 6) and asymptotic behavior as a function of time.

The ANC and BNC at the respective pH values of 2 and

8 for sample B1 are nearly completely exhausted with

about 97.8 and 97.9%, respectively, to be consumed after

96 h according to the modeling (Table 6).

With respect to sample B3, the modeling suggests that

99.9% of the ANC at pH = 2 is exhausted by the end of the

experiment (96 h). As the set-point pH increases gradually

from 6 to 10 the contribution of the rapid buffer (BC1) to

the BNC increases from 64 mmol/kg (pH 6) up to

535 mmol/kg (pH 10). An illustration of the BNC curve

and the rapid and slow buffer constants (respectively, BC1

and BC2) versus time is shown in Fig. 5 for the sample B3

at pH 10. The constant (BC2) related to the slow buffer

appears to play an important role especially towards the

end of the experiment in this sample.

The exhaustion constants for sample F3, as calculated

from the BNC values for the 96 h pHstat are overestimated

from the model (103 and 108% at the respective pH values

6 and 8). This is due to the higher BNC values calculated

from the model (453 and 893 mmol/kg at pH values 6 and

8, respectively) if compared to the measured values (440

and 829 mmol/kg).

The BNC in sample F4 at pH 6 and 8 is not exhausted

since the modeling indicates that only 92 and 97%,

respectively, are consumed after 96 h. During the modeling

of the BNC in this sample, the contribution of the slow

buffer (BC2) at pH 6 and 8 (i.e. 200 and 252 mmol/kg,

respectively) appears to play an important role in such a

BNC pattern (Table 6).

The release of the major elements like Ca, Fe and Al

might give an idea with respect to the buffering mecha-

nisms taking place during the release of the heavy metals.

The amount of these elements released at pH 2 after 96 h is

given in Table 5. These data indicate that at pH = 2 dis-

solution of Fe- and Al-phases hardly contributes to the

ANC, with respective values below 2.1 and 1.8% of their

total concentrations, compared to up to 52% of the total Ca

concentration.

Mathematical modeling of the Cr release

In order to infer the contribution of the different Cr-bearing

mineralogical phases in the soluble forms, the release of

trivalent and/or hexavalent Cr was mathematically mod-

eled for the samples B1 and B3 (Table 7). However, notice

that the mathematical model overestimates the Cr(III)

released (135%) at pH 2 for sample B1.

For sample B3 at pH = 2, the release of total Cr is

estimated to be 94%, from which 98% corresponds to

Fig. 4 a, b Leaching of Cr-

species from samples B1 and B3

at pH 2 as a function of time in

comparison with the measured

ANC

Table 6 Buffer capacities (BC)

and rate coefficients (k)

calculated in MATLAB by

fitting the ANC/BNC curves

according to Eq. 1

The percentage of the

(calculated) ANC/BNC which

was exhausted after the 96 h of

the pHstat is expressed as

BCex% = (BC1 + BC2)/ANC

or BNC 9 100

Sample pHstat

values

BC1

(mmol/kg)

BC2

(mmol/kg)

k1

(per h)

k2

(per h)

R ANC/BNC BCex %

Modeled

(mmol/kg)

Measured

(mmol/kg)

B1 pH 2 222 579 0.70 0.06 0.996 801 819 97.8

pH 8 67 43 0.71 0.05 0.990 111 113 97.9

B3 pH 2 83 315 15.91 0.04 0.994 398 398 99.9

pH 6 64 82 0.62 0.01 0.997 146 127 115.5

pH 8 153 138 0.29 0.02 0.998 290 262 110.9

pH 10 535 380 0.08 0.01 0.997 916 830 110.3

F3 pH 6 163 290 0.80 0.04 0.997 453 440 103.0

pH 8 277 616 0.41 0.02 0.999 893 829 107.7

F4 pH 6 51 200 697.66 0.06 0.975 252 272 92.4

pH 8 179 252 32.53 0.07 0.984 432 445 96.9

Environ Geol

123

Cr(VI). In the other pH values, there is a trend towards the

increase of the BNC towards the end of the experiment,

while the release of Cr(VI) (in pH 6 and 8) contributes with

89.6 and 77% in the respective pH values. At pH 10, the

release of Cr corresponds to 71.8% of the total Cr, of which

67.4% corresponds to hexavalent Cr (Table 7). To be

noticed in Table 7, is the impact of the slow release

capacity (RC2) of Cr in the presence of both Cr species,

while when only Cr(VI) is present (i.e. B3, pH 6 and 8) the

rapid release capacity (RC1) appears to be of primary

importance.

Since in samples F3 and F4, Cr(III) was released only in

the natural acidic pH conditions but not during the pHstat,

the Cr release in the latter samples was not modeled.

Sediment acidification

Data from the pHstat leaching tests jointly with data from

acid deposition, allow to predict (in long term) the time

required for sediment acidification. To estimate the time

period (years) over which the ANC and BNC (data from

Table 5) might become exhausted the following equation is

used:

Time ðyearsÞ ¼ ANC96h or BNC96h mmol=kgð Þ3000 mmol acid=ha=yearð Þ�2 hað Þ

mass kg=hað Þð4Þ

The calculation was carried out for the upper surface

(20 cm) of an area of 2 ha (study area of Porto-Romano).

Knowing that changes in the buffering capacity due to

acidification normally occur in the uppermost layers, the

selection of 20 cm thickness is of relevance, however, the

depth of the sediments can be extended further if needed.

With respect to the bulk density, a value of 1.4 ton/m3 was

used. In 1989, the deposition of the total acid in Albania

was estimated to be around 3000 mmol/ha/year (EDACS

1995).

It is clear from the data in Table 8 that it would take

between 40 and 310 years to have substantial changes in

the pH of the studied sediments. If the ANC of the lagoonal

or pond sediments is high, (Table 8) more than 306 and

149 years, respectively are needed in order to have sub-

stantial changes to reach, i.e. pH = 2 in the lagoonal (B1)

Table 7 Release capacities

(RC) and rate coefficients (k)

calculated in MATLAB by

fitting the Cr total and Cr(VI)

concentrations according to

Eq. 2

The percentage of the

(calculated) ANC/BNC which

was exhausted after the 96 h of

the pHstat is expressed as

BCex% = (BC1 + BC2)/

Cr 9 100

Sample pHstat

values

Cr

released

RC1

(mmol/kg)

RC2

(mmol/kg)

K1

(per h)

K2

(per h)

R Cr RCex

(%)Modeled

(mg/kg)

Measured

(mg/kg)

B1 pH 2 Cr (III) 99 241 0.10 0.01 0.999 341 252 135.2

B3 pH 2 Cr total 131 295 3.14 0.07 0.979 426 453 94.0

Cr (VI) 106 277 51.15 0.10 0.981 383 391 97.8

pH 6 Cr (VI) 208 39 75.69 0.04 0.934 246 275 89.6

pH 8 Cr (VI) 234 152 1.60 0.03 0.969 387 502 77.0

pH 10 Cr total 69 413 42.88 0.09 0.991 482 671 71.8

Cr (VI) 74 341 52.66 0.11 0.986 415 615 67.4

Fig. 5 Example of the measured and mathematical modeled BNC

and buffering capacities in the sample B3 at pH 10, calculated

according to Eq. 1

Table 8 Estimation of time (years) to predict the effect of acidifi-

cation in the behaviour of these sediments (for more explanation see

text)

Sample pHstat values ANC/BNC measured

(mmol/kg)

Time calculated

(years)

B1 2 819 306

8 113 42

B3 2 398 149

4 230 86

6 127 47

8 262 98

10 830 310

F3 6 440 164

8 829 310

F4 6 272 102

8 445 166

Environ Geol

123

and pond (B3) sediments. However, pH changes and time

frame for possible changes highly depend on the natural pH

of the samples and the sewage of pollutants occurring in

the area. Moreover, important to be stressed is the fact that

the result in years should be used as an order of magnitude

rather than exact years (van Herreweghe et al. 2002).

Mineralogical phases and modeling

The characterization of the soluble phases was assessed by

a combination of the XRD and geochemical modeling

Visual MINTEQ 2.40 (Gustafsson 2006) jointly with the

respective pH values. The database from Visual MINTEQ

2.40 was used. The solubility of Cr from lagoonal sediment

B1 is rather low under normal conditions (pH 6 and 8) as

observed from the pHstat leaching test results (Table 5).

Under acidic (pH 2, 4) conditions Cr occurs mainly as Cr+3

(84%) and CrSO4+ (62%) while in the alkaline (pH 10) the

Cr occurs as soluble hydroxide chromium (i.e. Cr(OH)3(aq))

(Table 9). The XRD pattern of the lagoonal sediment B1

indicated that at different pH values the content of specific

phases changed due to pH changes in solution. The most

important dissolution taking place in this sample is the

partial dissolution of clays (especially clinochlore) along

all the pH range investigated and the formation of chromite

ore in the pH 8 and 10. The signature of gypsum in this

sample is rather low due to the dominant signature of

quartz, which is weakly dissolved (Fig. 6a).

In sample B3, the pH changes during the pHstat batch

tests are also reflected in the mineralogical composition

changes through the dissolution and formation of secondary

precipitates. In this sample, the partial dissolution of gyp-

sum occurs in the complete range of the studied pH values

as indicated in Fig. 6b, however, most distinctive dissolu-

tion occurs in the alkaline environment as shown also by

Visual MINTEQ 2.40 (Table 9). The dissolved sulfates

occur predominantly as SO42- (i.e. 20% at pH 2 until up to

80% at pH 10), or as CaSO4(aq) predominantly in the acidic

(pH 2; 42%) and almost neutral environments (pH 6; 40%).

From the mineralogical data it is clear that the Cr(VI)

occurs associated mainly with sodium (sodium chromate

and sodium dichromate). However, the dominant soluble

species differ depending on the pH. In acidic environments

(pH = 3.8), the hydrogenchromate (HCrO4-) is the pre-

dominant soluble specie accounting for up to 82% of the

total soluble hexavalent Cr, while in the alkaline environ-

ments the chromate anion (CrO42-) is the dominant form

increasing from 23?38?61% if the pH is increasing

from 6?8?10, respectively. The soluble chromatite

(CaCrO4(aq)) appears to play an important role in the alka-

line environment accounting for 39% (pH 10) and up to

62% (pH 8). Barium solubility depends on the presence of

Table 9 The content (in %) of the main species based on the geochemical modeling with Visual MINTEQ 2.40 of the different soluble phases

leached during the pHstat experiments

Components Species name Sample

B1 B3 F3 F4

pH 2 4 6 7.7 8 10 2 3.8 6 8 10 3.9 6 8 2.7 6 8

Cr(III) Cr3+ 83.8 23.0 47.1 16.3 16.5

Cr(OH)3 (aq) 95.3 95.7 96.0

Cr(OH)4- 4.6 5.1

CrOH2+ 0.4 9.8 0.2 6.1

CrSO4+ 15.9 61.4 52.7 72.3 82.9

CrOHSO4 (aq) 4.8

Cr2(OH)24+ 94.8

Cr(VI) CrO42- 0.4 22.8 37.8 61.0

HCrO4- 71.0 82.4 28.7 0.5

Cr2O72- 22.9 15.9 2.9

CaCrO4 (aq) 0.6 45.5 61.5 38.7

CrO3SO42- 5.9 0.7

Sulfates SO42- 9.8 49.6 54.3 59.3 50.3 96.9 19.5 64.7 54.8 60.4 80.2 37.3 62.4 73.7 33.8 58.3 62.8

CaSO4 (aq) 36.9 49.5 44.4 31.8 49.0 3.0 41.6 34.9 40.7 36.4 18.8 45.0 2.5 19.6 37.6 27.8 25.2

MgSO4 (aq) 11.4 10.7 6.0 10.4 12.4 10.2

AlSO4+ 38.7

Barite Ba+2 36.5 29.8 3.6 28.1 16.0 31.6 26.3

BaSO4 (aq) 63.0 69.7 96.3 71.9 83.9 68.4 73.7

Environ Geol

123

barite (BaSO4) and hashemite (BaCrO4) in the analyzed

sample and the isomorphic substitution of the SO42- for

CrO42- as well as on the dissolution/recrystallization cycles

of these mineralogical phases. From Visual MINTEQ 2.40

calculations (Table 9), the solubility of Cr(III) appears to be

important at (pH 2) occurring mainly as Cr+3 (47%) or as

CrSO4+ (53%). The source of Cr(III) in solution appears to

be related to the presence of chromite ores.

In acidic environments, the samples F3 and F4, Cr is

mainly (up to 83%) occurring as CrSO4+, while in alkaline

environments the dominant specie is Cr(OH)3(aq). The

solubility of sulfates differs in the pre-selected samples by

dissolving maximum quantities in the alkaline (sample F3)

and acidic (sample F4) environments (Table 9). The clays

appear to be unaffected by the different pH values, while

the content of anhydrite is likely to be affected by the pH.

Its peaks increase their intensity at pH values 6 and 8 in

sample F4 and only in pH 6 in sample F3 (Fig. 6c,d).

However, due to the low detection limit of the XRD

analysis (Fig. 6) and the low content of the chromium, the

solubility of the main Cr-species is difficult to be observed.

Thus, the use of the Visual MINTEQ 2.40 highlights the

mineralogical sources, which occur even in small quantities

as well as on the speciation of the soluble phases.

Implications for the on-site leaching

Worldwide there are a wide range of methods (especially

leaching tests) addressing the same question: how much

does a waste material leach under normal or simulated

conditions? For a better understanding of the leaching

behavior, the information on the behavior of the major

elements (like Fe, Al, Ca, etc.) is crucial as they dictate the

leaching environment for the trace contaminants (van der

Sloot et al. 2006). Thus, the use of a differentiated

approach to study the sites highly contaminated with

COPR-wastes has shown how combination of geochemical

analysis, leaching tests, mineralogical characterization,

mathematical and geochemical modeling can provide

valuable insights into the processes occurring on-site or

under simulated pH conditions. The modeling of the sol-

uble species in combination with the mineralogical

determinations have been used to provide a qualitative

assessment of the mineral phases present and the deter-

mination of the different contributors with respect to

dissolution and precipitation in the complex materials such

as COPR.

Mineral solubility and pH have been identified as key

factors in the retention and the release of the Cr(VI)O42-

Fig. 6 XRD patterns of the samples B1 (a), B3 (b), F3 (c) and F4 (d) before and after the pHstat tests with the respective pH values as indicated.

The XRD patterns from the pHstat tests are off set for clarity

Environ Geol

123

and Cr(VI)2O7- from the COPR-rich sediments. The

release of elements (except Cr(VI)), increases by

decreasing the pH. The key mineral phases responsible

for the fast release of the Cr are the chromate salts (i.e.

sodium chromate and sodium dichromate), while the

long-term retention and comparatively slow but still

considerable release of Cr(VI) are chromatite (CaCrO4)

and hashemite (BaCrO4). In the case of Cr(VI), the

decrease in the Cr(VI) concentration in the equilibrium

solution may be possibly due to both direct adsorption of

Cr(VI) on soil/sediment particles and the reduction of the

Cr(VI) to Cr(III) and its subsequent adsorption as shown

in Bolan et al. (2003). Since there was no evidence for

the adsorption of Cr(VI) by the soil, the decrease in the

concentration of Cr(VI) (sample B3) in the acidic envi-

ronments during the pHstat batch experiment was

attributed mainly to the reduction of Cr(VI) to Cr(III) and

its subsequent adsorption into sediment particles. This is

expressed by rather similar total concentrations of Cr

(Table 2) recorded in the lagoonal and pond sediments

(samples B1 and B3, respectively).

The solubility of Cr in solution is highly susceptible

to changes in pH. There is a higher release of Cr(III) in

the pH value 2, but there is almost no release of Cr(III)

in the alkaline environments. The low Cr(III) concen-

trations occurring in the natural to slight alkaline

environments were attributed mainly to the adsorption–

precipitation of Cr(III) because there was no evidence for

the oxidation of Cr(III) to Cr(VI). The oxidation of

Cr(III) to Cr(VI) has been reported to occur in soils and

the presence of the high valence Mn oxides is a pre-

requisite, to act as an electron acceptor for the reaction

to occur (Bartlett and James 1979; Fendorf and Zasoski

1992). Since the sediments in this study contained very

small amounts of Mn oxides, the oxidation of Cr(III) is

unlikely to occur.

The strongly buffered lagoonal clays (sample B1)

appear to be a significant natural barrier for the further

migration of the COPR-pollutants in depth. Moreover, due

to the high buffering capacity of the COPR-rich sediments,

considerable changes of the pH are, on a time scale of

decades, unlikely to occur under the existing field

conditions.

Conclusions

Based on the investigated lagoonal and pond sediments, the

use of pHstat leaching tests, the characterization of solid-

state by means of XRD and soluble-state by means of

geochemical modeling is possible to depict the following

conclusions:

– In general, the solubility of heavy metals is highly

susceptible to changes in pH (increase by decreasing

pH). Apart from pH, for Cr also speciation plays a

significant role in its solubility.

– The strongly buffered lagoonal clays (sample B1) are a

strong natural barrier for the further migration of the

pollutants (i.e. pond sediments and COPR-wastes) in

depth due to their low permeability. Moreover, due to

the high buffering capacity of these lagoonal sediments,

considerable changes of the pH are, on a time scale of

decades, unlikely to occur under the existing field

conditions as shown by the results from the sediment

acidification.

– However, if extreme acidic conditions (pH = 2) occur

in the lagoonal sediments (sample B1) at pH \ 4 will

enhance the leaching of only mainly Cr(III) species.

The release of Cr(III) in acidic conditions corresponds

to some extend to the dissolution of the iron oxides and

clays (Fe, Al, Ca and K), which are known to play an

important role in the adsorption of Cr precipitates

especially in the form of Cr(III) hydroxides.

– The release of the Cr(VI), Cr(III), SO42-, Ca, Mg, Fe

and other elements under natural or pre-defined pH

conditions, strongly depends on the dissolution of solid

phases from COPR-wastes.

– The limning of the Cr(VI)-rich pond sediments (sample

B3) will result in a higher concentrations of the Cr(VI)

soluble species, while their acidification will minimize

the solubility of the Cr(VI) species by four times. The

results from the pHstat batch leaching tests and

geochemical modeling suggest that leaching of Cr(VI)

can be increased by unsuitable treatments (such as

uncontrolled liming of the pond sediments).

– On contrary, the limning of the Cr(III)-rich pond

sediments (samples F3 and F4) will result in immobi-

lization of Cr(III) species in solution for all the duration

of the pHstat experiment.

– Mineral solubility has been identified as key factor in

the retention and release of the different Cr species

from the pond and lagoonal sediments. However, due

to the low detection limit of the XRD analysis the

solubility of different Cr containing species is diffi-

cult to be recorded accurately, thus the use of the

Visual MINTEQ 2.40 appears to help in the inves-

tigation of species partition, which occur even in

small quantities.

– The site-specific conditions of Porto-Romano play an

important role in the speciation/fractionation of dis-

solved different Cr-species. Thus, the mobility of the

different Cr species (i.e. Cr(VI), Cr(III)) and pH shall

be taken into account in the decision to remediate or

isolate this site.

Environ Geol

123

– Moreover, the findings reported here along with a

comprehensive presentation of Cr speciation and other

relevant data for these COPR-contaminated sites, may

also be significant with respect to the Cr behaviour and

the remediation of Cr-contaminated sites elsewhere in

the world.

Thus, improving the knowledge with respect to processes

controlling the release of Cr, the chemical speciation can

be addressed even in the most complex and heterogeneous

waste materials as is the case for COPR.

Acknowledgments This research is realized with the financial

support of the Katholieke Universiteit of Leuven, East European

Initiatives Project 3E000659.

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