Environmental Pollution 186 (2014) 195e202
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Environmental Pollution
journal homepage: www.elsevier.com/locate/envpol
Assessing the influence of compost and biochar amendments on the
mobility and toxicity of metals and arsenic in a naturally
contaminated mine soil
Luke Beesley a, *, Onyeka S. Inneh b, Gareth J. Norton b, Eduardo Moreno-Jimenez c,
Tania Pardo d, Rafael Clemente d, Julian J.C. Dawson a
a
The James Hutton Institute, Craigiebuckler, Aberdeen AB15 8QH, UK
Institute of Biological and Environmental Sciences, University of Aberdeen, Aberdeen AB24 3UU, UK
Universidad Autónoma de Madrid, 28049 Madrid, Spain
d
CEBAS-CSIC, PO Box 164, 30100 Espinardo, Murcia, Spain
b
c
a r t i c l e i n f o
a b s t r a c t
Article history:
Received 28 September 2013
Received in revised form
19 November 2013
Accepted 27 November 2013
Amending contaminated soils with organic wastes can influence trace element mobility and toxicity.
Soluble concentrations of metals and arsenic were measured in pore water and aqueous soil extracts
following the amendment of a heavily contaminated mine soil with compost and biochar (10% v:v) in a
pot experiment. Speciation modelling and toxicity assays (Vibrio fischeri luminescence inhibition and
Lolium perenne germination) were performed to discriminate mechanisms controlling metal mobility
and assess toxicity risk thereafter. Biochar reduced free metal concentrations furthest but dissolved
organic carbon primarily controlled metal mobility after compost amendment. Individually, both
amendments induced considerable solubilisation of arsenic to pore water (>2500 mg l1) related to pH
and soluble phosphate but combining amendments most effectively reduced toxicity due to simultaneous reductions in extractable metals and increases in soluble nutrients (P). Thus the measureemonitor-model approach taken determined that combining the amendments was most effective at
mitigating attendant toxicity risk.
Ó 2013 Elsevier Ltd. All rights reserved.
Keywords:
Soil contamination
Organic amendments
Trace elements
Speciation
Pore water
1. Introduction
Contaminated, industrially impacted, mining and urban lands
are not only characterised by young, poorly developed soils but
often by their scarcity or absence of vegetation cover (Mench et al.
2010) associated with heavy metal toxicity. As well as restoring
natural cycling of organic matter and nutrients, re-vegetation of
contaminated soils is key to onward remediation. The presence of a
vegetative cover over bare soil reduces the potential for migration
of contaminants to proximal watercourses or inhalation following
soil erosion and windblow (Tordoff et al., 2000; Arienzo et al., 2004;
Ruttens et al., 2006) but a major limitation to re-vegetation is
phyto-toxic concentrations of heavy metals in soils (Pulford and
Watson, 2003). Organic soil amendments, such as composts, manures and sludges are now established amongst in-situ alternatives
to expensive and/or disruptive hard-engineered removal or
* Corresponding author.
E-mail addresses: luke.beesley@hutton.ac.uk,
(L. Beesley).
luke_beesley@hotmail.com
0269-7491/$ e see front matter Ó 2013 Elsevier Ltd. All rights reserved.
http://dx.doi.org/10.1016/j.envpol.2013.11.026
capping of contaminated substrates to reduce contaminantassociated risk (Brown et al., 2003; Hartley et al., 2009). The
contaminated site remediation agenda now relies more heavily on
assisted natural attenuation measures, such as promotion of soil
stability using retro-applied organic materials, increasingly viewed
as both more environmentally harmonious and cost-effective than
ex-situ works.
Composts are produced by spontaneous microbial bio-oxidation
of raw wastes to produce a biologically stable, humified organic
matter end-product from, amongst myriad of other sources, green
and agro-food industrial wastes (Bernal et al., 2007). In the latter
category, ‘alperujo’, a waste derived from olive oil production, is
abundantly available in Mediterranean regions and known for its
fertilisation qualities (Fornes et al., 2009). Once conveniently
composted, it is able to increase organic matter (OM), total-organic
carbon (TOC), and microbial biomass C and N in soils, which
stimulates plant growth on bare contaminated substrate (Clemente
et al., 2012). These provisions may be particularly useful to old mine
sites, typically existing with degraded or skeletal soils, depleted in
organic matter and nutrients, but abundant with phyto-toxic
metalliferous spoils (Wong, 2003). Other organic materials, such
196
L. Beesley et al. / Environmental Pollution 186 (2014) 195e202
Fig. 1. a) Panoramic view of the wider soil sampling area, b) experimental pot set-up showing rhizon pore water samplers in-situ and c) petri-dishes for toxicity seed germination
bio-assay in preparation.
as biochars, which are biomass pyrolysed under limited oxygen
supply, have also gained favour recently in the same context due
mainly to their ability to sorb metals, reducing phyto-toxic effects,
which would otherwise be a barrier to initial re-vegetation of bare
soils (Beesley et al., 2011; Gomez-Eyles et al., 2013). General benefits demonstrated by the experimental application of biochars to
soils have been increased water holding capacities (Thies and Rillig,
2009), C, N and P status (Lehmann, 2007; Chan and Xu, 2009;
Borchard et al., 2012), enhanced availability of Ca, Mg and Zn
(Major et al., 2010; Gartler et al., 2013), but reductions in the
leaching of some macronutrients in solution (Novak et al., 2009;
Laird et al., 2010). In the context of pollution control, the removal
of heavy metals and As from waste-waters (Mohan et al., 2007) and
heavy metals from soil leachates (Beesley et al., 2010; Beesley and
Marmiroli, 2011; Fellet et al., 2011) have also been reported as a
consequence of biochar additions.
Both alperujo composts and biochars have been proven to
contain low concentrations of some metals and As, often below
limits of detection, especially in the case of biochars (Clemente
et al., 2012; Freddo et al., 2012). This means the risk of introducing extra contaminant load is minimal after their addition to
soils. For example, As concentrations in redwood, maize, rice straw
and bamboo biochars of <0.3 mg kg1 were reported by Freddo
et al. (2012). However, both amendments have tended to increase
available As concentrations when added to pluri-contaminated
soils (Pardo et al., 2011; Clemente et al., 2012; Beesley et al.,
2013). Unlike metals, As may be mobilised following an increase
in pH (Fitz and Wenzel, 2002; Moreno-Jimenez et al., 2012) which
can be induced by some composted wastes and biochars and this is
of particular concern because As is toxic, even in low concentrations (World Health Organisation drinking water standard is
10 mg l1). Inorganic As (arsenite (As III) and arsenate (As V)) is
categorised as a class 1, non threshold carcinogen, but different As
species have different levels of toxicity (Tamaki and Frankenberger,
1992; Carbonell-Barrachina et al., 1999a,b; Meharg and HartleyWhitaker, 2002), and therefore, knowing the form of As that occurs in soils treated with organic materials is essential to understanding the associated post treatment risk.
The aims of the present study were to carry out an ecotoxicological pre-screening to identify potential risks posed by adding
alperujo compost and biochar to a heavily pluri-contaminated
substrate in the context of 1) the potential for leaching of metals
and As, 2) confounding factors affecting toxicity and 3) the potentially efficacious effects of combining the two amendments
together.
2. Materials and methods
2.1. Soil, amendments and treatments
Bulk (30 kg) surface samples of soils (0e10 cm) were randomly collected from
around a sporadically vegetated part of the La Mina Monica mine site area (Fig. 1a)
close to the village of Bustarviejo (40 520 07.0600 N; 3 430 48.8700 W), Madrid
(Spain). Previous studies have examined the distribution and fate of metals and As
in soils and vegetation from the surrounding tailings and mine drainage areas
(Moreno-Jiménez et al., 2009, 2010, 2011). Those studies identified a large spatial
variation in pH, organic matter and metal and As contents as well as uptake of Cd
and Zn to various locally abundant vascular plant species. For the present study
collected soils were mixed and homogenised into one composite sample, air dried
(24 C) for 4 days and sieved to a particle size of <2 mm. The alperujo compost
(olive mill waste compost) was prepared from a mixture of olive husk (alperujo)
and cow manure (10% fresh weight) whilst the biochar (BC) was produced from
residues of orchard prunings pyrolysed at 500 C; both amendments have been
described previously by Clemente et al. (2012) and Fellet et al. (2011), respectively.
For the present study new characterisation was carried out; organic matter was
determined by the loss on ignition method, whilst trace element total concentrations were determined on dried sub-samples (approx. 3 g) of soil, compost and
biochar using portable X-ray fluorescence (PXRF). Samples were held in a 25-mm
diameter plastic cup with a 4-mm thick polypropylene window (TF-240 film,
Fluxana, Germany) and analysed using a Bruker S-1 TurboSD PXRF instrument
(Bruker Nano Gmbh, Germany). The instrument was used in bench-top mode and
analyses carried out using the manufacturer’s soil programme. A certified reference
Chinese mineral soil (GBW07402; LGC, UK) was included periodically in analyses to
verify instrument accuracy.
Biochar and compost were mixed individually with soil at 10% (v/v) and in
combination 5% (v/v) of each amendment to equal a total amendment volume of
10%. An end-over-end shaker was used to thoroughly homogenize the mixtures
before quadruplicated treatments of approx.1.5 kg were placed into one litre pots,
compacted by light hand pressure and watered to reach 100% water holding capacity
(WHC). Thereafter pots were allowed to drain for 48 h to reach approx. 60% WHC
and maintained by weighing and addition of aqueous losses every 48 h. The treatments consisted of control soil without amendment (S), soil plus biochar (S þ BC),
soil plus compost (S þ C) and soil, compost and biochar combined (S þ C þ BC).
L. Beesley et al. / Environmental Pollution 186 (2014) 195e202
2.2. Pore water sampling and analysis
One rhizon sampler of 10 cm length (Eijkelkamp Agrisearch equipment, The
Netherlands) was inserted into each pot, at an angle of 45 and each pot was covered
with ParafilmÒ to prevent losses of water through evaporation. The soils were
maintained in a controlled environment chamber (Conviron, USA) in darkness at a
temperature of 22 C and 28% relative humidity. Soil pore water was extracted by
attaching 30 ml plastic syringes to each rhizon sampler after one (T1) and four (T4)
weeks (Fig. 1b). Approximately 5 ml of each pore water sample were taken for
analysis of pH (Jenway, UK) prior to elemental analyses. Dissolved organic carbon
(DOC) was determined using an aqueous carbon analyser (LabTOC, Pollution and
Process Monitoring, UK). Phosphate (PO4eP) was analysed using a flow injection
analyser FIAstart 5000 System (FOSS Tecator, Denmark). Total element analysis was
carried out using inductively coupled plasma mass spectroscopy (ICP-MS; Agilent
Technologies, USA). A suitable reference material (BCR-610) was used to verify accuracy. The remaining pore-water samples (3e5 ml) were prepared for arsenic
speciation analysis by performing a 1:10 dilution and acidifying the samples to give
a nitric acid concentration of 1%. Prior to analysis 0.1 ml of hydrogen peroxide was
added to 0.9 ml of acidified pore-water and stored at 4 C overnight. Arsenic
speciation was quantified by HPLCeICP-MS as described by Williams et al. (2007).
Separation was performed on a PRP-X100 10-mm anion-exchange column
(250 4.6 mm) with a mobile phase of 6.66 mM ammonium hydrophosphate and
6.66 mM ammonium nitrate, adjusted to pH 6.2 using ammonia. An arsenic species
mix of As III, As V, DMA and MMA (10 mg l1) was used to establish the retention time
for the arsenic species. The sum of species correlated with the total As measurements (p < 0.001) and the mean percentage recovery of the sum of species
compared to the total arsenic analysis was 89.5% 4.5% (mean SD).
2.3. Toxicity bio-assays
Samples of fresh soil and treatments following the final pore water sampling
(T4) were taken for toxicity testing using two bio-assays, applied to water extracts.
Firstly water extracts were prepared for trace element determination (1:10 w:v) by
shaking overnight, centrifuging and filtering to remove remaining particulate
matter. An aliquot of 9 ml was separated and mixed with 1 ml of 5 mM EDTA solution, and kept refrigerated until As speciation analysis (HPLCeAFS Millennium
Excalibur, PS Analytical, UK) whilst the rest of the extract was used for trace elements determination by inductively coupled plasma optical emission spectroscopy
(ICP-OES, Thermo Scientific, UK). For the germination testing a second batch of
extracts were prepared as above but these were further centrifuged at 750 rpm for
10 min to remove suspended solids but retain colloids (>10 mm) following the
method of Bao et al. (2011). A germination success and root emergence test, using
Lolium perenne L. var Cadix was performed in triplicate petri dishes containing
cellulose paper saturated with extracted solution. Fifteen seeds were placed between an upper and lower paper (Fig. 1c) and petri dishes were sealed with ParafilmÒ to prevent moisture losses and incubated for 96 h in darkness at 28 C and 60%
relative humidity in environmental chambers. Successfully emerged plants (those
with root elongation) were counted, their three longest root lengths measured
(following Moreno-Jimenez et al. (2011)) and the emergence success calculated as a
percentage of the 15 seeds applied having germinated.
A final batch of water extracts was prepared as described above for trace elements determination. These were diluted with NaCl solution (2% w/v) at the
following range: 0, 6.25, 12.5, 25, 50 and 70% (v/v), and used in a toxicity bioassay
that determined luminescence inhibition of the bacteria Vibrio fischeri (ISO 11348-2,
1998), using a BioToxÔ Kit (Aboatox Oy, Finland). The decrease of luminescence was
measured in duplicate after 30 min contact with the extracts at 15 1 C. The results
are thus expressed as the % of the soil extract that is added to the NaCl solution to
cause a 50% reduction in luminescence in the contact time period.
2.4. Statistical analyses and metal speciation modelling
Linear regressions were performed using pore water pH and trace elements,
DOC and phosphate concentrations (SPSS v.15.0). Visual MINTEQ (v.3.0, 2012) was
used to predict metal (Cd, Cu, Pb and Zn) and As speciation in pore water using input
parameters pH, DOC, Fe, P, Ca and Mg in soil solution.
3. Results
3.1. Effects of amendments on pore water and water extractable
metals and As
Low organic matter content but high total concentrations of As
and metals, especially Zn, were notable features of this mine soil
(Table 1) in contrast to compost and biochar, whose organic matter
contents exceeded 50% and total element concentrations were
<500 mg kg1 (Table 1). Concentrations of elements in pore water
extracted from pots showed various amendment responses; with
compost (S þ C) or compost and biochar (S þ C þ BC) mobilisation
197
Table 1
Organic matter content (%) and total concentrations (mg kg1) of heavy metals and
As in soil (S), alperujo compost (C) and biochar (BC) (mean n ¼ 3; se). ND ¼ Below
the limit of detection.
S
C
BC
Organic matter
As
Cd
Cu
Pb
Zn
1.7 (0.1)
62 (0.8)
51 (0.6)
7490 (32)
63 (2)
13 (2)
74 (16)
ND
ND
2940 (35)
94 (4)
449 (8)
4170 (33)
64 (6)
17 (7)
13,200 (55)
302 (4)
483 (7)
of As occurred (S ¼ <200 mg l1, S þ C ¼ >2500 mg l1), whilst
biochar’s effect was comparatively muted on this metalloid
(S þ BC ¼ <1000 mg l1; Fig. 2). Common to control and treatments,
concentrations of As in pore water sampled at T4 (4 weeks)
exceeded those sampled at T1 (1 week). Speciation analysis of pore
water samples confirmed that all As was in inorganic form (As III
and V). Metal concentrations in pore water demonstrated the
opposite general trend to those of As; for Cd, Cu, Pb and Zn biochar
proved most effective in reducing concentrations in pore water (up
to 175 fold for Cu), especially at T1 except for Cd, which was
reduced furthest at T4 (Fig. 2).
The proportion of the total concentrations that were water
extractable were generally low for soil (Table 2) but amendment
variously affected extractability. Compost addition, whether alone
or combined with biochar (S þ C and S þ C þ BC) resulted in highest
concentrations of water extractable As, Cd, Cu, and Pb where biochar’s effect alone was muted in comparison. The amendments
themselves exhibited variable element specific water extractabilities; for compost, Cu and Zn were 10 fold more extractable than for
biochar (Table 2), as a proportion of the total concentration
(Table 1). For Pb >1.5% of the total was water extractable from
biochar (Table 2) and for As extractability was equal between
compost and biochar (0.5% of total).
3.2. Temporal trends in pH, and DOC and phosphate concentrations
in pore water
The addition of all amendments to the soil promoted an increase
in pore water pH in excess of 3 units, with compost and biochar
combined affecting the biggest increase (S þ C þ BC; Fig. 3). This
was slightly more pronounced at T4 than T1 which was also the
case for DOC, where concentrations at T4 exceeded those at T1,
apart from for biochar amendment alone (S þ BC) where concentrations were 10e20 fold lower than those from soil plus compost
(S þ C) or biochar and compost (S þ C þ BC). Phosphate concentrations in pore water from soils with all treatments exceeded those
from the control at both T1 and T4; compost (S þ C) had the
greatest magnitude of effect, promoting an increase in pore water
phosphate concentrations (w0.4 mg l1) compared to the control
without amendment (<0.05 mg l1; data not shown). In terms of the
influence of pH, DOC and PO4eP on pore water metal and As concentrations, linear regression analysis demonstrated the predominant influences of pH, DOC and P on As, and pH and DOC on metal
mobility (Table 3). Speciation modelling using pore water data
resulted in metal specific influences of the amendments; in all
cases percentages of free metals were reduced by the amendments,
to the greatest extent for Cu and Pb by compost (S þ C and
S þ C þ BC). Biochar alone (S þ BC) was relatively ineffective for
reducing free Cd and Zn whilst all of the amendments also provoked the deprotonation of arsenate in solution with respect to the
control soil (Table 4). Chemical analyses of pore water and water
extracts (HPLCeAFS) confirmed that As(V) was the predominant
species in all treatments, so amendments did not provoke reduction/methylation of As.
198
L. Beesley et al. / Environmental Pollution 186 (2014) 195e202
4000
140
As
Cd
120
100
80
2000
10
-1
---------------------------------- Pore water concentration (µg l ) ---------------------------------
T1
T4
0
0
1400
140
Cu
Pb
10000
9000
1200
120
1000
100
800
80
600
60
400
40
200
20
Zn
8000
7000
6000
500
0
0
S
S+C
S+BC S+C+BC
0
S
S+C
S+BC
S+C+BC
S
S+C
S+BC
S+C+BC
-------------------------------------------------- Soil treatment ----------------------------------------------------Fig. 2. Concentration of arsenic (As) and selected heavy metals in pore water from soil (S), soil plus compost (S þ C), soil plus biochar (S þ BC), and soil plus compost and biochar
(S þ C þ BC) sampled 1 week (T1) and 4 weeks (T4) following commencement of the experiment (mean n ¼ 4; se).
3.3. Toxicity assessments
Combining soil with compost (S þ C) more than doubled Lolium
perenne germination success in water extracts (from 15.5% for S, to
35.6% in S þ C; Table 5); biochar alone (S þ BC) and compost and
biochar (S þ C þ BC) also resulted in an enhanced success, but to a
lesser extent (Table 5). Root lengths mirrored germination success,
with similar percentage improvements after amendment addition
compared to the control (40e57%).
In the bacteria Vibrio fischeri luminescence inhibition test, the
combination of compost and biochar (S þ C þ BC) resulted in the
Table 2
Water extractable concentrations (mg kg1) of heavy metals and As following the
experimental period (T4) in soil (S), soil plus compost (S þ C), soil plus biochar
(S þ BC), soil plus compost and biochar (S þ C þ BC), compost only (C), and biochar
only (BC) (mean n ¼ 2; sd). ND ¼ Below the limit of detection.
Treatment
As
S
C
BC
SþC
S þ BC
S þ C þ BC
1.16
0.3
0.07
23
14
28
Cd
(0.09)
(0.01)
(0.06)
(0.44)
(0.95)
(1.32)
0.02
ND
ND
0.05
0.01
0.06
Cu
(0.01)
(0.00)
(0.01)
(0.00)
0.29
0.94
0.44
0.57
0.14
0.47
Pb
(0.07)
(0.02)
(0.08)
(0.00)
(0.03)
(0.03)
0.24
0.12
0.27
2.28
1.26
2.5
Zn
(0.12)
(0.01)
(0.08)
(0.42)
(0.49)
(0.04)
4.31
6.89
0.79
0.43
0.29
0.44
(0.28)
(2.01)
(0.17)
(0.07)
(0.05)
(0.06)
largest decrease in toxicity, assessed as the percentage of the soil
extract required, added to a non-toxic solution, to cause a 50%
decrease in luminescence (w50% for S þ C þ BC compared to <40%
for S þ BC, and w30% for S þ C; Table 5). The control soil without
amendment (S) required < 30% of the soil extract to induce halving
of luminescence intensity (Table 5). Values presented were
measured at 30 min contact time, but similar results were also
obtained after 15 min contact time (data not shown).
4. Discussion
4.1. Primary effects of amendments on soluble metals and As
Neither total nor water-extractable concentrations of metals or
As in compost or biochar exceeded those of soil (Tables 1and 2), so
the contrasting influence of the amendments on As and metal
extractability therefore must be related to the amendment effects
on geochemically confounding factors; this is likely to be especially
prevalent at mine sites with inherently low organic matter content
(Wong, 2003), which will likely yield a low cation exchange capacity (CEC) and weak retention of trace metals in the soil matrix. It
has previously been noted that amending soils with highly organic
materials can generate large concentrations of soluble organic
L. Beesley et al. / Environmental Pollution 186 (2014) 195e202
300
8
pH
DOC
T1
T4
250
-1
Pore water concentration (mg l )
199
7
200
6
150
5
100
4
50
0
3
S
S+C
S+BC
S+C+BC
S
S+C
S+BC
S+C+BC
---------------------------------------------------- Soil treatment ------------------------------------------------Fig. 3. Concentration of dissolved organic carbon (DOC) and pH of pore water from soil (S), soil plus compost (S þ C), soil plus biochar (S þ BC), and soil plus compost and biochar
(S þ C þ BC), sampled 1 week (T1) and 4 weeks (T4) following commencement of the experiment (mean n ¼ 4; se).
matter to which free ions can complex and co-mobilise with
organic ligands during leaching events (Bernal et al., 2007).
Compared to biochar alone (S þ BC), in the present study the
treatments including compost (S þ C and S þ C þ BC) stimulated
greatest DOC concentrations in pore water (Fig. 3); regression
analysis confirmed the influence of DOC on metal mobility
(Table 4). Comparatively the influence of biochar on DOC in pore
water was small (Fig. 3) whilst in other field (Jones et al., 2012) and
pot (Karami et al., 2011) studies no significant differences in DOC
concentrations were attributed to biochar application suggesting,
as is commonly reported, that its carbon pool is relatively stable and
insoluble. Similarly Clemente et al. (2012) did not find significant
changes in the concentrations of DOC in field collected pore water
from alperujo compost treated plots over two years, so some
longevity of the effect of this compost can also be supposed. Given
that both amendments had an organic matter content of >50%
(Table 1) it can be concluded that the organic binding effect of
metals to the amendments is more stable to biochar than to
compost due to lower DOC concentrations emanating from the
former.
The influence of DOC on metal mobility may be intraamendment as well as metal specific. Table 4 shows the proportion of major species of metals in the pore water predicted using
modelling of element solubility data. Clearly, free ion metals are
abundant in this soil without the addition of amendments due to its
low pH and organic matter content yielding a low inherent binding
capacity, as discussed previously. The majority of the metal pool is
complexed by DOC in the presence of compost (S þ C and
S þ C þ BC) in the case of all metals (Table 4), more noticeable for Pb
and Cu than for Zn and Cd. Biochar treatment did not enhance
metal complexation in pore water to the same extent as compost
which may explain the more effective retention of metals on solid
phases than with compost; previous studies have noted that biochars were especially effective for retaining Zn (Beesley et al., 2010;
Beesley and Marmiroli, 2011). This effect may rather depend on the
ability of the soil to retain DOC which is important in predicting
longer term efficacy of the amendments as divalent cationic metals
are quite reactive with soil particles, but complexes are less likely to
interact with soil constituents and can leach to ground and surface
waters (Nowack et al., 2006; Liu et al., 2007) posing increased risks.
4.2. Specific influences on arsenic
Competition can occur between DOC and As for retention sites
on soil surfaces (Fitz and Wenzel, 2002), resulting in an increase in
soluble As with increasing concentrations of DOC (Hartley et al.,
2009). This appears to be the case in the present study, as
confirmed by the regression analysis (Table 3) and explaining, or
partially explaining the amendments’ effects on As. As well as DOC,
liming effects can increase As in soil solution (Fitz and Wenzel,
2002), an effect confirmed in the present study by regression
analysis (Table 3). The mobilisation of As to soil pore waters
following alperujo compost addition has also been found
(Clemente et al., 2012). Given that it is widely reported that the
addition of biochars to soils has resulted in pH increases (Yamato
et al., 2006; Chan et al., 2007; Van Zweiten et al., 2010; Bell and
Worrall, 2011; Jones et al., 2012) it can be assumed that a generalised, albeit possibly transient effect of composts and biochar additions to contaminated soils will be to decrease the risk of metal
leaching, but potentially increase As leaching depending on local
geochemical conditions.
Biochars have been proven as sources of, or as enhancing the
bioavailability of P (Sohi et al., 2010; Cui et al., 2011; Ippolito et al.,
2012; Wang et al., 2012) as demonstrated in myriad pot and field
trials (Fellet et al., 2011; Hass et al., 2012; Quilliam et al., 2012;
Beesley et al., 2013). In the present study PO4eP concentrations
in water-extracts from compost were more than double those of
biochar (103 and 213 mg kg1, respectively; data not shown). As
phosphate is chemically analogous to arsenate (As(V)), increases in
P availability result in the release of As from soil surfaces into solution (Meharg and Macnair, 1992); regression analysis confirmed
the relationship of P with As in pore water in the present study
(Table 2) whilst speciation confirmed As(V) as the sole species in
water-extracts (data not shown). Previous study has identified
mainly (80%) Al and Fe (oxyhydr)oxides retained As in soils from
the vicinity of the sample area (Moreno-Jiménez et al., 2010); some
solubilisation of Al and Fe was affected by the addition of compost
in the present study (pore water and water-extracts; data not
shown) and, as Mikutta and Kretzschmar (2011) observed ternary
complex formation between arsenate and ferric iron complexes of
humic substances extracts, it is likely that a combination of DOC, P
200
L. Beesley et al. / Environmental Pollution 186 (2014) 195e202
Table 3
Linear regression equations between trace element concentrations in pore water
(pw) and other parameters in pore water (pH, DOC, P) (data from pore water samples taken at T1 and T4).
Equations
R2adj
Sig
[As]pw ¼ 7.1e3099$[P] þ 214$pH e 536
[Cd]pw ¼ 234e33.5$pH þ 0.076e33.3$[P]
[Cu]pw ¼ 2016e294$pH þ 1.18
[Pb]pw ¼ 156e22.2$pH þ 0.088$
[Zn]pw ¼ 14,763e2073$pH þ 3.69e2226$[P]
0.67
0.96
0.91
0.84
0.97
p
p
p
p
p
<
<
<
<
<
F-value
0.001
0.001
0.001
0.001
0.001
Table 5
Germination success and root length of Lolium perenne and EC50 (% of soil extract
compared to non-toxic control, that caused a 50% inhibition in luminescence) of
Vibrio fischeri in aqueous extracts of soil (S), soil plus compost (S þ C), soil plus
biochar (S þ BC), and soil plus compost and biochar (S þ C þ BC) (mean n ¼ 3; se).
Treatment
20.5
260
136
75
327
and Fe solubility controlled As mobility, mediated by pH increase in
the present study.
4.3. Relevance of toxicity risk assessment
For a range of organisms (including humans, other animals and
aquatic plants) inorganic As species are more toxic than organic As
species (Tamaki and Frankenberger, 1992; Meharg and HartleyWhitaker, 2002). However, for two terrestrial plant species tested
(turnip and radish) organic As species were demonstrated to be
more toxic than inorganic ones (Carbonell-Barrachina et al.,
1999a,b). All the detectable As species measured in pore water in
the present study were inorganic and in water-extracts only As(V)
could be detected so it may be concluded that the risk associated
with As after applying amendments to the present soil would be
from leaching to waters and aquatic ecosystems. In a previous
study, utilising the same biochar on an As contaminated mine soil,
Beesley et al. (2013) also found considerable As mobilisation, but
this did not equate to high concentrations of the metalloid in the
fruit of Solanum lycopersicum L. (tomato) grown in this soil. In the
present study it could have been hypothesised that some organic As
species would be detectable in soil pore water after amendment,
especially with compost, which generated a high concentration of
DOC; it has been shown that the addition of organic matter to soil
induced the production of volatile As species, arsines (Mestrot et al.,
2011) and that DOC correlates with organic As in pore water
(Williams et al., 2011). It has also been shown that the addition of
farm yard manure to soil has led to the production of organic As
species in pore water from anaerobic soil (Norton et al., 2013).
However, the process of As methylation in soils appears to occur
predominantly in anaerobic soils (Takamatsu et al., 1982; Blodau
et al., 2008; Moreno-Jiménez et al., 2013) and as the soils in the
present study were closely maintained at 60% WHC this may
explain why there was no detectable organic As species present as
redox conditions were likely never to have been reducing.
Re-vegetation stabilises contaminated sites (Arienzo et al.,
2004; Ruttens et al., 2006), by introducing vegetative cover over
bare soils (Tordoff et al., 2000). Therefore toxicity assays should
evaluate the likelihood of improving germination of vegetation
weighed against the risk of increasing contaminant leaching and
affecting aquatic toxicity. The observed toxicity reductions to both
seed germination and bacteria luminescence found in the present
study after adding amendments to the soil (Table 4) suggests that
Success (%)
Root length (cm)
EC50 (%)
0.8
1.67
1.33
1.47
27.4
31.4
37.5
48.4
Vibrio fischeri
Lolium perenne
S
SþC
S þ BC
S þ C þ BC
15.5
35.6
26.7
33.3
(4.4)
(5.88)
(3.85)
(3.85)
(0.37)
(0.32)
(0.03)
(0.22)
(3.6)
(3.6)
(1.2)
(7.9)
metal solubility was determinant in toxicity as both tests were
carried out on water-extracts. Many studies have reported close
relationships between plant and other organisms’ metal uptakes
and the free metal ion concentrations in the soil (Parker et al., 1995;
Thakali et al., 2006; Ashworth and Alloway, 2007). In the present
study compost induced the formation DOC-metal complexes to the
greatest extent, reducing free metals to <50% of the total predominant species in pore water (Table 4), explaining the greater
seed germination success and root elongation of Lolium perenne
(Table 5). Other factors are likely also to have enhanced seed
germination chances; Karami et al. (2011) alluded to a priming
effect finding that water soluble nitrogen concentration were
increased to the greatest extent after co-amending a Pb contaminated mine soil with a combined green waste compost and biochar
amendment. Therefore, in the present study the effect of the
amendments could also have been to enhance nutrient provision in
the case of the germination success and root length assay; this
theory tends to be supported by the water soluble P and also K
concentrations, which were greatest in the combined compost and
biochar amended soil (data not shown). Similar suggestions were
made by Pardo et al. (2011, 2013) to explain both an increase in
germination of L. sativum after alperujo compost amendment of a
Pb and Zn contaminated mine soil, and enhanced plant growth in
field plots amended with the same compost in a calcareous
contaminated soil.
Lastly, many benefits are only seen when organic or inorganic
fertilisers are added together with biochar amendment, suggesting
that biochar alone is unsuitable as a soil ameliorant from an
available nutrient point of view (Yamato et al., 2006; Chan et al.,
2007; Steiner et al., 2008; Asai et al., 2009; Van Zwieten et al.,
2010; Beesley et al., 2013). As degraded soils, such as those at
former mine or industrial sites, or many heavily disturbed urban
soils often lack basic functionality, such as sufficient nutrient capital to restart natural processes, biochars may not be the most
suitable single amendment for these sites. A solution appears to be
combining compost with biochar; in the present study this reduced
soluble metal concentrations compared to the control without
amendment, whilst also reducing toxicity effectively in both bioassays. Therefore to provide available nutrients for plant growth
whilst ensuring metals are not leached from soils and risk is
increased, combining alperujo compost and biochar seems
efficacious.
Table 4
Predicted speciation of metals and As in pore water using MINTEQ (% of total species; data from pore water samples taken at T4). Predominant species for each treatment are
highlighted with bold text.
Cd
S
SþC
S þ BC
S þ C þ BC
Cu
Pb
Zn
As
Cd2þ
DOC-Cd
[CdOH]þ
Cu2þ
DOC-Cu
[CuOH]þ
Pb2þ
DOC-Pb
[PbOH]þ
Zn2þ
DOC-Zn
[ZnOH]þ
99.3
32.7
96.1
56.2
0.74
67.3
3.8
43.7
0.03
0.05
0.07
77.2
1.2
35.9
3.1
22.8
98.4
56.7
95.4
0.01
0.42
7.4
1.6
62.9
0.61
23.1
1.6
37.1
99.2
73.1
97.8
0.17
3.8
0.6
98.8
23.4
93.5
44.5
1.61
76.3
5.9
54.7
0.26
0.61
0.84
H3AsO4
H2AsO4 -
HAsO4 2-
3.6
96.3
38.4
46.4
27.4
0.06
61.6
53.6
72.6
L. Beesley et al. / Environmental Pollution 186 (2014) 195e202
5. Conclusions
There are risks when applying organic materials to contaminated soils as this study identifies; primarily both alperujo compost
and biochar increased the potential for As leaching due to their
effects on pH, DOC and soluble P concentrations as confounding
factors in trace element geochemistry. Even discounting potentially
increasing the bioavailability of As, which was not directly assessed
by this study, leaching of this metalloid to waters is nonetheless
undesirable as it could protract risks beyond site boundaries. Of
course this may be countered by an increase in soil fertility, and
reduced phyto-toxicity, promoting a deal of plant growth within
catchments and stabilising soils to reduce windblown soil transport. Thus, for metal contaminated substrates, the present study
supports the combined application of compost and biochar. How
this relates to efficacy in the field over time needs to be determined
but the screening approach presented forms a useful precursor.
Therefore we recommend establishment of field trials on metal
contaminated sites to further evaluate the effectiveness of the
amendments investigated here with the eventual aim of their
incorporation into remediation strategies.
Acknowledgements
The authors wish to thank Claire Deacon, Ken Cruickshank,
rová (University of
Anette Moran, Dave Hadwen and Lenka Made
Aberdeen) and Renate Wendler (The James Hutton Institute) for
their assistance with analysis of samples. The authors also gratefully acknowledge the support of COST Action TD 1107 ‘Biochar as
option for sustainable resource management’ for granting a short
term scientific mission (STSM), which assisted the completion of
experimental work for this paper.
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