Natural Attenuation Potential of Cyanide in Groundwater Near a SPL ...

Natural Attenuation Potential of Cyanide in Groundwater Near a SPL
Landfill
I.Gagnon
Department of Chemical Engineering, École Polytechnique de Montréal, Montreal, Quebec, Canada
G.J.Zagury
Department of Civil, Geological and Mining Engineering, École Polytechnique de Montréal, Montreal,
Quebec, Canada
L.Deschênes
Department of Chemical Engineering, École Polytechnique de Montréal, Montreal, Quebec, Canada
ABSTRACT: Spent pot lining (SPL), a waste product containing cyanide, has been deposited in a landfill at
an aluminum refining site in British Columbia (Canada). Since the capping of the landfill, a continuous decline
of the cyanide level in the groundwater was observed for more than a decade, suggesting that natural
attenuation might be taking place. The information gathered from a critical review of the existing literature
along with available groundwater characterization data identified adsorption and biodegradation as two possible
attenuation phenomena. Collection and analysis of additional groundwater samples showed that the majority
of the cyanide in the groundwater was under the form of iron-cyanide complexes. Using the groundwater
characterization data, geochemical modeling using Visual MINTEQ was performed in order to predict the
concentration of dissolved cyanide species. The results were close to the field data and they confirmed the occurrence
of significant concentrations of Fe(CN) 6
4-
and Na Fe(CN) 6 3- along the contamination plume.
1 INTRODUCTION
1.1 Background
Cyanide occurs as a groundwater contaminant at
various current and former industrial sites, including
electroplating facilities, aluminum production plants,
manufactured-gas plants (MGP), and gold mining
industries. The study site is located in British Columbia
(Canada) and is primarily an aluminum
smelter, but also functions as an aluminum production
facility.
Spent pot lining (SPL) is a waste product of aluminum
refining. It contains, in addition to carbon,
fluoride salts and cyanide. In British Columbia, SPL
is classified as a special waste by the provincial
Ministry of Water, Lands and Air Protection
(MWLAP), and is therefore subject to particular
handling, storage and transfer requirements. Disposal
of SPL is an industry wide problem.
Prior to the mid-1980s, SPL from the smelter was
deposited in a landfill on the smelter site. The landfill
was capped at the end of the 1980s to prevent
leaching of cyanide to groundwater. Over the next
decade, a continuous decline of the cyanide level in
the groundwater was observed, suggesting that natural
attenuation (NA) might be taking place. The
main objectives of this study are: (1) to identify the
main NA phenomena which could explain the cyanide
decrease ; and (2) to assess the NA potential of
cyanide in groundwater near the landfill.
1.2 Cyanide chemistry
Speciation of the cyanide present in groundwater is
very important when determining groundwater toxicity.
Cyanides are toxic and their toxicity is related
to their physicochemical speciation (Zagury et al.
2004). Cyanide can exist in aqueous solution as free
cyanide (HCN, CN - ) or as complexes with metals
such as cadmium, copper, iron, gold and nickel
among others, or as thiocyanate. Free cyanide, and
weak acid dissociable cyanide (CN WAD ) (complexes
with metals such as copper, zinc, nickel) are classified
as the most toxic because of their high metabolic
inhibition potential whereas strong acid dissociable
cyanide (CN SAD ) (complexes with cobalt,
iron, gold) are considered to be relatively less toxic
(Shifrin et al.1996; Zagury et al. 2004). Thiocyanates
are not considered to be very toxic when compared
with cyanide.
Available information on chemical speciation of
cyanide in contaminated groundwater indicates that
iron-cyanide complexes are often dominant (Dzombak
et al. 1996; Meeussen et al. 1992; Theis et al.
1994). These species are stable in the dark at neutral
to high pH (Meeussen et al. 1992) and are resistant
to biodegradation (Aronstein et al. 1994). However,
Meehan et al. (1999) have shown that cyanide can
be degraded in aerobic and anaerobic conditions and
Cherryholmes et al. (1985) observed that K 3 Fe(CN) 6
can be degraded by P. aeruginosa in the dark. In addition,
at neutral pH conditions, iron cyanide species

exhibit little adsorption onto iron oxides (Theis &
West 1986) but perhaps have greater adsorption on
aluminum oxides and kaolin clay (Alesii & Fuller
1976), which can be important adsorbents in
groundwater systems. Ghosh et al. (1999a) investigated
fate and transport of cyanide at a manufactured-gas
plant (MGP) site and their results showed
that cyanide in groundwater was primarily in the
form of iron-cyanide complexes which are transported
as non-reactive solutes in the sand-gravel aquifer
material. Also, chemical decomposition of iron
cyanide complexes can occur, but the decomposition
kinetics are extremely slow.
Free cyanide has been shown to react with various
forms of sulphur in the environment to form thiocyanate.
The two forms of sulphur most likely to react
with cyanide are polysulphides (S 2 ) and thiosulphate
(S 2 O 3 ) (Smith & Mudder 1991). They react
according to the following equations:
S x -2 + CN - → S (x-1) -2 + SCN - (1)
S 2 O 3 -2 + CN - → SO 3 -2 + SCN - (2)
In neutral to basic solutions, both polysulphides and
thiosulphate are oxidation products of sulphides. As
such, these products could possibly be present in
oxidizing environments, such as the vadose zone in
soils. The concentrations of polysulphides and thiosulphate
in groundwater are strongly dependent on
the sulphur content and the Eh-pH conditions in that
groundwater.
1.3 Natural attenuation
Natural attenuation phenomena such as volatilization,
chelation and precipitation, chemical decomposition,
adsorption, photolytic reactions, and biodegradation
with indigenous microorganisms can occur
in various environments. In groundwater, volatilization
and photolytic reactions are less likely to occur
and chemical decomposition has very slow kinetics.
1.3.1 Adsorption
A key point in the case of cyanide is that most aqueous
species are anionic (except HCN), and will
therefore only be sorbed on soils with high anion
exchange capacity. The surfaces of most soil particles
have low anion exchange capacity. Retardation
of cyanide release due to sorption processes is generally
of minor importance in most soils (Kjeldsen
1999). However, some surfaces such as Fe, Al and
Mn oxides, clay minerals and organic matter may
provide anion exchange sites (McLean & Bledsoe
1992).
1.3.2 Precipitation
Most of the cyanide at aluminum plant sites is originally
in solid form as spent pot lining, which contains
primarily solid cyanide complexes and thiocyanate
compounds. Thiocyanates are quite soluble
in water and can produce relatively high thiocyanate
concentrations in pores and groundwater. Iron cyanide
complexes are generally slightly soluble. The
potassium and sodium ferrocyanides are readily
soluble, but can reprecipitate in the presence of iron
(III). The solubility of iron cyanide in soil is dependent
on pH. Thus, re-precipitation of cyanide can
occur if conditions change. The potential for the
cyanide precipitation process to occur can be evaluated
with the use of computer-based chemical speciation
models. In this study, modeling of thermodynamic
equilibrium conditions in groundwater
samples collected from four wells will be performed
using Visual MINTEQ software .
Ghosh et al. (1999b) showed that precipitation of
iron cyanide complexes using a reactive barrier has
the potential to be used for in situ treatment of cyanide-contaminated
groundwater.
1.3.3 Microbial degradation
There are many examples of cyanide microbial degradation
under both aerobic and anaerobic conditions
(Wang et al. 1996; Barclay et al. 1997). Simple
cyanides, in particular, are relatively easily degraded,
especially under aerobic conditions. Degradation
of iron cyanides under aerobic conditions also
occurs, but at a slower rate than that of simple cyanide
degradation. To our knowledge, bacterial degradation
of iron cyanides in groundwater under anaerobic
conditions has not yet been investigated.
Aerobic degradation of CN- is mediated by a
number of bacteria, fungi, algae, and yeast (Young
& Jordan 1995). The first step involves oxidation of
cyanide to cyanate:
CN - + ½ O 2 (aq) → OCN - (3)
This is followed by hydrolysis of the cyanate, which
requires a pH below 7:
OCN - + 3 H 2 O → NH 4 + + HCO 3 - + OH - (4)
Aerobic degradation of CN SAD is mediated by a
number of organisms, generally Pseudomonas bacteria
(Young & Jordan 1995). The relevant reaction is:
M(CN) x y-x + 3xH 2 O + x/2 O 2 (aq) → M y+ + x NH 4 + +
x HCO 3 - + xOH - (5)
Thiocyanate is also subject to bacterially-mediated
aerobic degradation:
SCN - + 3H 2 O + 2 O 2 (aq) → SO 4 2- + NH 4 + + HCO 3
-
+ H + (6)
Anaerobic biodegradation of cyanide and hydrogen
cyanide is restricted to the moderately to strongly
reduced portions of the environment and can only

occur if HS - or H 2 S(aq) are present. The sulphur
species present will depend on pH. At a pH value
greater than 7, HS - is the dominant species. At a
lower pH, H 2 S(aq) will be present. These equations
illustrate the anaerobic biodegradation of cyanide:
CN - + H 2 S (aq) → HCNS + H + (7)
HCN + HS - → HCNS + H + (8)
The HCNS will be hydrolyze to form NH 3 , H 2 S and
CO 2 . In comparison with the aerobic biodegradation
of cyanide, anaerobic degradation is much slower
and anaerobic bacteria have a cyanide toxicity
threshold of only 2 mg/L compared to 200 mg/L for
aerobic bacteria (Smith & Mudder 1991). Consequently
anaerobic biodegradation would be a less
effective cyanide removal mechanism.
Presumably, natural attenuation may play an important
role in cyanide reduction in groundwater.
According to Kjeldsen (1999), studies on natural attenuation
of cyanide compounds in groundwater at
relevant contaminated sites must be carried out.
Few investigations have been conducted in a realistic
soil/groundwater environment. The presence
of another substrate, or nutrients, as well as the accessibility
of cyanide for the bacteria, can vary
greatly in a natural soil/groundwater environment.
Hence, the interest in conducting experiments under
field conditions using indigenous material is evident.
2 CHARACTERIZATION OF WATER
SAMPLES
2.1 Sampling
Groundwater from the study site was sampled at
four different locations (wells PN2A, PN3B,
MW12A, and PZ3) in May 2002 and May 2003 in
order to assess the potential for natural attenuation
near source, middle, end, and just off the contamination
plume.
2.2 Groundwater characterization results
Samples from each of the four wells were analyzed
for different physicochemical characteristics and for
various cyanide species according to Standard
Methods (Clesceri et al. 1998). The most significant
results obtained in May 2003 are presented in Table
1. Results for the four different wells are presented
in the same order as the flow direction. This allows
to follow the evolution of the contamination within
the plume.
Table 1. Physicochemical properties and concentration of cyanide species in groundwater (PN2A, PN3B,
MW12A, and PZ3).
FLOW DIRECTION
Well location PN2A PN3 B MW12A PZ3
2003-05- 2003-05- 2003-05-
Sampling date
12
12
12
2003-05-
12
Units
Dissolved oxygen (in-situ) % 0.85 1.09 0.63 1.52
pH (in-situ) 9.48 8.84 8.91 6.70
Redox potential (in-situ measurement)
MV -72 123 -188 -36
On-site temperature °C 9.3 13.4 9.3 11.2
Dissolved anions
Sulphate mg/l 130 15 32 678
Cyanides
Total cyanide mg/l 4.8 0.056 2.6 0.032
Free cyanide mg/l

Total cyanide values (CN Total ) include free cyanide,
weak acid dissociable cyanide (CN WAD ) and strong
acid dissociable cyanide (CN SAD ). To obtain the
concentration of strong metal-cyanide complexes,
free cyanide and CN WAD concentrations must be
substracted from the total cyanide. It should be
noted that thiocyanates are not part of the total cyanide
concentration.
2.3 Discussion of groundwater characterization
Using the data presented in Table 1 and the conclusions
drawn from the literature review, one can foresee
which cyanide species should be present in the
groundwater as well as which attenuation mechanisms
are likely to occur.
Site conditions and species distribution can play a
major role in the adsorption/precipitation potential.
For example, pH influences the sorption potential by
creating negatively or positively charged surfaces
and redox potential determines the oxidation state of
some species (ex. Fe 2+ / Fe 3+ ). Here are some observations
made while analyzing the data obtained from
the groundwater laboratory characterization:
- CN Total content ranges from 0.03 mg/l
(off the contamination plume) to 4.8 mg/l
(near source).
- In all wells, CN Total content is far greater
than that obtained for CN WAD , indicating
that the majority of cyanide in groundwater
is under the form of strong metal
complexes;
- Of the four main metals forming strong
metal complexes with cyanide (Fe, Co,
Ag, Au), only Fe is present in significant
concentration;
- Two of the four wells (PN2A and
MW12A) show a high concentration of
thiocyanates. These two wells are the
ones having the greatest quantity of dissolved
organic carbon and the lowest dissolved
oxygen (DO) concentration. Furthermore,
the measured redox potential
was negative in these wells;
- Groundwater pH is alkaline in the plume
and is near neutrality off the plume
(PZ3).
According to the observations stated above and the
detailed analysis, it is possible to assume that the
majority of the cyanide (excluding thiocyanates)
present in the groundwater at the four wells is under
the form of strong metal complexes because the free
cyanide and weak-acid dissociable cyanide concentrations
are very low. Also, dissolved metals analysis
show that these strong complexes are more likely
to form with iron.
High concentrations of thiocyanate at wells
MW12A and PN2A can be explained by different
factors. These wells show the highest concentrations
of dissolved sulphate and dissolved organic
carbon, and low concentrations of dissolved oxygen.
All these characteristics suggest that sulphate reducing
conditions prevail in the groundwater in both
wells MW12 and PN2A. Furthermore, dissolved organic
carbon can provide adsorption sites and can
serve as a substrate for bacterial activity.
Groundwater characteristics will directly influence
microbial activity. As mentioned in the literature
review, cyanide biological degradation pathways
differ depending on redox conditions; aerobic
biodegradation occurring at a much faster rate than
its counterpart. Here are some additional observations
(in relation to bacterial activity) made while
analyzing the data obtained from the groundwater
geochemical characterization:
- Dissolved oxygen concentration in the
plume is very low and vary between
0.6 and 1.1 % of saturation;
- The well off the plume (PZ3) showed a
relatively elevated DO concentration (3.5
% in May 2002).
It will be interesting to study the biological degradation
of iron cyanide complexes in the groundwater
because this is the dominant cyanide species. The
wells having the lowest redox potentials in 2003
(MW12A and PN2A) show elevated thiocyanate and
sulphate concentrations along with high DOC. These
conditions are very suitable for sulfate-reducing
bacteria activity in groundwater.
3 GEOCHEMICAL MODELING USING
VMINTEQ CODE
Using the results of the detailed groundwater characterization,
geochemical modeling of the groundwater
composition was performed with Visual
MINTEQ. This software is used to predict the concentrations
of various dissolved species in the
groundwater once the thermodynamic equilibrium is
reached. It gives the saturation status, providing information
on the probability of precipitation.
The modeling was performed using May 2002
data. According to the Visual MINTEQ modeling
results, the dominant cyanide species in the groundwater
at each of the four wells once thermodynamic
equilibrium has been reached are the following:
MW12A - Thiocyanates (SCN - ); PN3B - Free cya-

nide (HCN) and iron cyanide complexes (Fe(CN) 6
4-
and Na Fe(CN) 6 3- ); PN2A - Thiocyanates (SCN - );
PZ3 - Free cyanide (HCN).
In accordance with the results obtained from the
geochemical characterization, the dominant species
is thiocyanates in wells MW12A and PN2A. It
should be noted that in these wells, the dominant
cyanide species (other than thiocyanates) are
Fe(CN) 6
4-
and Na Fe(CN) 6 3- ). Thiocyanates are quite
soluble in water and can produce relatively high
thiocyanate concentrations in pores and groundwater
(Kjeldsen 1999), as observed in wells PN2A and
MW12A. PZ3, for its part, shows free cyanide
(HCN) as its dominant cyanide species which can be
explained by the groundwater pH of 6.2. At such a
pH, more than 99.9 % of the free cyanide is under
the form of molecular hydrogen cyanide (the pKa is
9.2 at 25 o C).
To help compare the modeling results with the
characterization results obtained in the laboratory,
Table 2 shows the concentration of various cyanide
species using the two methods.
The concentrations provided by the modeling are
close to those measured in the field. The total cyanide
concentrations calculated using the equilibrium
modeling results are slightly higher than those obtained
during the 2002 geochemical characterization.
This result allows the assumption that the reactions
in the groundwater under the SPL landfill may have
reached a pseudo equilibrium. However, as site conditions
change (ex. pH, redox), equilibrium will shift
and cyanide species concentrations might be different.
The modeling does not include the influence of
reaction kinetics and the occurrence of phenomena
such as sorption and microbial degradation.
Furthermore, the modeling results indicate that in
the contamination plume, iron cyanide complexes
and specifically Fe(CN) 6
4-
and NaFe(CN) 6 3- are the
dominant cyanide species. The geochemical modeling
of the groundwater composition therefore completes
and refines the groundwater characterization
performed in the laboratory.
4 CONCLUSIONS AND FUTURE WORK
Characterization of groundwater samples collected
at four different locations (near source, middle,
end, and just off the contamination plume) in the
Spring of 2002 and 2003 showed total cyanide concentrations
(CN Total ) ranging from 0.02 mg l -1 (off
the plume) to 8.9 mg l -1 (near the source of the
plume). In all wells, CN Total was far greater than
weak acid dissociable cyanide (CN WAD ) suggesting
that the majority of the cyanide in the groundwater
was under the form of iron-cyanide complexes.
However, the more toxic free cyanide species (HCN,
CN - ) were below detection. Two out of the four
wells steadily showed high concentrations of thiocyanate
(SCN - ). These wells had conditions very
suitable for sulfate-reducing bacteria activity (high
dissolved sulphate and dissolved organic carbon
concentrations, and low concentrations of dissolved
oxygen).
Using the groundwater characterization data,
geochemical modeling with Visual MINTEQ was
performed in order to predict the concentration of
dissolved cyanide species at thermodynamic equilibrium.
The results were close to the field data and
they confirmed the occurrence of significant concentrations
of Fe(CN) 6 and NaFe(CN) 4- 6 along the
4-
contamination plume. Further modeling will be
performed with the 2003 data, to better understand
the fate of cyanide.
The results of the groundwater characterization
and the geochemical equilibrium modeling enabled
us to identify the main natural attenuation phenomena
which should be tested in the laboratory. Adsorption
and microbial degradation will be investigated.
The adsorption potential will be assessed using
batch and column experiments. Available geological
material from the site will be used along with
groundwater from the site and synthetic cyanidecontaminated
water.
Heterotrophic and cyanide-resistant bacteria
counts will be performed to investigate the presence
of indigenous microflora in groundwater. Microbial
degradation of iron cyanide complexes and free cyanide
will be evaluated using radiolabeled mineralization
tests in microcosms (Oudjehani et al. 2002).
Upon completion of these experiments, we should
have a better understanding of cyanide fate and natural
attenuation potential in the groundwater near
spent pot lining landfills.
Table 2. Total cyanide, strong acid dissociable cyanide and SCN - in groundwater (mg/l) following laboratory
characterization and thermodynamic modeling.

* Calculated as the difference between total cyanide and WAD cyanide
** Calculated by summing every WAD, SAD and free cyanide species concentration
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Sampling Wells Laboratory characterization (May 2002
sampling campaign)
Modeling results (Visual MINTEQ)
Total CN *CN SAD SCN - **Total CN Iron-cyanide SCN -
Complexes
PN2A 8.9 8.83 185 12.7 12.7 184.78
PN3B 0.584 0.576 < 5 0.696 0.512 5.8E-18
MW12A 3.2 3.176 17 4.33 2.37 16.98
PZ3 0.02 0.02 < 0.5 0.021 < 10 -5 5.8E-18