Ecotoxicology DOI 10.1007/s10646-014-1335-2
TECHNICAL NOTE
Water-extractable priority contaminants in LUFA 2.2 soil: back to basics, contextualisation and implications for use as natural standard soil A. C. Bastos • M. Prodana • J. M. M. Oliveira • C. F. Calhoˆa • M. J. G. Santos • A. M. V. M. Soares S. Loureiro
•
Accepted: 13 August 2014 Ó Springer Science+Business Media New York 2014
Abstract The natural LUFA 2.2 standard soil has been extensively used in hazard assessment of soil contaminants, combining representation with ecological relevance for accurate risk evaluation. This study revisited the waterextractable fraction of LUFA 2.2 soil, through consecutive soil wet–dry cycles and discusses implications of use as standard substrate in derivation of ecotoxicological data and toxicity thresholds. Potentially bioavailable contents of metals (177.9–888.7 lg/l) and the 16 polycyclic aromatic hydrocarbons (PAHs; 0.064–0.073 lg/l) were dependent on the number of soil wetting–drying cycles applied. Such contents were screened based on current EU guidelines for surface waters and reported toxicological benchmarks for aquatic organisms. Aqueous concentrations generally fit within recommended Environmental Quality Standards (EQS), except for Hg (0.13–0.22 lg/l; [Maximum Allowable Concentration—MAC—of 0.07 lg/l) and for the sum of benzo(g,h,i)perylene and indeno(1,2,3-cd)pyrene (0.005 lg/l; [double the Annual Average of 0.002 lg/ l). Further, aqueous As, Zn, Cd, Ni and Cr concentrations exceeded ‘lower benchmark’ values for aquatic organisms, possibly reflecting an inadequate derivation for ecotoxicological data. In turn, PAHs in LUFA 2.2 soil aqueous extracts, whilst individually, are not likely to constitute a
A. C. Bastos (&) M. Prodana J. M. M. Oliveira C. F. Calhoˆa M. J. G. Santos A. M. V. M. Soares S. Loureiro Department of Biology and CESAM, University of Aveiro, Campus Universita´rio de Santiago, 3810-193, Aveiro, Portugal e-mail: [email protected]; [email protected] A. M. V. M. Soares Programa de Po´s-Graduac¸a˜o em Produc¸a˜o Vegetal, Universidade Federal do Tocantins, Campus de Gurupi, Cx. Postal 66, Gurupi, TO CEP: 77402-970, Brazil
hazard to test biota exposed to its aqueous fractions. This study urges for potentially bioavailable fractions of reference and standard natural soils to be adequately characterized and addressed as part of the research aim, experimental approach and design, as well as the expected scope of the outcomes. Keywords LUFA 2.2 Standard soil Potential bioavailability Contaminants Toxicity thresholds
Introduction Generally, there is a need for a tiered and iterative approach in the environmental risk assessment posed by contaminated soils, in view of current EU regulations for Soil [COM (2006) 232] and Water (2008/105/EC) protection. Risk characterization traditionally relies on an initial screening that identifies potential scenarios of concern, which can be subsequently evaluated at a more targeted level. This involves comparing contaminant concentrations in the designated environmental matrix (e.g. soil, water) to generic quality targets that reflect in a possible significant risk, if surpassed (Weeks and Comber 2005; Merrington et al. 2006; Gonc¸alves et al. 2013). In a subsequent tier, ecological testing based on soil bioassays and aquatic tests using soil leachates, allows evaluating to which extent detrimental effects are acceptable at the organism and ecosystem levels (van Gestel et al. 2001, 2008; Loureiro et al. 2005a, b; Weeks and Comber 2005; Merrington et al. 2006). The derivation of toxicity (and contamination) thresholds relies on the use of standardized test procedures and reference soils, to ensure reproducibility, comparability, site-independence and general acceptance (Jensen and Pedersen 2006). This translates in accurate evaluation and
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A. C. Bastos et al.
extrapolation of risks to terrestrial ecosystems or aquatic bodies exposed to surface runoff or groundwater leaching, resulting in appropriate risk management and mitigation decisions (e.g. Loureiro et al. 2005a, b; van Gestel 2008). In this context, various standard substrates aiming at harmonizing ecotoxicological testing have already been developed by a number of European and international organizations, including the OECD artificial soil (OECD guideline no. 207 1984), the natural LUFA 2.2 soil (Speyer, Germany) and a set of European natural soils (EUROSOILS; Kuhnt et al. 1994). LUFA 2.2 soil is the most widely used natural standard and reference soil, due to large supply, representativeness and ecological relevance, thus ensuring no contamination and adequate performance of test biota (e.g. Domene et al. 2011; Ro¨mbke et al. 2006; Caetano et al. 2012). In light of the equilibrium partitioning theory and the concept of bioavailability, exposed organisms are affected by the metabolically available contaminant fraction, rather than by the total contaminant concentration in soil (ISO 17402:2008; Holmstrup et al. 1998; Frische et al. 2003; van Straalen et al. 2005; van Gestel 2008; Rocha et al. 2011). Natural soil spatial and temporal variability and the complexity of biotic and abiotic interactions in the soil ecosystem, pose a challenge to the development of approaches to effectively account and assess contaminant bioavailability in soils (e.g. Sijm et al. 2000; Frische et al. 2003; Hund-Rinke and Ko¨rdel 2003; Maliszewska-Kordybach et al. 2008). Nonetheless, there is consensus that waterextractable soil fractions can be regarded as potentially bioavailable, based on the relationship between observed toxicity and measured chemical concentrations in soil extracts (e.g. ISO 17402:2008; Kelsey et al. 1997; Frische 2002; Hund-Rinke and Ko¨rdel 2003). Additionally, waterbased extractions have the advantage of accounting for the exposure route (soil pore-water) to terrestrial organisms (e.g. Ma et al. 1998; Hund-Rinke and Ko¨rdel 2003; Lanno et al. 2004; van Gestel and Koolhaas 2004; ter Laak et al. 2006). The present study is a proof of concept and highlights the possibility of water-extractable components from the LUFA 2.2 standard soil to pose an additional stress and/or confounding factor to the test biota, when inferring biological responses to chemicals and toxicity thresholds. Firstly, we assessed the contribution of LUFA 2.2 soil as a direct potential source of priority contaminant groups (metals and the 16 PAHs) in ecotoxicological testing, by examining their release into aqueous soil solution, as influenced by consecutive soil wet–dry cycles. Such a procedure aimed at simulating a ‘worst-case scenario’ in respect to their potential bioavailability over time, being therefore in line with established approaches for characterization of soils and soil materials (Eisentraeger et al.
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2004; Jablonowski et al. 2012). Subsequently, a two-step screening of aqueous contaminant contents was performed based on current regulatory EU guidelines for surface waters (2008/105/EC) and reported ecotoxicological benchmarks for aquatic organisms (Suter II and Tsao 1996). By combining the information from both steps in a single strategy, we provide a straightforward and generic screening approach to identify a potential scenario of concern linked to the use of reference and standard natural soils, whose water-extractable fractions are not adequately characterized.
Materials and methods Soil characteristics The sandy loam LUFA 2.2 soil (\2 mm; LUFA Speyer, Germany) was the selected natural standard soil, with the following physico-chemical characteristics: pH (0.01 M, CaCl2): 5.5 ± 0.1; soil organic C (%): 1.93 ± 0.2; cation exchange capacity (100 cmol?/kg): 10.0 ± 0.8; sand (%): 81.3 ± 2.3; silt (%): 12.1 ± 1.3; clay (%): 6.60 ± 1.3; maximum water holding capacity (g/100 g): 45.2 ± 5.0; density (g/ml): 1.13 ± 0.045. Individual and total metal and PAHs contents in LUFA 2.2 soil are presented in Table 1. Soil wet–dry cycles and water-extraction The procedure used was a combination and adaptation of the methodologies described by Eisentraeger et al. (2004) and Jablonowski et al. (2012). Briefly, de-ionised water (1:2 mass:volume) was added to 100 g (oven-dry basis) of LUFA 2.2 soil (\2 mm) in triplicates and homogenized. Following agitation (overnight, 150 rpm, 20 ± 2 °C, dark) using a bench top orbital shaker, suspended soil was allowed to settle (15 min) and subsequently centrifuged (3000 rpm, 15 min; adapted from Eisentraeger et al. 2004). Collection of the supernatant was performed through vacuum-driven filtration using MilliporeÒ 0.45 lm paper filter, after which centrifugation of the supernatant (4000 rpm, 10 min) allowed clearing suspended organic and particulate matter. In turn, the wet soil residue was oven-dried (50 °C, 24 h) to its initial volumetric moisture content (8 % vol:vol), based on early experimental data (not shown). The described procedure constituted 1 cycle. Treatments corresponding to 6 and 12 cycles refer to 6 or 12 wetting–drying events prior to water-extraction, each initiating with fast rewetting of the oven-dry soil residue, obtained in the previous cycle (adapted from Jablonowski et al. 2012). While not in use, soil extracts were stored in glass containers, in the dark at 4 °C.
Water-extractable priority contaminants in LUFA 2.2 soil Table 1 Individual and total contents (dry weight) of metals and metalloids, as well as of the 16 priority PAHs in the reference soil Lufa 2.2, as a mean of duplicate measurements. When available, standard errors of means are given as ±SE
Parameters (lg/kg) Metals Cr
8831
Mn
139.0
Co
4.500
Ni
2452
Cu
1531
Zn
17220
As
11.60
Cd
140.1
Pb
15981
Hg P metals
46378
67.70
Screening approach Following characterization of the water-extractable LUFA soil fraction, a straightforward and generic two-step screening approach was performed. An initial screening of aqueous contaminant contents was performed based on current regulatory EU guidelines for surface waters (2008/ 105/EC), comparably to the first level in a tiered risk characterization of soils. To increase the ecological relevance of such an approach, a second screening step took place based on reported ecotoxicological benchmarks (Suter II and Tsao 1996), by which aqueous contents could pose an effective risk to aquatic biota in a non-specific testing scenario.
PAHs
NAP naphthalene, ACY acenaphthylene, ACE acenaphthene, FLU fluorene, PHE phenanthrene, ANT anthracene, FLT fluoranthene, PYR pyrene, CHR chrysene, BaA benz(a)anthracene, BbF benzo(b)fluoranthene, BkF benzo(k)fluoranthene, BaP benzo(a)pyrene, DBA dibenz(a,h)anthracene, BGP benzo(g,h,i)perylene, IND indeno(1,2,3-cd)pyrene
NAP
43.90 ± 3.59
ACY ACE
30.10 ± 0.33 29.30 ± 0.10
Results and discussion
FLU
31.70 ± 0.43
PHE
34.80 ± 0.23
Characterization and screening of the potentially bioavailable LUFA 2.2 fraction
ANT
22.40 ± 0.42
FLT
41.80 ± 0.94
PYR
43.20 ± 0.04
CHR
40.90 ± 0.11
BaA
15.30 ± 0.10
BbF
7.700 ± 0.15
BkF
6.800 ± 0.61
BaP
7.100 ± 0.24
DBA
6.400 ± 0.33
BGP
2.330 ± 0.41
IND P PAHs
366.0 ± 1.27
2.650 ± 0.13
Characterization of aqueous LUFA 2.2 soil extracts Extract characterization was performed in relation to pH (CaCl2) and dissolved organic carbon (DOC) using a TOC analyser. Total contents in the aqueous soil extracts were estimated for the 16 priority PAHs (by SPME-solid phase micro-extraction coupled to GC/MS-gas chromatography/ mass spectrometry) based on Campos et al. (2012), as well as metals and metalloids (by ICP/AES-inductively coupled plasma-atomic emission spectroscopy), namely cadmium (Cd), cobalt (Co), chromium (Cr), copper (Cu), manganese (Mn), zinc (Zn), nickel (Ni), lead (Pb) and arsenic (As), as described by Rodrigues et al. (2010). Mercury (Hg) was quantified by AAS-atomic absorption spectroscopy. Two replicate extractions were conducted and three replicate measurements were taken in each analysis (standard deviation between replicate measurements \10 %).
Table 2 summarizes the main physico-chemical characteristics of LUFA 2.2 soil aqueous extracts and compares aqueous metal and PAH concentrations with the EU water quality criteria and toxicological benchmarks for aquatic organisms. While aqueous metal concentrations were deeply influenced by the number of consecutive soil wet– dry events (177.87, 888.7 and 689.43 lg/l for 1, 6 and 12 cycles, respectively), extracted PAHs were largely independent of that (0.073, 0.071 and 0.064 lg/l for 1, 6 and 12 cycles, respectively). Metals and PAHs are ubiquitous in the environment. Background levels for such compounds are generally dictated by interactions in local geology, biogeochemistry, climate and anthropogenic activity (Reimann and Garrett 2005). It is not therefore, surprising the occurrence of such fractions in LUFA 2.2 aqueous extracts, which is in agreement with previous reports on background contaminant levels in LUFA’s solid matrix (e.g. \0.05 mg/kg of PAHs; Frische et al. 2003). Potentially bioavailable contaminant fractions in reference and standard substrates have been more scarcely addressed (exceptions include Loureiro et al. 2005a, b). The occurrence of such elements constitutes a challenge that should be addressed as part of the ecotoxicological testing approach, to ensure optimal performance of the test biota and accurate determination of toxicity thresholds. Aquatic regulations have been commonly considered in soil protection approaches, as discussed by Ferna´ndez et al. (2006). In the present study, aqueous concentration ranges were at least a factor of 10 below the EQS and MAC for metals and metalloids (Ni, Cd, Pb, Zn) as well as for the six PAHs (NAP, BaP, BbF, BkF, BGP, IND) regulated by the
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123
0.490
5.900
6.500
44.00 \1.000
0.270
Co
Ni
Cuc
Zn As
Cd
0.003 ± 1 9 10-4
0.006 ± 4 9 10-4
ANT
BaP
\LD
0.005 ± 6 9 10-4
BkF
a
\LD \LDa
0.004 ± 5 9 10-4
\LD
a
a
BbF
2.030 ± 0.33
\LDa
2.080 ± 0.15
1.160 ± 0.16
\LDa \LDa
\LDa \LDa
\LDa \LDa
CHR BaA
0.003 ± 3 9 10-5
0.002 ± 0.04
PYR
3.580 ± 0.02
2.400 ± 0.19
11.27 ± 1.45
\LD
a
5.490 ± 0.53
\LD
a
35.80 ± 5 9 10-3
689.4
0.130
1.300
0.100
18.00 4.000
8.300
10.00
1.300
643.0
3.300
103.0
0.007 ± 0.18
0.007 ± 9 9 10
FLT
-4
0.015 ± 8 9 10-4
0.011 ± 2 9 10-4
0.001 ± 3 9 10
\LD
PHE
-5
FLU
a
0.005 ± 1 9 10-4 \LDa
ACE
\LD
a
0.004 ± 3 9 10
-4
0.032 ± 4 9 10-3
888.7
177.8
0.027 ± 4 9 10-3
0.220
0.180
2.300
0.150
15.00 4.500
10.00
8.900
2.100
846.0
4.000
ACY
NAP
PAHs
Hgd P Metals
Pb
0.930
117.0
Mn
c
1.600
Cra
Metals
96.80
0.0500
R (BbF,BkF) = 0.030
2100 330.0
n/a
0.100
0.100
n/a
n/a
n/a
n/a
2.400–1.200
0.050
0.1000
n/a
n/a n/a
n/a
1.000
0.400
n/a
n/a
n/a
n/a
n/a
0.070
n/a
0.450–1.500
0.080–0.250c 7.200
n/a n/a
n/a n/a
n/a
20.00b
20.00b n/a
n/a
n/a
n/a
n/a
n/a
n/a
88.50
DOC (mg/l)
7.24
6.98
5.32
pH (0.01 M CaCl2)
MAC EQS
AA EQS
6 cycles
1 cycle
12 cycles
EU Directive for inland and other surface waters (lg/l)
Aqueous extract (lg/l)
n/a
n/a
33.60
n/a f
30.00f
n/a
80.00f
n/a
2.400
82.00
3.900c
1200 n/a
18.00
1400c
n/a
16.00
f
n/a
n/a
6.160
n/a
6.300f
n/a
23.00f
n/a
0.010
3.200
1.100c
110.0 n/a
12.00
160.0c
n/a
11.00
0.240
0.490
n/a
13.00
n/a
70.00
n/a
190.0
n/a
n/a
n/a
n/a 66.00
n/a
n/a
1500
2300
n/a
h
0.014
0.027
n/a
0.730
n/a
3.900
n/a
12.00
h
1.300e
n/a
n/a
n/a 3.100
n/a
n/a
23.00
120.0
n/a
Chronic
Acute
Acute
Chronic
Tier II (Secondary chronic) values
NAWQ Criteria
n/a
n/a
30.00
0.090
n/a
n/a
74.00
620.0
\0.23
18.88
1.700
36.41 892.0
3.800
\35.0
290.0
1780
73.18
Fish
g
0.300
0.650
15.00
\2.10
200.0
n/a
6.646
1.163
0.960
12.26
0.150
46.73 450.0
0.230
\5.00
g
g
g
g
n/a
n/a
54400
n/a
n/a
n/a
520.0
33000
5.000
500.0
2.000
30.00 48.00
1.000
5.000
2.000
n/a 5.100
2.000
6.132
Aquatic plants
\1100
Daphnids
Lowest chronic value for aquatic biota
Conventional toxicological benchmarks (lg/l) (Suter II and Tsao 1996)
Table 2 Screening of metals, metalloids and the 16 priority PAHs in aqueous extracts of standard LUFA 2.2 soil (after 1, 6 or 12 wet–dry and water-extraction cycles, 20 °C), based on the current EU regulation for surface waters (2008/105/EC) and the reported toxicological benchmarks (reviewed by Suter II and Tsao 1996)
A. C. Bastos et al.
0.0727
0.003 ± 1 9 10 0.003 ± 9 9 10 0.0713
0.064
\LD
a
R (BGP,IND) = 0.002 n/a
Chronic
Acute
Acute
Chronic
Tier II (Secondary chronic) values
NAWQ Criteria
Fish
Daphnids
Lowest chronic value for aquatic biota Aquatic plants
Conventional toxicological benchmarks (lg/l) (Suter II and Tsao 1996)
h
g
f
e
d
c
b
a
Estimated value, as described in Suter II and Tsao (1996) OSWER (Office of Solid Waste and Emergency Response) threshold values for aquatic life (U.S. EPA 1996)
Final acute and chronic values in derivation of sediment quality criteria (U.S. EPA 1993b)
Chronic NAWQC value for Hg (0.012 lg/l) is estimated based on final residue values (Suter II and Tsao 1996)
Values refer to inorganic Hg
Criteria dependent on water hardness and normalized to 100 mg/l
Value refers to Ni and its compounds
Indicated contents for As (V) and Cr(VI), which hold higher relevance in natural waters when compared to As (III) and Cr(III) (e.g. Economou-Eliopoulos et al. 2011)
NAP naphthalene, ACY acenaphthylene, ACE acenaphthene, FLU fluorene, PHE phenanthrene, ANT anthracene, FLT fluoranthene, PYR pyrene, CHR chrysene, BaA benz(a)anthracene, BbF benzo(b)fluoranthene, BkF benzo(k)fluoranthene, BaP benzo(a)pyrene, DBA dibenz(a,h)anthracene, BGP benzo(g,h,i)perylene, IND indeno(1,2,3-cd)pyrene, AA annual average, MAC maximum allowable concentration, NAWQ National Ambient Water Quality, U.S. water quality criteria, LD limit of detection
IND P PAHs
-5
0.003 ± 1 9 10-4 \LDa
-4
0.002 ± 4 9 10-4
BGP
\LDa
\LDa
\LDa
DBA
MAC EQS
AA EQS
6 cycles
1 cycle
12 cycles
EU Directive for inland and other surface waters (lg/l)
Aqueous extract (lg/l)
Table 2 continued
Water-extractable priority contaminants in LUFA 2.2 soil
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A. C. Bastos et al.
EU Directive for inland and other surface waters (2008/ 105/EC). Exceptions included the sum BGP ? IND, which was twice as high (0.005 and 0.006 lg/l for 1 and 6 cycles) as the annual average EQS for inland surface waters (2 lg/l; 2008/105/EC), with no MAC available; the Hg contents (0.18, 0.22, 0.13 lg/l for 1, 6 and 12 cycles, respectively) have also slightly exceeded the recommended MAC levels for inorganic or total mercury (0.07 lg/l) for inland and other surface waters (2008/105/EC). Total As contents (1.0, 4.5 and 4.0 lg/l respectively), while not targeted in the above Directive, were within the established as ‘typical’ for background natural waters (Smedley and Kinniburgh 2002). When compared to generic quality guidelines, screening based on reported aquatic toxicology benchmarks (Suter II and Tsao 1996) on either single or combinations of multiple endpoints may be an equally valuable tool to assess the likelihood of an effective risk to the test biota (Maliszewska-Kordybach et al. 2008). Conventional benchmarks are grounded on regulatory standards or on the responsespecific endpoints used to derive them, such as the acute or chronic NAWQC, the SCV (estimated with 80 % confidence as to not exceed the NAWQC for compounds lacking such criteria) and the LCV (lowest acceptable chronic values, recommended by U.S. EPA as substitutes for water quality criteria) for aquatic organisms (U.S. EPA 1992, 1993a, b). In respect to aqueous metal and metalloid contents (independently of the number of wet–dry cycles), these were 1–2 orders of magnitude below the corresponding NAWQC, suggesting that a significant risk to the test (aquatic) species can be discarded. Nevertheless, some metallic compounds were detected at concentrations above the SCV or the LCV for aquatic organisms, while aqueous Pb contents (0.93, 2.30, 1.3 lg/l for 1, 6 and 12 cycles respectively) were closely comparable to the chronic NAWQ value (3.2 lg/l; derived from median lethal concentrations -LC50- and Chronic Values -CV-) for this compound (Suter II and Tsao 1996). For instance, As contents slightly surpassed the SCV for aquatic organisms (3.1 lg/l), after 6 and 12 wet–dry cycles. Cd and Ni generally exceeded the LCV for daphnids (0.15 and \0.5 lg/l, respectively), while Cr contents were above the LCV for aquatic plants (2 lg/l). Amongst the detected metallic compounds, Zn concentrations surpassed the LCV targets for fish (36.41 lg/l), non-daphnid invertebrates (5.24 lg/l) and aquatic plants (30 lg/l) after the first extraction. Having surpassed ‘lower benchmark’ values (U.S. EPA 1992, 1993a), it suggests that Zn, As, Cd, Ni and Cr may represent an additional stress and/or confounding factor when inferring the biological response to aqueous extracts of LUFA 2.2 soil (Suter II and Tsao 1996). This may occur preferentially under acidic to slightly acidic (\pH 6.0) extract conditions and/or increasing DOC levels, due to
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enhanced metal solubility and extractability (e.g. Chuan et al. 1996; Antoniadis and Alloway 2002). For the PAHs which toxicological benchmarks were revised by Suter II and Tsao (1996; ACE, ANT, BaA, BaP, FLU, FLT, NAP, PHE) concentrations were generally, a factor of 10–100 below the respective NAWQC, SV and LCV, albeit ANT contents being only marginally below the LCV for fish (0.09 lg/l). Overall, PAHs in LUFA 2.2 soil aqueous extracts, whilst individually, are not likely to constitute a hazard to aquatic biota exposed to LUFA 2.2 aqueous extracts (Suter II 1995). Contextualization and potential implications for testing The choice of a standard or reference soil is of prime importance in estimating ecotoxicological thresholds, providing grounds for accurately inferring or refining ecological risk, with consideration of both physico-chemical and biological criteria. Generally, the standard soil represents a substrate for spiking with contaminants of interest to derive soil and water screening values (as solid or aqueous soil extracts) or a negative control in dilutions or aqueous extracts of contaminated soils, soil materials or wastes in aquatic and terrestrial biological assessment (van Gestel et al. 2001; Loureiro et al. 2005a, b; Caetano et al. 2012). For example, the derivation of soil and water quality criteria as well as predicted no-effect concentrations (PNECs) is grounded on standard bioassays or obtained by statistical extrapolation (EC 1996). The contaminant background concentration in the standard substrate can therefore influence that process, despite quantification of bioavailable fractions, often being poorly addressed. Either of such procedures generally also involves consideration of additional factors, including toxicity-dependent parameters, such as contaminant bioavailability and biomagnification potential, as well as species-specific sensitivities (e.g. van Gestel et al. 2001; Ferna´ndez et al. 2006). Therefore, questioning the suitability of the standard substrate due to background noise and/or added stress to the test biota, can lead to compromising the integrity of estimated toxicity thresholds. In the same way, it can confound data interpretation and limit reproducibility in tests and comparability between studies. It further adds to the already-challenging derivation of critical levels for specific contaminants, reference doses and contamination thresholds in the selected soil and aqueous matrices. Ultimately, if not adequately addressed, it can lead to inappropriate evaluation and prioritization of risks relating to contaminated soils, and consequently, to poor decision-making for risk management or mitigation (Suter II 1995). To the best of our knowledge, this is the first comprehensive characterization of water-extractable priority contaminant groups in a widely used natural standard and
Water-extractable priority contaminants in LUFA 2.2 soil
reference soil, which are generally not accounted for in risk characterization approaches and experimental set-ups. The procedure used was a simple and straightforward approach, which served to demonstrate the importance of considering the potentially bioavailable LUFA 2.2 fraction over time and the possibility of long-term leaching and runoff of contaminant fractions, as influenced by a natural soil process (consecutive dry-wet events). It further considers that: (i) such elements have dissimilar interactions with, and desorption rates from the soil matrix; (ii) there may be relative variations in background concentrations in the aqueous extracts to which aquatic test biota are exposed; (iii) aqueous concentrations may represent pore-water relevant exposures to terrestrial organisms, as influenced by soil wetting and drying. Based on comparison to generic water quality standards and available aquatic ecotoxicological data, screening of LUFA 2.2 water-extracts has helped identifying a potential conceptual bottleneck, in the context of tiered risk assessment and environmental quality criteria derivation procedures. It suggests that potentially bioavailable fractions should be adequately addressed in view of one’s research aim, experimental design and scope of outcomes. Further, it proposes a simple, yet conservative procedure that could be adapted to match specific assessment requirements and goals, in order to maximize the level of robustness and confidence in the estimated ecotoxicological criteria, independently of the standard soil used. The present study and output recommendations are intended to encourage discussion on how risk evaluation and data generation can cope with, adapt to and overcome the use of natural soils, in view of the ERA framework and current European Directives for Soil [COM (2006) 232] and Water (2008/105/EC) protection. Rather than undermining the use of LUFA 2.2 soil as a natural standard and reference substrate, it highlights the necessity to fully characterize any standard soil’s water-extractable fraction to include and help manage sub-optimal traits related to its inherent variability (Jensen and Pedersen 2006). In this way, the experimental design and considerations can be adjusted to specific study requirements, endpoints and goals, to improve relevance and refinement. Components that are more likely to contribute the most (or the least) to a biological response can be easily identified and ranked, whereas added stressors that can compromise the test biological integrity can be excluded or adequately accounted for. Most of the present discussion can also be applied to terrestrial test organisms exposed via the water pathway, by ingestion of suspended particles in soil pore water (EC 2003). Nevertheless, there are limitations for the described approach in the context of derivation of terrestrial ecotoxicological data. Despite the reasoning behind the
described extraction procedure overlapping to some extent with the concept of soil pore-water (Eisentraeger et al. 2004), aqueous contaminant contents and implications to terrestrial biota may not extrapolate straightforwardly and must be considered critically (Sijm et al. 2000; Frische et al. 2003; Hund-Rinke and Ko¨rdel 2003). Similarly, while addressing wet–dry cycles can account for long-term desorption as influenced by natural soil processes, long time-scales are not generally covered by conventional acute and chronic ecotoxicological tests. Estimated waterextractable metal and PAH fractions in this study should be regarded only in the context of the present discussion. Further, a certain degree of variability between LUFA 2.2 soil batches is expected, including pH, clay and organic matter contents, all of which are key parameters in controlling chemical availability of metals and organic compounds in soils (Jensen and Pedersen 2006; van Gestel 2008).
Conclusions Using a simple and generic screening approach to address and contextualize potentially bioavailable contaminants in LUFA 2.2 soil, this study highlights the need for suitably characterizing its water-extractable fractions. Potentially bioavailable fractions for priority metals and PAHs in LUFA 2.2 soil were generally in agreement with recommended EQS for surface waters (except Hg and the sum BGP ? IND) and the reported NAWQ criteria. Whilst individually, PAHs are not likely to constitute a hazard to the test biota exposed to LUFA 2.2 aqueous fractions. Nevertheless, As, Zn, Cd, Ni, Cr are among the metals which concentrations exceeded lower ecotoxicological benchmarks (SCV or LCV) for aquatic organisms. This study urges for potentially bioavailable fractions of the natural standard or reference soil to be adequately characterized and addressed for derivation of ecotoxicological data and contamination thresholds, according to the study aim, experimental approach and design, as well as the expected scope of outcomes. Acknowledgments This study was supported by European Funds through COMPETE and National Funds through the Portuguese Science Foundation (FCT), within projects PEst-C/MAR/LA0017/ 2013, FUTRICA (FCOMP-01-0124-FEDER-00008600) and CLIMAFUN (FCOMP-01-0124-FEDER-008656), which included the post-doctoral fellowships of Ana C. Bastos (BPD/CESAM/PTDC/ AAC-AMB/104666/2008) and Jacinta M.M. Oliveira (BPD/UI88/ 6463/2013), respectively. The authors also wish to acknowledge the FCT-funded post-doctoral fellowships of Carla F. Calhoˆa (SFRH/ BPD/74232/2010) and Miguel J.G. Santos (SFRH/BPD/72380/2010), as well as the European Commission through the Erasmus Mundus Program for the MSc Grant to Marija Prodana. Amadeu M.V.M. Soares is ‘Bolsista CAPES/BRASIL’, Project No. A058/2013.
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A. C. Bastos et al. Conflict of interest Authors declare that they have no conflict of interest.
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TECHNICAL NOTE
Water-extractable priority contaminants in LUFA 2.2 soil: back to basics, contextualisation and implications for use as natural standard soil A. C. Bastos • M. Prodana • J. M. M. Oliveira • C. F. Calhoˆa • M. J. G. Santos • A. M. V. M. Soares S. Loureiro
•
Accepted: 13 August 2014 Ó Springer Science+Business Media New York 2014
Abstract The natural LUFA 2.2 standard soil has been extensively used in hazard assessment of soil contaminants, combining representation with ecological relevance for accurate risk evaluation. This study revisited the waterextractable fraction of LUFA 2.2 soil, through consecutive soil wet–dry cycles and discusses implications of use as standard substrate in derivation of ecotoxicological data and toxicity thresholds. Potentially bioavailable contents of metals (177.9–888.7 lg/l) and the 16 polycyclic aromatic hydrocarbons (PAHs; 0.064–0.073 lg/l) were dependent on the number of soil wetting–drying cycles applied. Such contents were screened based on current EU guidelines for surface waters and reported toxicological benchmarks for aquatic organisms. Aqueous concentrations generally fit within recommended Environmental Quality Standards (EQS), except for Hg (0.13–0.22 lg/l; [Maximum Allowable Concentration—MAC—of 0.07 lg/l) and for the sum of benzo(g,h,i)perylene and indeno(1,2,3-cd)pyrene (0.005 lg/l; [double the Annual Average of 0.002 lg/ l). Further, aqueous As, Zn, Cd, Ni and Cr concentrations exceeded ‘lower benchmark’ values for aquatic organisms, possibly reflecting an inadequate derivation for ecotoxicological data. In turn, PAHs in LUFA 2.2 soil aqueous extracts, whilst individually, are not likely to constitute a
A. C. Bastos (&) M. Prodana J. M. M. Oliveira C. F. Calhoˆa M. J. G. Santos A. M. V. M. Soares S. Loureiro Department of Biology and CESAM, University of Aveiro, Campus Universita´rio de Santiago, 3810-193, Aveiro, Portugal e-mail: [email protected]; [email protected] A. M. V. M. Soares Programa de Po´s-Graduac¸a˜o em Produc¸a˜o Vegetal, Universidade Federal do Tocantins, Campus de Gurupi, Cx. Postal 66, Gurupi, TO CEP: 77402-970, Brazil
hazard to test biota exposed to its aqueous fractions. This study urges for potentially bioavailable fractions of reference and standard natural soils to be adequately characterized and addressed as part of the research aim, experimental approach and design, as well as the expected scope of the outcomes. Keywords LUFA 2.2 Standard soil Potential bioavailability Contaminants Toxicity thresholds
Introduction Generally, there is a need for a tiered and iterative approach in the environmental risk assessment posed by contaminated soils, in view of current EU regulations for Soil [COM (2006) 232] and Water (2008/105/EC) protection. Risk characterization traditionally relies on an initial screening that identifies potential scenarios of concern, which can be subsequently evaluated at a more targeted level. This involves comparing contaminant concentrations in the designated environmental matrix (e.g. soil, water) to generic quality targets that reflect in a possible significant risk, if surpassed (Weeks and Comber 2005; Merrington et al. 2006; Gonc¸alves et al. 2013). In a subsequent tier, ecological testing based on soil bioassays and aquatic tests using soil leachates, allows evaluating to which extent detrimental effects are acceptable at the organism and ecosystem levels (van Gestel et al. 2001, 2008; Loureiro et al. 2005a, b; Weeks and Comber 2005; Merrington et al. 2006). The derivation of toxicity (and contamination) thresholds relies on the use of standardized test procedures and reference soils, to ensure reproducibility, comparability, site-independence and general acceptance (Jensen and Pedersen 2006). This translates in accurate evaluation and
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extrapolation of risks to terrestrial ecosystems or aquatic bodies exposed to surface runoff or groundwater leaching, resulting in appropriate risk management and mitigation decisions (e.g. Loureiro et al. 2005a, b; van Gestel 2008). In this context, various standard substrates aiming at harmonizing ecotoxicological testing have already been developed by a number of European and international organizations, including the OECD artificial soil (OECD guideline no. 207 1984), the natural LUFA 2.2 soil (Speyer, Germany) and a set of European natural soils (EUROSOILS; Kuhnt et al. 1994). LUFA 2.2 soil is the most widely used natural standard and reference soil, due to large supply, representativeness and ecological relevance, thus ensuring no contamination and adequate performance of test biota (e.g. Domene et al. 2011; Ro¨mbke et al. 2006; Caetano et al. 2012). In light of the equilibrium partitioning theory and the concept of bioavailability, exposed organisms are affected by the metabolically available contaminant fraction, rather than by the total contaminant concentration in soil (ISO 17402:2008; Holmstrup et al. 1998; Frische et al. 2003; van Straalen et al. 2005; van Gestel 2008; Rocha et al. 2011). Natural soil spatial and temporal variability and the complexity of biotic and abiotic interactions in the soil ecosystem, pose a challenge to the development of approaches to effectively account and assess contaminant bioavailability in soils (e.g. Sijm et al. 2000; Frische et al. 2003; Hund-Rinke and Ko¨rdel 2003; Maliszewska-Kordybach et al. 2008). Nonetheless, there is consensus that waterextractable soil fractions can be regarded as potentially bioavailable, based on the relationship between observed toxicity and measured chemical concentrations in soil extracts (e.g. ISO 17402:2008; Kelsey et al. 1997; Frische 2002; Hund-Rinke and Ko¨rdel 2003). Additionally, waterbased extractions have the advantage of accounting for the exposure route (soil pore-water) to terrestrial organisms (e.g. Ma et al. 1998; Hund-Rinke and Ko¨rdel 2003; Lanno et al. 2004; van Gestel and Koolhaas 2004; ter Laak et al. 2006). The present study is a proof of concept and highlights the possibility of water-extractable components from the LUFA 2.2 standard soil to pose an additional stress and/or confounding factor to the test biota, when inferring biological responses to chemicals and toxicity thresholds. Firstly, we assessed the contribution of LUFA 2.2 soil as a direct potential source of priority contaminant groups (metals and the 16 PAHs) in ecotoxicological testing, by examining their release into aqueous soil solution, as influenced by consecutive soil wet–dry cycles. Such a procedure aimed at simulating a ‘worst-case scenario’ in respect to their potential bioavailability over time, being therefore in line with established approaches for characterization of soils and soil materials (Eisentraeger et al.
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2004; Jablonowski et al. 2012). Subsequently, a two-step screening of aqueous contaminant contents was performed based on current regulatory EU guidelines for surface waters (2008/105/EC) and reported ecotoxicological benchmarks for aquatic organisms (Suter II and Tsao 1996). By combining the information from both steps in a single strategy, we provide a straightforward and generic screening approach to identify a potential scenario of concern linked to the use of reference and standard natural soils, whose water-extractable fractions are not adequately characterized.
Materials and methods Soil characteristics The sandy loam LUFA 2.2 soil (\2 mm; LUFA Speyer, Germany) was the selected natural standard soil, with the following physico-chemical characteristics: pH (0.01 M, CaCl2): 5.5 ± 0.1; soil organic C (%): 1.93 ± 0.2; cation exchange capacity (100 cmol?/kg): 10.0 ± 0.8; sand (%): 81.3 ± 2.3; silt (%): 12.1 ± 1.3; clay (%): 6.60 ± 1.3; maximum water holding capacity (g/100 g): 45.2 ± 5.0; density (g/ml): 1.13 ± 0.045. Individual and total metal and PAHs contents in LUFA 2.2 soil are presented in Table 1. Soil wet–dry cycles and water-extraction The procedure used was a combination and adaptation of the methodologies described by Eisentraeger et al. (2004) and Jablonowski et al. (2012). Briefly, de-ionised water (1:2 mass:volume) was added to 100 g (oven-dry basis) of LUFA 2.2 soil (\2 mm) in triplicates and homogenized. Following agitation (overnight, 150 rpm, 20 ± 2 °C, dark) using a bench top orbital shaker, suspended soil was allowed to settle (15 min) and subsequently centrifuged (3000 rpm, 15 min; adapted from Eisentraeger et al. 2004). Collection of the supernatant was performed through vacuum-driven filtration using MilliporeÒ 0.45 lm paper filter, after which centrifugation of the supernatant (4000 rpm, 10 min) allowed clearing suspended organic and particulate matter. In turn, the wet soil residue was oven-dried (50 °C, 24 h) to its initial volumetric moisture content (8 % vol:vol), based on early experimental data (not shown). The described procedure constituted 1 cycle. Treatments corresponding to 6 and 12 cycles refer to 6 or 12 wetting–drying events prior to water-extraction, each initiating with fast rewetting of the oven-dry soil residue, obtained in the previous cycle (adapted from Jablonowski et al. 2012). While not in use, soil extracts were stored in glass containers, in the dark at 4 °C.
Water-extractable priority contaminants in LUFA 2.2 soil Table 1 Individual and total contents (dry weight) of metals and metalloids, as well as of the 16 priority PAHs in the reference soil Lufa 2.2, as a mean of duplicate measurements. When available, standard errors of means are given as ±SE
Parameters (lg/kg) Metals Cr
8831
Mn
139.0
Co
4.500
Ni
2452
Cu
1531
Zn
17220
As
11.60
Cd
140.1
Pb
15981
Hg P metals
46378
67.70
Screening approach Following characterization of the water-extractable LUFA soil fraction, a straightforward and generic two-step screening approach was performed. An initial screening of aqueous contaminant contents was performed based on current regulatory EU guidelines for surface waters (2008/ 105/EC), comparably to the first level in a tiered risk characterization of soils. To increase the ecological relevance of such an approach, a second screening step took place based on reported ecotoxicological benchmarks (Suter II and Tsao 1996), by which aqueous contents could pose an effective risk to aquatic biota in a non-specific testing scenario.
PAHs
NAP naphthalene, ACY acenaphthylene, ACE acenaphthene, FLU fluorene, PHE phenanthrene, ANT anthracene, FLT fluoranthene, PYR pyrene, CHR chrysene, BaA benz(a)anthracene, BbF benzo(b)fluoranthene, BkF benzo(k)fluoranthene, BaP benzo(a)pyrene, DBA dibenz(a,h)anthracene, BGP benzo(g,h,i)perylene, IND indeno(1,2,3-cd)pyrene
NAP
43.90 ± 3.59
ACY ACE
30.10 ± 0.33 29.30 ± 0.10
Results and discussion
FLU
31.70 ± 0.43
PHE
34.80 ± 0.23
Characterization and screening of the potentially bioavailable LUFA 2.2 fraction
ANT
22.40 ± 0.42
FLT
41.80 ± 0.94
PYR
43.20 ± 0.04
CHR
40.90 ± 0.11
BaA
15.30 ± 0.10
BbF
7.700 ± 0.15
BkF
6.800 ± 0.61
BaP
7.100 ± 0.24
DBA
6.400 ± 0.33
BGP
2.330 ± 0.41
IND P PAHs
366.0 ± 1.27
2.650 ± 0.13
Characterization of aqueous LUFA 2.2 soil extracts Extract characterization was performed in relation to pH (CaCl2) and dissolved organic carbon (DOC) using a TOC analyser. Total contents in the aqueous soil extracts were estimated for the 16 priority PAHs (by SPME-solid phase micro-extraction coupled to GC/MS-gas chromatography/ mass spectrometry) based on Campos et al. (2012), as well as metals and metalloids (by ICP/AES-inductively coupled plasma-atomic emission spectroscopy), namely cadmium (Cd), cobalt (Co), chromium (Cr), copper (Cu), manganese (Mn), zinc (Zn), nickel (Ni), lead (Pb) and arsenic (As), as described by Rodrigues et al. (2010). Mercury (Hg) was quantified by AAS-atomic absorption spectroscopy. Two replicate extractions were conducted and three replicate measurements were taken in each analysis (standard deviation between replicate measurements \10 %).
Table 2 summarizes the main physico-chemical characteristics of LUFA 2.2 soil aqueous extracts and compares aqueous metal and PAH concentrations with the EU water quality criteria and toxicological benchmarks for aquatic organisms. While aqueous metal concentrations were deeply influenced by the number of consecutive soil wet– dry events (177.87, 888.7 and 689.43 lg/l for 1, 6 and 12 cycles, respectively), extracted PAHs were largely independent of that (0.073, 0.071 and 0.064 lg/l for 1, 6 and 12 cycles, respectively). Metals and PAHs are ubiquitous in the environment. Background levels for such compounds are generally dictated by interactions in local geology, biogeochemistry, climate and anthropogenic activity (Reimann and Garrett 2005). It is not therefore, surprising the occurrence of such fractions in LUFA 2.2 aqueous extracts, which is in agreement with previous reports on background contaminant levels in LUFA’s solid matrix (e.g. \0.05 mg/kg of PAHs; Frische et al. 2003). Potentially bioavailable contaminant fractions in reference and standard substrates have been more scarcely addressed (exceptions include Loureiro et al. 2005a, b). The occurrence of such elements constitutes a challenge that should be addressed as part of the ecotoxicological testing approach, to ensure optimal performance of the test biota and accurate determination of toxicity thresholds. Aquatic regulations have been commonly considered in soil protection approaches, as discussed by Ferna´ndez et al. (2006). In the present study, aqueous concentration ranges were at least a factor of 10 below the EQS and MAC for metals and metalloids (Ni, Cd, Pb, Zn) as well as for the six PAHs (NAP, BaP, BbF, BkF, BGP, IND) regulated by the
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0.490
5.900
6.500
44.00 \1.000
0.270
Co
Ni
Cuc
Zn As
Cd
0.003 ± 1 9 10-4
0.006 ± 4 9 10-4
ANT
BaP
\LD
0.005 ± 6 9 10-4
BkF
a
\LD \LDa
0.004 ± 5 9 10-4
\LD
a
a
BbF
2.030 ± 0.33
\LDa
2.080 ± 0.15
1.160 ± 0.16
\LDa \LDa
\LDa \LDa
\LDa \LDa
CHR BaA
0.003 ± 3 9 10-5
0.002 ± 0.04
PYR
3.580 ± 0.02
2.400 ± 0.19
11.27 ± 1.45
\LD
a
5.490 ± 0.53
\LD
a
35.80 ± 5 9 10-3
689.4
0.130
1.300
0.100
18.00 4.000
8.300
10.00
1.300
643.0
3.300
103.0
0.007 ± 0.18
0.007 ± 9 9 10
FLT
-4
0.015 ± 8 9 10-4
0.011 ± 2 9 10-4
0.001 ± 3 9 10
\LD
PHE
-5
FLU
a
0.005 ± 1 9 10-4 \LDa
ACE
\LD
a
0.004 ± 3 9 10
-4
0.032 ± 4 9 10-3
888.7
177.8
0.027 ± 4 9 10-3
0.220
0.180
2.300
0.150
15.00 4.500
10.00
8.900
2.100
846.0
4.000
ACY
NAP
PAHs
Hgd P Metals
Pb
0.930
117.0
Mn
c
1.600
Cra
Metals
96.80
0.0500
R (BbF,BkF) = 0.030
2100 330.0
n/a
0.100
0.100
n/a
n/a
n/a
n/a
2.400–1.200
0.050
0.1000
n/a
n/a n/a
n/a
1.000
0.400
n/a
n/a
n/a
n/a
n/a
0.070
n/a
0.450–1.500
0.080–0.250c 7.200
n/a n/a
n/a n/a
n/a
20.00b
20.00b n/a
n/a
n/a
n/a
n/a
n/a
n/a
88.50
DOC (mg/l)
7.24
6.98
5.32
pH (0.01 M CaCl2)
MAC EQS
AA EQS
6 cycles
1 cycle
12 cycles
EU Directive for inland and other surface waters (lg/l)
Aqueous extract (lg/l)
n/a
n/a
33.60
n/a f
30.00f
n/a
80.00f
n/a
2.400
82.00
3.900c
1200 n/a
18.00
1400c
n/a
16.00
f
n/a
n/a
6.160
n/a
6.300f
n/a
23.00f
n/a
0.010
3.200
1.100c
110.0 n/a
12.00
160.0c
n/a
11.00
0.240
0.490
n/a
13.00
n/a
70.00
n/a
190.0
n/a
n/a
n/a
n/a 66.00
n/a
n/a
1500
2300
n/a
h
0.014
0.027
n/a
0.730
n/a
3.900
n/a
12.00
h
1.300e
n/a
n/a
n/a 3.100
n/a
n/a
23.00
120.0
n/a
Chronic
Acute
Acute
Chronic
Tier II (Secondary chronic) values
NAWQ Criteria
n/a
n/a
30.00
0.090
n/a
n/a
74.00
620.0
\0.23
18.88
1.700
36.41 892.0
3.800
\35.0
290.0
1780
73.18
Fish
g
0.300
0.650
15.00
\2.10
200.0
n/a
6.646
1.163
0.960
12.26
0.150
46.73 450.0
0.230
\5.00
g
g
g
g
n/a
n/a
54400
n/a
n/a
n/a
520.0
33000
5.000
500.0
2.000
30.00 48.00
1.000
5.000
2.000
n/a 5.100
2.000
6.132
Aquatic plants
\1100
Daphnids
Lowest chronic value for aquatic biota
Conventional toxicological benchmarks (lg/l) (Suter II and Tsao 1996)
Table 2 Screening of metals, metalloids and the 16 priority PAHs in aqueous extracts of standard LUFA 2.2 soil (after 1, 6 or 12 wet–dry and water-extraction cycles, 20 °C), based on the current EU regulation for surface waters (2008/105/EC) and the reported toxicological benchmarks (reviewed by Suter II and Tsao 1996)
A. C. Bastos et al.
0.0727
0.003 ± 1 9 10 0.003 ± 9 9 10 0.0713
0.064
\LD
a
R (BGP,IND) = 0.002 n/a
Chronic
Acute
Acute
Chronic
Tier II (Secondary chronic) values
NAWQ Criteria
Fish
Daphnids
Lowest chronic value for aquatic biota Aquatic plants
Conventional toxicological benchmarks (lg/l) (Suter II and Tsao 1996)
h
g
f
e
d
c
b
a
Estimated value, as described in Suter II and Tsao (1996) OSWER (Office of Solid Waste and Emergency Response) threshold values for aquatic life (U.S. EPA 1996)
Final acute and chronic values in derivation of sediment quality criteria (U.S. EPA 1993b)
Chronic NAWQC value for Hg (0.012 lg/l) is estimated based on final residue values (Suter II and Tsao 1996)
Values refer to inorganic Hg
Criteria dependent on water hardness and normalized to 100 mg/l
Value refers to Ni and its compounds
Indicated contents for As (V) and Cr(VI), which hold higher relevance in natural waters when compared to As (III) and Cr(III) (e.g. Economou-Eliopoulos et al. 2011)
NAP naphthalene, ACY acenaphthylene, ACE acenaphthene, FLU fluorene, PHE phenanthrene, ANT anthracene, FLT fluoranthene, PYR pyrene, CHR chrysene, BaA benz(a)anthracene, BbF benzo(b)fluoranthene, BkF benzo(k)fluoranthene, BaP benzo(a)pyrene, DBA dibenz(a,h)anthracene, BGP benzo(g,h,i)perylene, IND indeno(1,2,3-cd)pyrene, AA annual average, MAC maximum allowable concentration, NAWQ National Ambient Water Quality, U.S. water quality criteria, LD limit of detection
IND P PAHs
-5
0.003 ± 1 9 10-4 \LDa
-4
0.002 ± 4 9 10-4
BGP
\LDa
\LDa
\LDa
DBA
MAC EQS
AA EQS
6 cycles
1 cycle
12 cycles
EU Directive for inland and other surface waters (lg/l)
Aqueous extract (lg/l)
Table 2 continued
Water-extractable priority contaminants in LUFA 2.2 soil
123
A. C. Bastos et al.
EU Directive for inland and other surface waters (2008/ 105/EC). Exceptions included the sum BGP ? IND, which was twice as high (0.005 and 0.006 lg/l for 1 and 6 cycles) as the annual average EQS for inland surface waters (2 lg/l; 2008/105/EC), with no MAC available; the Hg contents (0.18, 0.22, 0.13 lg/l for 1, 6 and 12 cycles, respectively) have also slightly exceeded the recommended MAC levels for inorganic or total mercury (0.07 lg/l) for inland and other surface waters (2008/105/EC). Total As contents (1.0, 4.5 and 4.0 lg/l respectively), while not targeted in the above Directive, were within the established as ‘typical’ for background natural waters (Smedley and Kinniburgh 2002). When compared to generic quality guidelines, screening based on reported aquatic toxicology benchmarks (Suter II and Tsao 1996) on either single or combinations of multiple endpoints may be an equally valuable tool to assess the likelihood of an effective risk to the test biota (Maliszewska-Kordybach et al. 2008). Conventional benchmarks are grounded on regulatory standards or on the responsespecific endpoints used to derive them, such as the acute or chronic NAWQC, the SCV (estimated with 80 % confidence as to not exceed the NAWQC for compounds lacking such criteria) and the LCV (lowest acceptable chronic values, recommended by U.S. EPA as substitutes for water quality criteria) for aquatic organisms (U.S. EPA 1992, 1993a, b). In respect to aqueous metal and metalloid contents (independently of the number of wet–dry cycles), these were 1–2 orders of magnitude below the corresponding NAWQC, suggesting that a significant risk to the test (aquatic) species can be discarded. Nevertheless, some metallic compounds were detected at concentrations above the SCV or the LCV for aquatic organisms, while aqueous Pb contents (0.93, 2.30, 1.3 lg/l for 1, 6 and 12 cycles respectively) were closely comparable to the chronic NAWQ value (3.2 lg/l; derived from median lethal concentrations -LC50- and Chronic Values -CV-) for this compound (Suter II and Tsao 1996). For instance, As contents slightly surpassed the SCV for aquatic organisms (3.1 lg/l), after 6 and 12 wet–dry cycles. Cd and Ni generally exceeded the LCV for daphnids (0.15 and \0.5 lg/l, respectively), while Cr contents were above the LCV for aquatic plants (2 lg/l). Amongst the detected metallic compounds, Zn concentrations surpassed the LCV targets for fish (36.41 lg/l), non-daphnid invertebrates (5.24 lg/l) and aquatic plants (30 lg/l) after the first extraction. Having surpassed ‘lower benchmark’ values (U.S. EPA 1992, 1993a), it suggests that Zn, As, Cd, Ni and Cr may represent an additional stress and/or confounding factor when inferring the biological response to aqueous extracts of LUFA 2.2 soil (Suter II and Tsao 1996). This may occur preferentially under acidic to slightly acidic (\pH 6.0) extract conditions and/or increasing DOC levels, due to
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enhanced metal solubility and extractability (e.g. Chuan et al. 1996; Antoniadis and Alloway 2002). For the PAHs which toxicological benchmarks were revised by Suter II and Tsao (1996; ACE, ANT, BaA, BaP, FLU, FLT, NAP, PHE) concentrations were generally, a factor of 10–100 below the respective NAWQC, SV and LCV, albeit ANT contents being only marginally below the LCV for fish (0.09 lg/l). Overall, PAHs in LUFA 2.2 soil aqueous extracts, whilst individually, are not likely to constitute a hazard to aquatic biota exposed to LUFA 2.2 aqueous extracts (Suter II 1995). Contextualization and potential implications for testing The choice of a standard or reference soil is of prime importance in estimating ecotoxicological thresholds, providing grounds for accurately inferring or refining ecological risk, with consideration of both physico-chemical and biological criteria. Generally, the standard soil represents a substrate for spiking with contaminants of interest to derive soil and water screening values (as solid or aqueous soil extracts) or a negative control in dilutions or aqueous extracts of contaminated soils, soil materials or wastes in aquatic and terrestrial biological assessment (van Gestel et al. 2001; Loureiro et al. 2005a, b; Caetano et al. 2012). For example, the derivation of soil and water quality criteria as well as predicted no-effect concentrations (PNECs) is grounded on standard bioassays or obtained by statistical extrapolation (EC 1996). The contaminant background concentration in the standard substrate can therefore influence that process, despite quantification of bioavailable fractions, often being poorly addressed. Either of such procedures generally also involves consideration of additional factors, including toxicity-dependent parameters, such as contaminant bioavailability and biomagnification potential, as well as species-specific sensitivities (e.g. van Gestel et al. 2001; Ferna´ndez et al. 2006). Therefore, questioning the suitability of the standard substrate due to background noise and/or added stress to the test biota, can lead to compromising the integrity of estimated toxicity thresholds. In the same way, it can confound data interpretation and limit reproducibility in tests and comparability between studies. It further adds to the already-challenging derivation of critical levels for specific contaminants, reference doses and contamination thresholds in the selected soil and aqueous matrices. Ultimately, if not adequately addressed, it can lead to inappropriate evaluation and prioritization of risks relating to contaminated soils, and consequently, to poor decision-making for risk management or mitigation (Suter II 1995). To the best of our knowledge, this is the first comprehensive characterization of water-extractable priority contaminant groups in a widely used natural standard and
Water-extractable priority contaminants in LUFA 2.2 soil
reference soil, which are generally not accounted for in risk characterization approaches and experimental set-ups. The procedure used was a simple and straightforward approach, which served to demonstrate the importance of considering the potentially bioavailable LUFA 2.2 fraction over time and the possibility of long-term leaching and runoff of contaminant fractions, as influenced by a natural soil process (consecutive dry-wet events). It further considers that: (i) such elements have dissimilar interactions with, and desorption rates from the soil matrix; (ii) there may be relative variations in background concentrations in the aqueous extracts to which aquatic test biota are exposed; (iii) aqueous concentrations may represent pore-water relevant exposures to terrestrial organisms, as influenced by soil wetting and drying. Based on comparison to generic water quality standards and available aquatic ecotoxicological data, screening of LUFA 2.2 water-extracts has helped identifying a potential conceptual bottleneck, in the context of tiered risk assessment and environmental quality criteria derivation procedures. It suggests that potentially bioavailable fractions should be adequately addressed in view of one’s research aim, experimental design and scope of outcomes. Further, it proposes a simple, yet conservative procedure that could be adapted to match specific assessment requirements and goals, in order to maximize the level of robustness and confidence in the estimated ecotoxicological criteria, independently of the standard soil used. The present study and output recommendations are intended to encourage discussion on how risk evaluation and data generation can cope with, adapt to and overcome the use of natural soils, in view of the ERA framework and current European Directives for Soil [COM (2006) 232] and Water (2008/105/EC) protection. Rather than undermining the use of LUFA 2.2 soil as a natural standard and reference substrate, it highlights the necessity to fully characterize any standard soil’s water-extractable fraction to include and help manage sub-optimal traits related to its inherent variability (Jensen and Pedersen 2006). In this way, the experimental design and considerations can be adjusted to specific study requirements, endpoints and goals, to improve relevance and refinement. Components that are more likely to contribute the most (or the least) to a biological response can be easily identified and ranked, whereas added stressors that can compromise the test biological integrity can be excluded or adequately accounted for. Most of the present discussion can also be applied to terrestrial test organisms exposed via the water pathway, by ingestion of suspended particles in soil pore water (EC 2003). Nevertheless, there are limitations for the described approach in the context of derivation of terrestrial ecotoxicological data. Despite the reasoning behind the
described extraction procedure overlapping to some extent with the concept of soil pore-water (Eisentraeger et al. 2004), aqueous contaminant contents and implications to terrestrial biota may not extrapolate straightforwardly and must be considered critically (Sijm et al. 2000; Frische et al. 2003; Hund-Rinke and Ko¨rdel 2003). Similarly, while addressing wet–dry cycles can account for long-term desorption as influenced by natural soil processes, long time-scales are not generally covered by conventional acute and chronic ecotoxicological tests. Estimated waterextractable metal and PAH fractions in this study should be regarded only in the context of the present discussion. Further, a certain degree of variability between LUFA 2.2 soil batches is expected, including pH, clay and organic matter contents, all of which are key parameters in controlling chemical availability of metals and organic compounds in soils (Jensen and Pedersen 2006; van Gestel 2008).
Conclusions Using a simple and generic screening approach to address and contextualize potentially bioavailable contaminants in LUFA 2.2 soil, this study highlights the need for suitably characterizing its water-extractable fractions. Potentially bioavailable fractions for priority metals and PAHs in LUFA 2.2 soil were generally in agreement with recommended EQS for surface waters (except Hg and the sum BGP ? IND) and the reported NAWQ criteria. Whilst individually, PAHs are not likely to constitute a hazard to the test biota exposed to LUFA 2.2 aqueous fractions. Nevertheless, As, Zn, Cd, Ni, Cr are among the metals which concentrations exceeded lower ecotoxicological benchmarks (SCV or LCV) for aquatic organisms. This study urges for potentially bioavailable fractions of the natural standard or reference soil to be adequately characterized and addressed for derivation of ecotoxicological data and contamination thresholds, according to the study aim, experimental approach and design, as well as the expected scope of outcomes. Acknowledgments This study was supported by European Funds through COMPETE and National Funds through the Portuguese Science Foundation (FCT), within projects PEst-C/MAR/LA0017/ 2013, FUTRICA (FCOMP-01-0124-FEDER-00008600) and CLIMAFUN (FCOMP-01-0124-FEDER-008656), which included the post-doctoral fellowships of Ana C. Bastos (BPD/CESAM/PTDC/ AAC-AMB/104666/2008) and Jacinta M.M. Oliveira (BPD/UI88/ 6463/2013), respectively. The authors also wish to acknowledge the FCT-funded post-doctoral fellowships of Carla F. Calhoˆa (SFRH/ BPD/74232/2010) and Miguel J.G. Santos (SFRH/BPD/72380/2010), as well as the European Commission through the Erasmus Mundus Program for the MSc Grant to Marija Prodana. Amadeu M.V.M. Soares is ‘Bolsista CAPES/BRASIL’, Project No. A058/2013.
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A. C. Bastos et al. Conflict of interest Authors declare that they have no conflict of interest.
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