# Assessment of Mercury-Polluted Soils Adjacent to  an Old Mercury-Fulminate Production Plant

Authors: M. Camps Arbestain; L. Rodríguez-Lado; M. Bao; F. Macías
Journal: Applied and Environmental Soil Science
 (2009)

---

## Abstract

Mercury contamination of soils and vegetation close to an abandoned Hg-fulminate production plant was investigated. Maximum concentrations of Hg (>6.5 gkg−1 soil) were found in the soils located in the area where the wastewater produced during the washing procedures carried out at the production plant used to be discharged. A few meters away from the discharge area, Hg concentrations decreased to levels ranging between 1 and 5 g kg−1, whereas about 0.5 ha of the surrounding soil to the NE (following the dominant surface flow direction) contained between 0.1 and 1 g kg−1. Mercury contamination of soils was attributed (in addition to spills from Hg containers) to (i) Hg volatilization with subsequent condensation in cooler areas of the production plant and in the surrounding forest stands, and (ii) movement of water either by lateral subsurface flow through the contaminated soils or by heavy runoff to surface waters.


---

## Body

## 1. Introduction

Mercury is often found
in soils as “hot spots” located close to industrial facilities that either use
Hg in their fabrication processes   (e.g., chlor-alkali plants) or produce Hg
compounds (e.g., Hg-fulminate plants). The type of reactions that take place
during the production process,  as well as during transportation and disposal,
largely determines the chemical composition and distribution of Hg in the
surrounding environment [1]. Mercury-fulminate (Hg(OCN)2)  used to be
produced as a primary explosive for percussion caps and as a detonator [2].
Formation of this detonating compound involves the dissolution of Hg in nitric
acid and the addition of ethanol. Acid vapors containing ethanol and Hg are
generated during this process, although they were usually condensed and
collected within the production facilities. Wastewaters produced—either after
filtering the reacting mixtures or through washing activities—were
historically disposed of in the surroundings of the production plants. This explains
why the soils surrounding many of these old facilities contain high levels of
Hg contamination.Mercury can undergo
changes in speciation that are either physicochemically or biologically induced,
which results in changes in solubility, toxicity, and bioavailability [3].
Thus, the weathering of Hg materials disposed in soils may redistribute Hg in
other chemical forms and facilitate its dispersal in watersheds or atmospheric
emissions [4]. This further complicates the characterization of these contaminated
sites, which is already complex because of the very heterogeneous distribution
of this type of pollutant in the environment and within samples. Moreover, the
sampling of soils contaminated with primary explosives, such as Hg-fulminate,
is risky because of the extreme instability of these compounds [5].Mercury is naturally present in
soils at concentrations  ranging between
0.003 and4.6mgkg−1 [6]—in most cases
below 0.5mgkg−1 [7]—whereas in
contaminated sites, concentrations of up to 11500 and 14000mgkg−1 have been reported [8, 9]. In these contaminated areas—where Hg entrance
to the system is mainly via surface spills, wastewater discharge, and/or by
condensation of atmospheric Hg—the element tends
to accumulate in the soil surface horizons, and is mainly retained by sorption
onto organic compounds and, to a lesser extent, clays [3, 10]. Maximum
sorption onto soil organic surfaces occurs in the range of pH 3 to 5 [11, 12],
whereas as pH increases, sorption decreases, mainly because of the increase in
dissolved organic matter complexed with Hg [12]. Thioligands appear to be
mainly responsible for Hg binding to organic compounds [13] and, in general,
organic matter exerts a dominant influence on Hg binding, transformation, and
transport processes [14]. Other factors affecting Hg retention in surface
soils, in addition to organic matter, are (i) chemical properties, such as soil
pH and redox potential, which affect Hg speciation and solubility [15], (ii)
amount and type of mineral colloids [16], (iii) presence of Cl− ligands [12, 17], and (iv) soil temperature.In the present study,
Hg contamination of soils and vegetation in the surroundings of an abandoned
Hg-fulminate production plant was investigated. Digital maps of the
distribution of Hg in the soils in the study area were generated for the
different depths studied. Distribution of Hg in different particle-size
fractions was also investigated. Additionally, the geochemical evolutionary
trends of Hg in the contaminated soils were estimated from Eh and pH
determinations.


## 2. Materials and Methods

### 2.1. Site History

The site under study (see Figure1) is
located 6 km from the city of Oviedo (Asturias, North West Spain), and has an extension of 90 ha. The mean annual
temperature in the area is 12°C, and total annual precipitation is 1100 mm. Soils are
classified as “Urbi-anthropic Regosols” [18]. The natural soils in nearby areas
are Umbrisols developed from poorly developed metamorphic rocks. The plant
began operations in 1866, although since then, the type of products
manufactured has changed greatly. Since the plant became operational, a number
of products have been manufactured, including sulphuric acid, nitroglycerine,  nitroglycol, dynamite, dinitrotoluene, thrilite,  and emulsions, Ca superphosphates,
Hg-fulminate, and BNT-DNT. Production at the plant ceased in 1996, and the
facilities are currently used for the storage of commercial explosives produced
in other plants. Within the study site, the former Hg-fulminate production plant
is located on a low hill (220–240 m height) in
the NE of the property; the site covers an area of 4.3 ha, which is dominated
by a dense deciduous forest. The Hg-fulminate production facilities occupy an
area of 840m2.
In addition to this primary explosive, other materials, mainly penthrite (PETN)
and TNT, used
to be stored in the area.Figure 1
View of the study area (source: Google Earth).


### 2.2. Sampling and Sample Preparation

A total of 37 sampling
points (28 within the area of Hg-fulminate production and 9 in the surrounding area)
were sampled taking into account the position of possible sources of Hg
contamination (e.g., areas of storage, production, discharge, etc.) as well as
the possible sinks. Soil samples were collected from different depths, down to
the presence of a compacted layer (e.g., a rock, clay sediments, or concrete),
and a total of 127 soil samples were analyzed for Hg. All soils were found to
be highly disturbed by the construction of the explosive production facilities.
Soils were air-dried, thoroughly mixed, and ground to pass through a 2 mm sieve, before use.
Twenty-three of the soil samples were selected for a more detailed analysis. Of
these, Hg-contaminated samples covering the whole pH range of the soils from
the area were chosen. Particle-size fractionation of some soil samples was
carried out by sieving to separate the following fractions: coarse sand (1-2 mm), fine sand (0.2–1 mm), very fine
sand (0.2–0.05 mm), silt +
clay (<0.05 mm).
Organic matter was not removed from the soil samples.


### 2.3. Soil Chemical Analyses

The total Hg concentration was
determined in dry soil samples, with an LECO AMA-254 combustion Hg analyzer (LECO
Corp., St. Joseph, Mich,
USA).
This system
determines Hg directly by combustion, amalgamation, concentration on a gold
filter, and spectrometry. Several certified NIST standards were used (e.g.,
2782 industrial sludge and 1633 trace elements in coal fly ash).Soil samples in which the
concentration of Hg was >10mgkg−1 were diluted with commercial
kaolinite. Comparison of Hg concentrations obtained with and without dilution
with kaolinite showed a good recovery (data not shown). Soil pH was measured in
H2O and KCl in a soil:solution ratio of 1:2.5. The pH of oxidation
was also measured  6 hours after the
addition of 100 mL of  H2O2 to 5 g of soil [19]. Organic C in
the selected soil samples was analyzed by combustion with an LECO carbon
analyzer (model CHN-1000, LECO Corp.) (soil samples of pH  > 5.6 were
previously treated with concentrated HCl to eliminate carbonates for organic C
determination). The redox potential (Eh) of the selected soil samples was
measured in the laboratory as follows. Distilled water was added to the dried
and sieved soil until a saturated paste was achieved; the mixture was then
allowed to dry with the Eh electrode immersed in it. The Eh potential was read
once the soil reached field capacity (24–36 hours later),
when changes in Eh were ≤2mVmin−1. The Eh values obtained are
approximations, as because with this methodology the effects of soil structure
and of many biotic processes on redox potential are overlooked. However,
experiments carried out with A horizons of forest soils from NW Spain showed
differences between field Eh measurements (at field capacity) and laboratory Eh
measurements (following the above described methodology) ≤50 mV (Macías,
unpublished data).


### 2.4. Plant Analyses

Foliar samples ofRubus fruticosus L., Osmunda cinnamomea (fern), and Acer sp. were collected at different
sites around the former production plant, which differed in terms of the Hg
concentrations in the soil. Foliar samples of the three species were also taken
from a noncontaminated site in Galicia,
under similar climatic conditions, but located some 300 km away from the study
area. Foliar samples were washed successively with distilled water, air-dried,
and ground before analyses. The total concentration of Hg was determined in dry
foliar samples with the same LECO AMA-254 combustion Hg analyzer.


### 2.5. Mapping/Kriging

A georeferenced soil database was
constructed using soil sample position and Hg concentration for each soil
layer. The distribution of the maximum Hg concentration in the area was firstly
calculated using ordinary kriging as the spatial interpolator. There was a
single spot with an extremely high Hg concentration (30gkg−1 soil),
which was not included in this process as the contamination was very local, and
this would have distorted the interpolation. Secondly, three levels of risk for
soil Hg concentrations (40, 100, and 1000mgkg−1)   were established
and, for each soil profile, the soil depth at which such values were reached
was determined, and the corresponding maps generated. The maps were overlain on
a digital elevation model so that the influence of topography on the
distribution of Hg in the study area could be inferred. Total concentration of
Hg of 40mgkg−1 corresponds to the threshold value for industrial
areas in several autonomous regions within Spain  (e.g., the Basque Country).


## 2.1. Site History

The site under study (see Figure1) is
located 6 km from the city of Oviedo (Asturias, North West Spain), and has an extension of 90 ha. The mean annual
temperature in the area is 12°C, and total annual precipitation is 1100 mm. Soils are
classified as “Urbi-anthropic Regosols” [18]. The natural soils in nearby areas
are Umbrisols developed from poorly developed metamorphic rocks. The plant
began operations in 1866, although since then, the type of products
manufactured has changed greatly. Since the plant became operational, a number
of products have been manufactured, including sulphuric acid, nitroglycerine,  nitroglycol, dynamite, dinitrotoluene, thrilite,  and emulsions, Ca superphosphates,
Hg-fulminate, and BNT-DNT. Production at the plant ceased in 1996, and the
facilities are currently used for the storage of commercial explosives produced
in other plants. Within the study site, the former Hg-fulminate production plant
is located on a low hill (220–240 m height) in
the NE of the property; the site covers an area of 4.3 ha, which is dominated
by a dense deciduous forest. The Hg-fulminate production facilities occupy an
area of 840m2.
In addition to this primary explosive, other materials, mainly penthrite (PETN)
and TNT, used
to be stored in the area.Figure 1
View of the study area (source: Google Earth).


## 2.2. Sampling and Sample Preparation

A total of 37 sampling
points (28 within the area of Hg-fulminate production and 9 in the surrounding area)
were sampled taking into account the position of possible sources of Hg
contamination (e.g., areas of storage, production, discharge, etc.) as well as
the possible sinks. Soil samples were collected from different depths, down to
the presence of a compacted layer (e.g., a rock, clay sediments, or concrete),
and a total of 127 soil samples were analyzed for Hg. All soils were found to
be highly disturbed by the construction of the explosive production facilities.
Soils were air-dried, thoroughly mixed, and ground to pass through a 2 mm sieve, before use.
Twenty-three of the soil samples were selected for a more detailed analysis. Of
these, Hg-contaminated samples covering the whole pH range of the soils from
the area were chosen. Particle-size fractionation of some soil samples was
carried out by sieving to separate the following fractions: coarse sand (1-2 mm), fine sand (0.2–1 mm), very fine
sand (0.2–0.05 mm), silt +
clay (<0.05 mm).
Organic matter was not removed from the soil samples.


## 2.3. Soil Chemical Analyses

The total Hg concentration was
determined in dry soil samples, with an LECO AMA-254 combustion Hg analyzer (LECO
Corp., St. Joseph, Mich,
USA).
This system
determines Hg directly by combustion, amalgamation, concentration on a gold
filter, and spectrometry. Several certified NIST standards were used (e.g.,
2782 industrial sludge and 1633 trace elements in coal fly ash).Soil samples in which the
concentration of Hg was >10mgkg−1 were diluted with commercial
kaolinite. Comparison of Hg concentrations obtained with and without dilution
with kaolinite showed a good recovery (data not shown). Soil pH was measured in
H2O and KCl in a soil:solution ratio of 1:2.5. The pH of oxidation
was also measured  6 hours after the
addition of 100 mL of  H2O2 to 5 g of soil [19]. Organic C in
the selected soil samples was analyzed by combustion with an LECO carbon
analyzer (model CHN-1000, LECO Corp.) (soil samples of pH  > 5.6 were
previously treated with concentrated HCl to eliminate carbonates for organic C
determination). The redox potential (Eh) of the selected soil samples was
measured in the laboratory as follows. Distilled water was added to the dried
and sieved soil until a saturated paste was achieved; the mixture was then
allowed to dry with the Eh electrode immersed in it. The Eh potential was read
once the soil reached field capacity (24–36 hours later),
when changes in Eh were ≤2mVmin−1. The Eh values obtained are
approximations, as because with this methodology the effects of soil structure
and of many biotic processes on redox potential are overlooked. However,
experiments carried out with A horizons of forest soils from NW Spain showed
differences between field Eh measurements (at field capacity) and laboratory Eh
measurements (following the above described methodology) ≤50 mV (Macías,
unpublished data).


## 2.4. Plant Analyses

Foliar samples ofRubus fruticosus L., Osmunda cinnamomea (fern), and Acer sp. were collected at different
sites around the former production plant, which differed in terms of the Hg
concentrations in the soil. Foliar samples of the three species were also taken
from a noncontaminated site in Galicia,
under similar climatic conditions, but located some 300 km away from the study
area. Foliar samples were washed successively with distilled water, air-dried,
and ground before analyses. The total concentration of Hg was determined in dry
foliar samples with the same LECO AMA-254 combustion Hg analyzer.


## 2.5. Mapping/Kriging

A georeferenced soil database was
constructed using soil sample position and Hg concentration for each soil
layer. The distribution of the maximum Hg concentration in the area was firstly
calculated using ordinary kriging as the spatial interpolator. There was a
single spot with an extremely high Hg concentration (30gkg−1 soil),
which was not included in this process as the contamination was very local, and
this would have distorted the interpolation. Secondly, three levels of risk for
soil Hg concentrations (40, 100, and 1000mgkg−1)   were established
and, for each soil profile, the soil depth at which such values were reached
was determined, and the corresponding maps generated. The maps were overlain on
a digital elevation model so that the influence of topography on the
distribution of Hg in the study area could be inferred. Total concentration of
Hg of 40mgkg−1 corresponds to the threshold value for industrial
areas in several autonomous regions within Spain  (e.g., the Basque Country).


## 3. Results and Discussion

### 3.1. General Soil Properties in the Study Area

The pH of the soils in
the surroundings of the Hg-fulminate facilities varied widely (see Table1).
Soil pH-H2O  values of these samples ranged from 3.9 to 7.8, and
those of pH-KCl from 3.6 to 8.0, whereas natural soils in the area are
moderately acidic (with surface horizons of pH 4-5 and subsurface horizons of
pH 5-6) [20]. The diverse activities carried out in the production plant have
caused changes in the acid-base conditions of the soils. In areas close to
where lime or concrete were applied, pH-H2O values are above 6,
whereas in areas with presence of untreated green pyrite and pyrite cinder
wastes—both of which are
wastes from the production of sulphuric acid—pH-H2O values are below 4. Organic C contents of mineral surface horizons of the
selected soils ranged from 12 to 120gkg−1, whereas those of subsurface horizons
ranged from 4 to 16mgkg−1 (see Table 1). Soils in the surroundings
of the production plant were also found to be contaminated with other heavy
metals in addition to Hg, such as Zn, Cu, Pb, Cd, and As (data not shown),
which are associated with the presence of pyrite cinder wastes, although the
contaminated areas did not always coincide. The present study focuses on the
area within the production plant that is contaminated with Hg.Table 1
Values of pH in water, KCl, andH2O2 (pH of oxidation),
Eh of selected soil samples, organic C content, and Hg concentration of
selected soil samples. Standard errors of Hg concentrations are indicated in
parentheses (n=4).
SiteHorizonDepthpH-H2OpH-KClpH-oxidationEhOrganic CHg conc.(cm)(mV)(gKg−1)(mgkg−1)I-8O5.906.214.09468182.0838 (22)I-8Ah10–106.736.634.9950318.2234 (6)P-6Ah10–106.526.406.4248457.06.96 (0.12)A-2Ah10–156.205.755.3427819.033.6 (1.2)M-2Ah10–107.107.586.4426875.03377 (39)M-3Ah10–107.597.465.9628878.05883 (252)M-4Ah10–57.587.275.93284104.06350 (135)M-5Ah10–57.687.676.4439615.01546 (81)M-6Ah10–207.076.835.8620056.01687 (222)M-9Ah10–305.954.945.3628112.026.4 (0.2)M-10Ah10–56.936.524.8327589.09043 (779)M-11Ah10–205.364.814.0553364.9392 (5)M-12Ah10–204.764.513.20485120.0280 (6)P-11Ah210–207.457.126.2742634.843.7 (1.1)L-4Ah210–204.243.933.7862030.250.7 (3.0)L-4CA70–807.797.446.5045415.8421 (15)P-13C10–204.224.164.205514.2109 (2)P-14C8–154.084.274.455437.4132 (2)P-15C10–185.645.866.135027.2150 (11)P-15C18–434.233.784.155493.9212 (9)P-15C43–933.884.713.935438.035.3 (1.6)P-16C10–303.873.613.326396.127.1 (1.9)P-8C78–2107.388.047.284235.02.46 (0.06)


### 3.2. Mercury Distribution in the Soils of the Study Area

Digital maps of Hg
distribution in  the surface horizons of soils in the area around the
Hg-fulminate facilities were generated (see Figure2). Extremely high levels of
Hg were detected in the discharge area for the wastewater produced during the
washing procedures in the production plant (see Figure 2(a)), with
concentrations higher than 6.5gkg−1 (with a very highly contaminated spot
in which Hg concentration in the first 5 cm depth was 30gkg−1, although
this was not included in the interpolation to avoid distortion of the Hg
concentration gradients). In this highly contaminated spot, elemental Hg was
visually identified as droplets. Some distance away from the trail of runoff
wastewater, Hg concentrations decreased to levels ranging between 1 and 5gkg−1. Mercury
concentrations were also high in the NE vicinity of the production plant, with
values above 0.1gkg−1, covering an extension of ~0.5 ha, in contrast with the
concentration of  0.003mgkg−1 Hg detected in a noncontaminated
parent material in the soils close to the study area. The results thus show a
typical point source distribution pattern, with Hg levels decreasing with
distance from the production plant.Digital maps of Hg distribution in the soils of the area
around the Hg-fulminate facilities. (a) Surface Hg concentrations, (b), (c),
and (d) depths to which Hg concentrations reached values above 1, 0.1, and0.04gkg−1,
respectively.
(a)(b)(c)(d)Mercury is generally of low mobility
because of its high density, which explains the high concentrations in the
vicinity of the disposal site, at the wastewater discharge
area, and some meters downstream. In addition to wastewater discharge and spills
from containers, Hg contamination of soils away from this point may be
attributed to Hg volatilisation—either through
the exothermic reactions of the Hg-fulminate production process or
physicochemical/microbial-induced reactions occurring in contaminated soils—with subsequent
condensation in cooler areas of the production plant and in the surrounding
forest stands (see Section3.5). The subsurface lateral movement of water
contaminated with Hg mineral and organic particles in suspension over the
soils, as well as heavy runoff to surface water, may also be important sources
of the metal downstream (NE direction).The furthest depths, at
which the concentration of Hg reached values above 1, 0.1, and0.04gkg−1, are
indicated in Figures 2(b), 2(c), and 2(d), respectively. In two contiguous
sampling points close to the production plant, Hg concentrations above 1gkg−1 were
observed down to a depth of 40 cm (see Figure 2(b)). In this highly contaminated spot,
concentrations above 0.1gkg−1 were observed even at 1 m depth (see Figure 2(c)). Moreover,
concentrations above 0.04gkg−1 were also observed between 10 and 40 cm depth in the N and NE
directions (see Figure 2(d)). The results obtained thus indicate high
accumulation of Hg in surface horizons, mainly attributed to the repeated entry
of the contaminant to the surface—through spills,
waterflow, or condensation of volatile Hg—and which was
probably retained in the soil by organic matter and to a lesser extent by clay
particles. The presence of Hg in deeper horizons in the sites indicated above
may be related to the downward movement of Hg associated with soluble organic
matter, as previously reported in [12, 17], although more research is needed
to confirm this.


### 3.3. Total Hg in the Particle-Size Subsamples

Comparison was made of
the concentrations of Hg within the different particle size subsamples of
selected soil samples (see Figure3). In general, the results show that Hg was
distributed within all the particle sizes studied, and followed a relatively
homogenous pattern, with a tendency for concentration to increase as the particle
size decreased in the P-11 and P-15 soil samples. Fernández-Martínez et al.
[21] observed a generally higher Hg concentration in the finest particle-size
subsamples, which was attributed to the higher Hg sorption capacity of clay
minerals, Fe and Al oxy-hydroxides, and humus surfaces, all of which tended to
concentrate in the finest grain sizes. Studies carried out to date indicate
that in acid soils (pH < 4.5–5.5) the organic material is the only effective
sorbent for inorganic Hg, whereas in nearly neutral soils (pH > 5.5–6), iron
oxides and clay minerals may become more effective [7, 11, 12]. In this case,
the four samples studied differed greatly in soil pH, organic matter content,
as well as in Hg content (see Table 1), and no relationship was found between
the Hg distribution in these particle sizes fractions and these soil
properties.Concentration of Hg within the different particle-size
subsamples of selected soil samples.
(a)(b)(c)(d)Both elemental Hg andHg2+ tend to be strongly sorbed to the humic
fraction of soils [7], although the former has less affinity for organic matter than Hg2+ species [22, 23]. Moreover, elemental Hg readily vaporizes and can thus be
reemitted into the atmosphere, especially during periods of high temperature
[10]. Under acidic conditions, Hg0 may be oxidized into Hg(I) and
Hg(II) [24], although Hg(I) does not seem to occur as a stable species in soils
[7]. On the other hand, because of the strong affinity of Hg2+ for
humic substances [12, 14, 25] only trace contents of Hg2+ are
generally found in soil solution [26], either as free Hg ions or as soluble Hg
complexes, which are bioavailable. Neither speciation nor sorption processes
were investigated in the present study, although the geochemical evolutionary
trends of Hg in the contaminated soils were inferred from pH-Eh diagrams (see Section 3.4).


### 3.4. Geochemical Evolutionary Trends of Hg in the Contaminated Soils

One of the techniques that can be
used to establish the geochemical evolutionary trends of Hg in the contaminated
soils is the consideration of pH and Eh values of the soil samples, and the
identification of thermodynamically stable Hg species by means of Eh-pH diagrams,
although it must be taken into consideration that these diagrams are simplified
models of very complex systems. The Eh-pH diagram for an  Hg-O-H-S-Cl system is
shown in Figure4 [27], and the Eh and pH values of selected soil
samples from
the study   area are represented. The results obtained (see Figure 4) show that
the group of soil samples with Eh values below 400 mV includes all the soil
samples with Hg concentrations above 1gkg−1, and all correspond to
surface horizons (see Table 1). According to the Eh-pH diagram, Hg0 is the most thermodynamically stable species in the first group of soils, which
is consistent with the fact that these soils were sampled close to the
discharge exit of wastewaters rich in Hg0.Figure 4
Mercury Eh-pH diagram for an Hg-O-H-S-Cl system. Values of Eh
and pH of selected soil samples are displayed. The assumed activities for
dissolved species are Hg:10−8 M, Cl: 10−3.5 M, and S: 10−3 M.On the other hand, the
concentrations of Hg in all soil samples with Eh values>400 mV were below 1gkg−1 (see Figure 4), and according to the Eh-pH diagram, Hg2Cl2, Hg22+, and Hg0 were the most thermodynamically stable species under
the conditions used. However, it is known that Hg(I) has the ability to
disproportionate and equilibrate according to the equation Hg(I) = Hg(0) + Hg(II), with the disproportionation reaction for soils shifted to the extreme
right side, because the high retention of Hg2+ [7]. Thus, Hg(I)
does not appear to occur as a stable species in soil [7]. Finally, within the
latter group of soil samples, the pH of those with Eh values above 550 mV was
below 4.3 (see Figure 4), which reveals the concurrence of very oxidant, or
even hyperoxidant conditions at high acidity. This is probably related to
pyrite oxidation processes, which give rise to the release of H+ and
SO42− into the environment.In order to assess the potential of
these soils to become further acidified by the oxidation of residual green
pyrite and pyrite cinder wastes, and thus, to estimate how this would affect
the future evolution of Hg species, the pH of oxidation was determined. The pH
of oxidation establishes the minimum pH value that could be produced if all
reduced substances were abruptly oxidized [19]. Values of pH in H2O,
KCl, and H2O2 (pH of oxidation) of the selected soil
samples are shown in Table 1. Comparison between values of pH-H2O
and values of pH of oxidation revealed a decrease in pH of more than 1 unit,
after oxidation with H2O2, in eight out of the 23 selected
soil samples, although pH values below 5 were reached in only four of the
soils. The results thus indicate a low-to-moderate potential of these soils for
further acidification processes.Finally, it should be
noted that several Hg species, such as elemental Hg and neutral organic Hg
(e.g., dimethyl-Hg), have a high vapor pressure and can be a significant source
of atmospheric Hg [28]. Over 90% of the mercury found in the atmosphere is
gaseous Hg0, whereas only a small amount occurs as methylated forms,
although the latter are of greater concern because of their high toxicity and
bioavailability in the environment [7]. Volatile forms of Hg may become
redistributed and deposited in nearby soils and plants as a result of
condensation under higher air humidity and cooler conditions [29]. Measurements
of Hg concentration over background vegetation tissue may thus indicate the
extent of these processes, as discussed in Section 3.5


### 3.5. Mercury Accumulation in Plants

Foliar concentrations
of Hg in the plants under study (Rubus fruticosus L., Osmunda cinnamomea, and Acer sp.) in the surroundings of the
Hg-fulminate production plant ranged between 0.3 to 12.7mgkg−1 (see
Table 2), whereas foliar Hg concentrations in the same species located in an
uncontaminated site ranged between 0.03 to 0.08mgkg−1 (see Table 2). Thus, the foliar Hg concentrations in vegetation at the contaminated site
were up to 3 times more (for Acer sp.), 8 times more (for Rubus sp.), and 17 time more (for fern) than those in plants at the uncontaminated
site. On the other hand, the data obtained also indicated that the Hg
concentration in leaves of Rubus sp. and fern increased linearly as the
soil Hg concentration increased up to 600mgkg−1  (r2 =
0.66, and 0.88, resp.), whereas no clear relationship was found between soil
and foliar Hg concentrations for Acer sp. Moreover, in each contaminated
site, Hg concentrations in fern leaves were consistently higher than those in
the other plants studied, suggesting a higher capacity of the former species to
accumulate Hg.Table 2
Mercury concentration of surface horizons and of leaf tissues of three
different species taken at different sites in the contaminated area, except
site 1, which is a noncontaminated site located 300 km away from the study
area but under similar climatic conditions.
Site 1Site 2Site 3Site 4Site 5Site 6Site 7Site 8mgkg−1Soil0.0314.7828.237.611518160714465fern0.71n.a.(a)3.143.364.7612.3712.6712.26Rubus sp.0.330.330.450.570.992.75n.a.2.34Acer sp.0.75n.a.0.370.39n.a.1.890.732.43n.a.(a) Not available.Unlike the majority of heavy metals,
most Hg present in above-ground biomass is taken up through leaves, either as
volatileHg0 [30] or to a lesser extent, as divalent gaseous Hg and
particulate Hg [31]. Uptake of Hg from the soil solution, through the roots, as
ionic Hg has also been reported, but translocation to aboveground biomass is
limited [32]. Thus, the high foliar Hg concentrations in forest stands close to
the Hg-fulminate production plant may be mainly attributed to deposition of
atmospheric Hg, as already indicated by other researchers [33]. Mercury-contaminated
plants, on the other hand, can also act as a source of Hg to (i) the atmosphere,  under low ambient air Hg concentrations [34],
and (ii) soils and waters through litterfall [7]. In the latter case, Hg tends
to accumulate more in forest soils than in open areas, because of the huge
amount of litter produced by forest species, giving rise to a large amount of
immobilized Hg on the forest floor. Temperature and temperature fluctuations as
well as air currents are lower under the forest canopy than in open areas,
whereas air humidity is higher, thus limiting Hg vaporization. Moreover, the
larger surface area of leaves in forest vegetation exposed to Hg air
deposition, as compared with nonforest ecosystems, may act as a large sink for
atmospheric Hg from other sources, whereas it may impede the loss of Hg
reemitted from the system by condensation of Hg as it reaches the leaves. This
may explain the accumulation of Hg observed in soils under the dense deciduous
forest vegetation of the study area located at a certain distance from the Hg
source.


### 3.6. Remediation Strategy

After the characterization study, a
plan for thorough cleaning up of the Hg contamination at the study site was
established. Cleanup has entailed the excavation and removal of all
contaminated material containing more than1000mgkg−1 Hg, from the
site and its transportation to a secure dump site. No other remediation
techniques such as treatment with Na sulphide or thermic treatments with vapor
recovery for in situ
remediation were implemented because of the proximity of the contaminated site
to a city and the urgent need to remove the dangerous material.


## 3.1. General Soil Properties in the Study Area

The pH of the soils in
the surroundings of the Hg-fulminate facilities varied widely (see Table1).
Soil pH-H2O  values of these samples ranged from 3.9 to 7.8, and
those of pH-KCl from 3.6 to 8.0, whereas natural soils in the area are
moderately acidic (with surface horizons of pH 4-5 and subsurface horizons of
pH 5-6) [20]. The diverse activities carried out in the production plant have
caused changes in the acid-base conditions of the soils. In areas close to
where lime or concrete were applied, pH-H2O values are above 6,
whereas in areas with presence of untreated green pyrite and pyrite cinder
wastes—both of which are
wastes from the production of sulphuric acid—pH-H2O values are below 4. Organic C contents of mineral surface horizons of the
selected soils ranged from 12 to 120gkg−1, whereas those of subsurface horizons
ranged from 4 to 16mgkg−1 (see Table 1). Soils in the surroundings
of the production plant were also found to be contaminated with other heavy
metals in addition to Hg, such as Zn, Cu, Pb, Cd, and As (data not shown),
which are associated with the presence of pyrite cinder wastes, although the
contaminated areas did not always coincide. The present study focuses on the
area within the production plant that is contaminated with Hg.Table 1
Values of pH in water, KCl, andH2O2 (pH of oxidation),
Eh of selected soil samples, organic C content, and Hg concentration of
selected soil samples. Standard errors of Hg concentrations are indicated in
parentheses (n=4).
SiteHorizonDepthpH-H2OpH-KClpH-oxidationEhOrganic CHg conc.(cm)(mV)(gKg−1)(mgkg−1)I-8O5.906.214.09468182.0838 (22)I-8Ah10–106.736.634.9950318.2234 (6)P-6Ah10–106.526.406.4248457.06.96 (0.12)A-2Ah10–156.205.755.3427819.033.6 (1.2)M-2Ah10–107.107.586.4426875.03377 (39)M-3Ah10–107.597.465.9628878.05883 (252)M-4Ah10–57.587.275.93284104.06350 (135)M-5Ah10–57.687.676.4439615.01546 (81)M-6Ah10–207.076.835.8620056.01687 (222)M-9Ah10–305.954.945.3628112.026.4 (0.2)M-10Ah10–56.936.524.8327589.09043 (779)M-11Ah10–205.364.814.0553364.9392 (5)M-12Ah10–204.764.513.20485120.0280 (6)P-11Ah210–207.457.126.2742634.843.7 (1.1)L-4Ah210–204.243.933.7862030.250.7 (3.0)L-4CA70–807.797.446.5045415.8421 (15)P-13C10–204.224.164.205514.2109 (2)P-14C8–154.084.274.455437.4132 (2)P-15C10–185.645.866.135027.2150 (11)P-15C18–434.233.784.155493.9212 (9)P-15C43–933.884.713.935438.035.3 (1.6)P-16C10–303.873.613.326396.127.1 (1.9)P-8C78–2107.388.047.284235.02.46 (0.06)


## 3.2. Mercury Distribution in the Soils of the Study Area

Digital maps of Hg
distribution in  the surface horizons of soils in the area around the
Hg-fulminate facilities were generated (see Figure2). Extremely high levels of
Hg were detected in the discharge area for the wastewater produced during the
washing procedures in the production plant (see Figure 2(a)), with
concentrations higher than 6.5gkg−1 (with a very highly contaminated spot
in which Hg concentration in the first 5 cm depth was 30gkg−1, although
this was not included in the interpolation to avoid distortion of the Hg
concentration gradients). In this highly contaminated spot, elemental Hg was
visually identified as droplets. Some distance away from the trail of runoff
wastewater, Hg concentrations decreased to levels ranging between 1 and 5gkg−1. Mercury
concentrations were also high in the NE vicinity of the production plant, with
values above 0.1gkg−1, covering an extension of ~0.5 ha, in contrast with the
concentration of  0.003mgkg−1 Hg detected in a noncontaminated
parent material in the soils close to the study area. The results thus show a
typical point source distribution pattern, with Hg levels decreasing with
distance from the production plant.Digital maps of Hg distribution in the soils of the area
around the Hg-fulminate facilities. (a) Surface Hg concentrations, (b), (c),
and (d) depths to which Hg concentrations reached values above 1, 0.1, and0.04gkg−1,
respectively.
(a)(b)(c)(d)Mercury is generally of low mobility
because of its high density, which explains the high concentrations in the
vicinity of the disposal site, at the wastewater discharge
area, and some meters downstream. In addition to wastewater discharge and spills
from containers, Hg contamination of soils away from this point may be
attributed to Hg volatilisation—either through
the exothermic reactions of the Hg-fulminate production process or
physicochemical/microbial-induced reactions occurring in contaminated soils—with subsequent
condensation in cooler areas of the production plant and in the surrounding
forest stands (see Section3.5). The subsurface lateral movement of water
contaminated with Hg mineral and organic particles in suspension over the
soils, as well as heavy runoff to surface water, may also be important sources
of the metal downstream (NE direction).The furthest depths, at
which the concentration of Hg reached values above 1, 0.1, and0.04gkg−1, are
indicated in Figures 2(b), 2(c), and 2(d), respectively. In two contiguous
sampling points close to the production plant, Hg concentrations above 1gkg−1 were
observed down to a depth of 40 cm (see Figure 2(b)). In this highly contaminated spot,
concentrations above 0.1gkg−1 were observed even at 1 m depth (see Figure 2(c)). Moreover,
concentrations above 0.04gkg−1 were also observed between 10 and 40 cm depth in the N and NE
directions (see Figure 2(d)). The results obtained thus indicate high
accumulation of Hg in surface horizons, mainly attributed to the repeated entry
of the contaminant to the surface—through spills,
waterflow, or condensation of volatile Hg—and which was
probably retained in the soil by organic matter and to a lesser extent by clay
particles. The presence of Hg in deeper horizons in the sites indicated above
may be related to the downward movement of Hg associated with soluble organic
matter, as previously reported in [12, 17], although more research is needed
to confirm this.


## 3.3. Total Hg in the Particle-Size Subsamples

Comparison was made of
the concentrations of Hg within the different particle size subsamples of
selected soil samples (see Figure3). In general, the results show that Hg was
distributed within all the particle sizes studied, and followed a relatively
homogenous pattern, with a tendency for concentration to increase as the particle
size decreased in the P-11 and P-15 soil samples. Fernández-Martínez et al.
[21] observed a generally higher Hg concentration in the finest particle-size
subsamples, which was attributed to the higher Hg sorption capacity of clay
minerals, Fe and Al oxy-hydroxides, and humus surfaces, all of which tended to
concentrate in the finest grain sizes. Studies carried out to date indicate
that in acid soils (pH < 4.5–5.5) the organic material is the only effective
sorbent for inorganic Hg, whereas in nearly neutral soils (pH > 5.5–6), iron
oxides and clay minerals may become more effective [7, 11, 12]. In this case,
the four samples studied differed greatly in soil pH, organic matter content,
as well as in Hg content (see Table 1), and no relationship was found between
the Hg distribution in these particle sizes fractions and these soil
properties.Concentration of Hg within the different particle-size
subsamples of selected soil samples.
(a)(b)(c)(d)Both elemental Hg andHg2+ tend to be strongly sorbed to the humic
fraction of soils [7], although the former has less affinity for organic matter than Hg2+ species [22, 23]. Moreover, elemental Hg readily vaporizes and can thus be
reemitted into the atmosphere, especially during periods of high temperature
[10]. Under acidic conditions, Hg0 may be oxidized into Hg(I) and
Hg(II) [24], although Hg(I) does not seem to occur as a stable species in soils
[7]. On the other hand, because of the strong affinity of Hg2+ for
humic substances [12, 14, 25] only trace contents of Hg2+ are
generally found in soil solution [26], either as free Hg ions or as soluble Hg
complexes, which are bioavailable. Neither speciation nor sorption processes
were investigated in the present study, although the geochemical evolutionary
trends of Hg in the contaminated soils were inferred from pH-Eh diagrams (see Section 3.4).


## 3.4. Geochemical Evolutionary Trends of Hg in the Contaminated Soils

One of the techniques that can be
used to establish the geochemical evolutionary trends of Hg in the contaminated
soils is the consideration of pH and Eh values of the soil samples, and the
identification of thermodynamically stable Hg species by means of Eh-pH diagrams,
although it must be taken into consideration that these diagrams are simplified
models of very complex systems. The Eh-pH diagram for an  Hg-O-H-S-Cl system is
shown in Figure4 [27], and the Eh and pH values of selected soil
samples from
the study   area are represented. The results obtained (see Figure 4) show that
the group of soil samples with Eh values below 400 mV includes all the soil
samples with Hg concentrations above 1gkg−1, and all correspond to
surface horizons (see Table 1). According to the Eh-pH diagram, Hg0 is the most thermodynamically stable species in the first group of soils, which
is consistent with the fact that these soils were sampled close to the
discharge exit of wastewaters rich in Hg0.Figure 4
Mercury Eh-pH diagram for an Hg-O-H-S-Cl system. Values of Eh
and pH of selected soil samples are displayed. The assumed activities for
dissolved species are Hg:10−8 M, Cl: 10−3.5 M, and S: 10−3 M.On the other hand, the
concentrations of Hg in all soil samples with Eh values>400 mV were below 1gkg−1 (see Figure 4), and according to the Eh-pH diagram, Hg2Cl2, Hg22+, and Hg0 were the most thermodynamically stable species under
the conditions used. However, it is known that Hg(I) has the ability to
disproportionate and equilibrate according to the equation Hg(I) = Hg(0) + Hg(II), with the disproportionation reaction for soils shifted to the extreme
right side, because the high retention of Hg2+ [7]. Thus, Hg(I)
does not appear to occur as a stable species in soil [7]. Finally, within the
latter group of soil samples, the pH of those with Eh values above 550 mV was
below 4.3 (see Figure 4), which reveals the concurrence of very oxidant, or
even hyperoxidant conditions at high acidity. This is probably related to
pyrite oxidation processes, which give rise to the release of H+ and
SO42− into the environment.In order to assess the potential of
these soils to become further acidified by the oxidation of residual green
pyrite and pyrite cinder wastes, and thus, to estimate how this would affect
the future evolution of Hg species, the pH of oxidation was determined. The pH
of oxidation establishes the minimum pH value that could be produced if all
reduced substances were abruptly oxidized [19]. Values of pH in H2O,
KCl, and H2O2 (pH of oxidation) of the selected soil
samples are shown in Table 1. Comparison between values of pH-H2O
and values of pH of oxidation revealed a decrease in pH of more than 1 unit,
after oxidation with H2O2, in eight out of the 23 selected
soil samples, although pH values below 5 were reached in only four of the
soils. The results thus indicate a low-to-moderate potential of these soils for
further acidification processes.Finally, it should be
noted that several Hg species, such as elemental Hg and neutral organic Hg
(e.g., dimethyl-Hg), have a high vapor pressure and can be a significant source
of atmospheric Hg [28]. Over 90% of the mercury found in the atmosphere is
gaseous Hg0, whereas only a small amount occurs as methylated forms,
although the latter are of greater concern because of their high toxicity and
bioavailability in the environment [7]. Volatile forms of Hg may become
redistributed and deposited in nearby soils and plants as a result of
condensation under higher air humidity and cooler conditions [29]. Measurements
of Hg concentration over background vegetation tissue may thus indicate the
extent of these processes, as discussed in Section 3.5


## 3.5. Mercury Accumulation in Plants

Foliar concentrations
of Hg in the plants under study (Rubus fruticosus L., Osmunda cinnamomea, and Acer sp.) in the surroundings of the
Hg-fulminate production plant ranged between 0.3 to 12.7mgkg−1 (see
Table 2), whereas foliar Hg concentrations in the same species located in an
uncontaminated site ranged between 0.03 to 0.08mgkg−1 (see Table 2). Thus, the foliar Hg concentrations in vegetation at the contaminated site
were up to 3 times more (for Acer sp.), 8 times more (for Rubus sp.), and 17 time more (for fern) than those in plants at the uncontaminated
site. On the other hand, the data obtained also indicated that the Hg
concentration in leaves of Rubus sp. and fern increased linearly as the
soil Hg concentration increased up to 600mgkg−1  (r2 =
0.66, and 0.88, resp.), whereas no clear relationship was found between soil
and foliar Hg concentrations for Acer sp. Moreover, in each contaminated
site, Hg concentrations in fern leaves were consistently higher than those in
the other plants studied, suggesting a higher capacity of the former species to
accumulate Hg.Table 2
Mercury concentration of surface horizons and of leaf tissues of three
different species taken at different sites in the contaminated area, except
site 1, which is a noncontaminated site located 300 km away from the study
area but under similar climatic conditions.
Site 1Site 2Site 3Site 4Site 5Site 6Site 7Site 8mgkg−1Soil0.0314.7828.237.611518160714465fern0.71n.a.(a)3.143.364.7612.3712.6712.26Rubus sp.0.330.330.450.570.992.75n.a.2.34Acer sp.0.75n.a.0.370.39n.a.1.890.732.43n.a.(a) Not available.Unlike the majority of heavy metals,
most Hg present in above-ground biomass is taken up through leaves, either as
volatileHg0 [30] or to a lesser extent, as divalent gaseous Hg and
particulate Hg [31]. Uptake of Hg from the soil solution, through the roots, as
ionic Hg has also been reported, but translocation to aboveground biomass is
limited [32]. Thus, the high foliar Hg concentrations in forest stands close to
the Hg-fulminate production plant may be mainly attributed to deposition of
atmospheric Hg, as already indicated by other researchers [33]. Mercury-contaminated
plants, on the other hand, can also act as a source of Hg to (i) the atmosphere,  under low ambient air Hg concentrations [34],
and (ii) soils and waters through litterfall [7]. In the latter case, Hg tends
to accumulate more in forest soils than in open areas, because of the huge
amount of litter produced by forest species, giving rise to a large amount of
immobilized Hg on the forest floor. Temperature and temperature fluctuations as
well as air currents are lower under the forest canopy than in open areas,
whereas air humidity is higher, thus limiting Hg vaporization. Moreover, the
larger surface area of leaves in forest vegetation exposed to Hg air
deposition, as compared with nonforest ecosystems, may act as a large sink for
atmospheric Hg from other sources, whereas it may impede the loss of Hg
reemitted from the system by condensation of Hg as it reaches the leaves. This
may explain the accumulation of Hg observed in soils under the dense deciduous
forest vegetation of the study area located at a certain distance from the Hg
source.


## 3.6. Remediation Strategy

After the characterization study, a
plan for thorough cleaning up of the Hg contamination at the study site was
established. Cleanup has entailed the excavation and removal of all
contaminated material containing more than1000mgkg−1 Hg, from the
site and its transportation to a secure dump site. No other remediation
techniques such as treatment with Na sulphide or thermic treatments with vapor
recovery for in situ
remediation were implemented because of the proximity of the contaminated site
to a city and the urgent need to remove the dangerous material.


## 4. Conclusions

The present study has shown the
extent of contamination of soils and vegetation close to an abandoned
Hg-fulminate production plant. A highly contaminated area was identified close
to the former discharge zone for the wastewater produced during the washing
procedures at the plant, where the concentrations of Hg in the surface horizons
were higher than1gkg−1. Analysis of the Hg Eh-pH diagram revealed that Hg0 is the most thermodynamically stable species in the highly contaminated surface
horizons in this area, which is consistent with the visual identification of Hg
droplets in the soil samples. On the other hand, about 0.5 ha of the surrounding
soil in the NE direction (following the dominant surface flow direction)
contained between 0.1 and 1gkg−1 Hg. In the latter area, the oxidized Hg species are more
thermodynamically stable than elemental Hg, as revealed by the Hg Eh-pH diagram.
It is possible that Hg(0) initially deposited in the soils was re-emitted with
subsequent condensation and oxidization in cooler areas of the production plant
and in the surrounding forest stands. Movement of Hg with water either by
lateral subsurface flow through the contaminated soils or by heavy runoff to
surface waters cannot be discounted. However, a more detailed investigation of
Hg speciation in the contaminated soils is required. In any case, it should be
considered that changes in atmospheric, soil climatic, physical, biological,
and chemical properties may lead to short- and long-term variability in the
speciation and total Hg concentrations in the soils in the study area.



---

*Source: 387419-2008-11-24.xml*