Abstract

Evaluation of the bioavailable fractions of organic contaminants such as polycyclic aromatic hydrocarbons (PAHs) is extremely important for assessing their risk to the environment. This available fraction, which can be solubilised and/or easily extracted, is believed to be the most accessible for bioaccumulation, biosorption and/or transformation. Sediment organic matter (OM) and clay play an important role in the biodegradation and bioavailability of PAHs. The strong association of PAHs with OM and clay in sediments has a great influence not only on their distribution but also on their long-term environmental impact. This paper investigates correlations between bioavailability and the clay and OM contents in sediments. The results show that OM is a better sorbent for pyrene (chosen as a model PAH) and that increasing the OM content reduces the bioavailable fraction. A mathematical model was used to predict the kinetic desorption, and these results showed that the sediment with the lowest content of OM had an Ffast value of 24%, whereas sediment with 20% OM gave a value of 9%. In the experiments with sediments with different clay contents, no clear dependence between clay and rate constants of the fast desorbing fractions was observed, which can be explained by the numerous possible interactions at the molecular level.

INTRODUCTION

Polycyclic aromatic hydrocarbons (PAHs) are organic contaminants with toxic, mutagenic and carcinogenic properties derived from natural sources and human inputs, which occur widely in coastal environments. Due to a high hydrophobicity and stable structure, PAHs are easily accumulated in sediment. The strong sorption of PAHs onto sediment particles could reduce their biodegradation rates and preserving them in sediment (McGroddy et al. 1996). Thus, the affinity of hydrophobic organic compounds for biotic and abiotic phases is an important determinant of both the rate of a lake's detoxification and its response time to changing loadings. PAH–sediment–water interactions occurring within a watershed and the associated aquatic system have been studied in a lot of research papers. The sorption of an organic chemical on a natural solid is a very complicated process described by a lot of authors, which involves many sorbent properties, besides the physicochemical properties of the chemical itself (Karickhoff 1984; Pignatello & Xing 1995). Also, different regions of a soil or sediment matrix may contain different types, amounts, and distributions of surfaces and of soil organic material, even at the particle scale (Weber et al. 1992). To model the biodegradation process, a non-linear model, namely the Monod equation, is commonly used to describe the kinetics of biodegradation of organic compounds such as PAHs (Knightes & Peters 2000; Dimitriou-Christidis & Autenrieth 2007; Liu et al. 2007). Based on the Monod equation, some other models were also developed for the PAH biodegradation process (Dimitriou-Christidis & Autenrieth 2007; Liu et al. 2007; Spasojević et al. 2015). A complicated model, including the microbial biodegradation activity of the PAHs in the aqueous phase and PAH sorption kinetics with respect to the organic carbon content, was developed by Artola-Garicano et al. (2003). The major difficulty when dealing with the concept of bioavailability is that there are many definitions and several methods to measure and to calculate it, depending on the specific scientific discipline (Ehlers & Loibner 2006; Reichenberg & Mayer 2006). Indeed, bioavailability depends on the physical, chemical and biological properties of contaminants, soil and receptors, and it is governed by three way interactions between contaminants, matrix and organism. Therefore, three distinct processes are involved: physicochemical, physiological uptake and toxicological. The physicochemical processes, which have been extensively discussed in recent years (Ehlers & Loibner 2006), include sorption, diffusion and partitioning and are controlled by soil and compound properties such as soil organic matter (SOM) content and quality, soil inorganic constituents and lipophilicity of compounds. The possible conditions of sorption/desorption process include the presence and nature of the natural sorbent mineral clay and especially organic matter (OM), as it was defined above. It is also affected by pH values, temperature, binding to dissolve OM, cosolvent effect, sorbent concentration, etc. For example, studies by Podoll et al. (1989) reported that the adsorption of naphthalene on soil decreases with increasing temperature from 15 to 50 °C and the isosteric enthalpy of adsorption is exothermic, while He et al. (1995) studied the sorption of fluoranthene on soils and lava. The sorption coefficients of fluoranthene were found to decrease with temperature between 5 and 25 °C. The relation between OM and clay is described in this paper on the molecular level, explained with Van der Waals forces (VdW), charge-transfer (CT), solvent-driven interactions, etc. Also the study conducted by Garcia-Falcon et al. (2006) reported that pH values have an effect on PAH removal and transfer. The physiological uptake processes depend on receptor type and specific parameters such as anatomy, feeding strategy or the lipid content of the organism, whereas toxicological processes are controlled by metabolism, detoxification or accumulation capacity. Experimental studies on the fate of persistent organic pollutants in environmental media are often conducted in the laboratory. Researchers use semi controlled conditions to follow specific processes of interest by maintaining control over appropriate factors/conditions (i.e., humidity, light, and temperature). This is especially the case when chemicals are being evaluated with respect to partitioning and sorption/desorption behavior, persistence, toxicity, bioavailability, and biodegradability. Experiments such as these often involve the study compound being introduced into the sample by various methods that are generically described as spiking. Spiking is defined by the American Society for Testing and Materials (ASTM) as ‘the experimental addition of a test material such as a chemical or mixture of chemicals, sewage sludge, oil, particulate matter, or highly contaminated sediment/soil to a clean negative control or reference sediment/soil to determine the toxicity of the material added. After the test material is added, sometimes with a solvent carrier, the sediment/soil is mixed to evenly distribute the test material throughout the sediment/soil’ (Northcott & Jones 2000).

Pyrene is a high molecular weight, four-ring PAH, and an EPA Priority Pollutant. Pyrene's release to the environment is ubiquitous, since it is a ubiquitous product of incomplete combustion. Because of concerns about their potential environmental impacts, EPA has made a list of priority pollutants, and pyrene is a part of that list. Particulate pyrene is well correlated with total PAH, because of which it is usually used as a model compound for other PAH compounds' behavior. Evaluation of the concentration of PAHs in soil samples was conducted by Garcia-Falcon et al. (2006). They reported higher PAH concentration in burnt soil versus unburnt soil. PAH levels in the 1–5 cm layer of the burnt soil remained more or less constant during the first 3 months after burning and then, over the following 7 months, fell to about 57% of their peak level. This decline was explained by PAH being transported to lower soil layers by rainfall, and increased mobilization of PAHs associated with dissolvable OM. Pontevedra-Pombal et al. (2012) and Rey-Salgueiro et al. (2009) reported that high levels of fluoranthene and pyrene were present in peat samples dating back to the 12th century A.D. The results in this study suggest that changes in sources, type of emission (global or local) and transport could be responsible for the different PAH content and composition of the sample. Also, although PAHs are formed during the combustion process and high concentrations were measured in fumes and particles emitted from wood fires, the ash itself was seldom studied, by Rey-Salgueiro et al. (2004) and Garcia-Falcon et al. (2004a). The fate of PAHs when they are released directly in the environment is largely associated with particulate matter of soils and sediments, and in that way it can be transported for considerable distances (Garcia-Falcon et al. 2004b). High concentration of organic pollutants, like pyrene, distant from primary sources indicates that it is reasonably stable in the atmosphere (Irwin 1997). When pyrene is found in a water/sediment system, it tends to adsorb very strongly to sediments and particulate matters, bioconcentrate in aquatic organisms slightly to moderately, but will not hydrolyze (Irwin 1997). Recent experimental studies that use chemical extraction methods in order to predict the bioavailability of organic pollutants in the environment instead of using living organisms showed that they can correlate with the uptake or degradation by a specific organism (Amezcua-Allieri et al. 2012; Ortega-Calvo et al. 2013; Barnier et al. 2014). Rey-Salgueiro et al. (2008) found that manures can be used for detection and quantification of PAHs and 3-hydroxybenzo[a]pyrene in animal husbandry. Results showed that all samples of manure didn't have alarming PAH concentration. However, 3-OH-B[a]P was quantified in all the samples except rabbit manure. In horse and cow manure, the highest total PAHs and the hydroxymetabolite were determined. These two animals graze outside every day. They intake, also, soil by grazing, which bounds high PAH levels. Differences between PAH levels in cows and horses could be explained by the microbiota in the horse gut, which are able to metabolize B[a]P at higher rates. Studies on the application of XAD4 resin for pyrene desorption and bioavailability assessment from sediments with different clay and organic contents are very limited, with only a few papers available to our knowledge. The aim of this study was to model rhw bioavailability of pyrene in artificially contaminated sediment depending on the content of clay and OM, using XAD4 resin as the desorption sorbent.

EXPERIMENTAL

Chemicals

Amberlite® XAD4 was obtained from Fluka, and all organic solvents were at least analytical grade and were purchased from Merck or J.T. Baker. Pyrene was obtained from Sigma-Aldrich.

Spiking and sediment characteristics

To take undisturbed sediment samples according to the standard method for sediment SRPS ISO 5667-12:2005, a corer sampler (Beeker, Eijkelkamp, The Netherlands) was used. Sediment samples were collected in appropriate containers. All samples were brought to the laboratory and stored at 4 °C until the moment of preparation of the samples for analysis. Sediment was air dried for four weeks and then passed through a 2 mm sieve, spiked with pyrene, homogenized and left to age. Subsamples were taken for the chemical analysis and other experiments after three months. The sample characteristics are given in Table 1. The contents of OM and clay were determined by the loss of ignition at 550 °C (SRPS EN 12879:2007) and by using the wet sieving method (ISO 11277:2009), respectively. Pyrene was extracted from sediments using ultrasound technique EPA3550b (US EPA 1996a) with a mixture of acetone/hexane (1:1). Elemental sulphur was removed by copper according to EPA 3660b (US EPA 1996b). Samples were fractionated on silica gel by the EPA3630c procedure (US EPA 1996c). GC/MS analysis was performed according to method EPA8270C (US EPA 1996d), on an Agilent 7890 gas chromatograph with a MSD 5975C mass spectrometer on an HP-5MS column (J&W Scientific), using phenanthrene-d 10 as an internal standard and the following conditions: pulsed splitless mode with a split ratio of 50:1, inlet temperature 300C, column 34.5 ml/min, initial oven temperature 55C for 1 min, then 25C min−1 to 300C for 3 min in splitless mode. The PAH concentrations were calculated by the internal standard method using target ion peak areas. The method detection limit for pyrene was 3.97 μg/kg.

Table 1

General properties of the sediments with different OM and clay contents

  Pyrene (μg/kg) Organic matter (%) Clay (%) 
Som1 3,106 8.23 30.3 
Som2 3,689 14.9 33.1 
Som3 3,533 17.2 33.3 
Som4 2,988 20.2 30.2 
Sc1 2,417 9.68 12.6 
Sc2 2,154 9.71 22.2 
Sc3 2,191 9.87 29.0 
Sc4 2,613 9.27 40.4 
  Pyrene (μg/kg) Organic matter (%) Clay (%) 
Som1 3,106 8.23 30.3 
Som2 3,689 14.9 33.1 
Som3 3,533 17.2 33.3 
Som4 2,988 20.2 30.2 
Sc1 2,417 9.68 12.6 
Sc2 2,154 9.71 22.2 
Sc3 2,191 9.87 29.0 
Sc4 2,613 9.27 40.4 

Desorption determination by XAD4 extraction

The methods of Cornelissen et al. (1997) for measuring desorption of non-polar organics in distilled water with Amberlite® XAD4 was used with some modifications (Spasojević et al. 2015). Preparation of resin: the XAD-4 resin was purified by Soxhlet extraction with ultra-pure water; methanol; a mixture of hexane/acetone 1:1; methanol and finally ultra-pure water. Each extraction took 6 h. After that, dry contaminated sediments (about 1 g), 1 ml HgCl2 (0.02%), 20 ml 5 mM KCl and 0.2 g of XAD4 resin were placed in 40 ml vials. After 2 h, 4 h, 6 h, 24, 48 h, 96 h, 144 h and 216 h, 0.8 g K2CO3 was added to each sample in order to promote flocculation, the supernatant was decanted and PAH analysis was conducted. At the same time, the XAD4 resin was separated and a fresh amount of resin was added to the vials. The separated XAD4 was extracted three times with acetone/hexane 1:1. A simultaneous blank analysis without XAD4 was conducted.

Desorption data modeling

The desorption of PAHs from soils and sediments has been described by the following first-order kinetics (Cornelissen et al. 1998; Barnier et al. 2014):  
formula
(1)
where St corresponds to the amount of PAHs sorbed to the sediment at desorption time t (h) and S0 is the total amount of sediment-associated PAHs immediately prior to desorption (obtained by sample oxidation). Ffast and Fslow (%) are the rapidly and slowly desorbing fractions, respectively, and kfast and kslow (h−1) are the corresponding rate constants of rapid and slow desorption. The model assumes that kslow is significantly less than kfast. Values of Ffast, Fslow, kfast and kslow were determined by exponential non-linear curve fitting.

RESULTS AND DISCUSSION

General properties of sediments

The sediment properties are summarized in Table 1. The first group of sediments, named Som1 –Som4, had similar clay contents and different OM contents, while sediments Sc1 –Sc4 had similar values of OM and different clay contents. These materials were chosen in order to determine the existence of a connection with OM and clay content with the bioavailability of pyrene. It is well known that numerous factors affect the sorption of organic pollutants by natural sorbents/soils, sediments, clays, humic materials, the dissolved OM and the sorption coefficients of selected pollutants (Delle Site 2001). Measured values for the first set of sediments varied from 8 to 20% OM and about 30% clay. For the second set of sediments, clay percentages varied from 12 to 40% and OM was about 9.5%.

Desorption kinetics and data modeling

The desorption kinetics of pyrene on natural sorbents such as clay or OM show a rapid initial phase followed by a slow approach to achieve equilibrium (Figure 1). In order to apply non-exhaustive extraction procedures, the time required to achieve equilibrium between the pyrene in water and pyrene in sediment has to be determined. Previous studies have determined that approximately 24 hours is enough to achieve equilibrium (Cornelissen et al. 1997; Spasojević et al. 2015). The shape of the obtained desorption curves (Figure 1) was similar to those observed by previous authors using the same or similar mild extraction techniques (Hawthorne et al. 2002; Spasojević et al. 2015). Kinetic curves were modelled using the two-fraction model, and they gave good correlation with desorption kinetic Equation (1), with the model fitting well to our data.

Figure 1

Desorption kinetics of pyrene from the sediments with different OM contents by XAD4 resin.

Figure 1

Desorption kinetics of pyrene from the sediments with different OM contents by XAD4 resin.

First-order kinetic reaction for sediment with different OM contents

The fitted rate constants and desorbing fractions of the two fraction model are given in Table 2. The rate constants of the rapidly desorbing fractions in different sediments (kfast) varied between 0.17 h−1 (for sediment Som2 and Som3) and 0.29 h−1 (for Som1 sediment). The rate constants of the slowly desorbing fractions (kslow) were roughly 1,000 times smaller than kfast, confirming the assumptions of the model. The value of the rapidly desorbing fractions for pyrene showed good correlation with OM content. In the sediment with the lowest OM content, the value of Ffast is 24% and sediment with 20% OM gave a value of 9%. This can be explained by the fact that OM can sorb organic pollutants (including pyrene) and thus decrease their bioavailability.

Table 2

Kinetic parameters for the two-fraction model for sediment with different OM contents

  Kfast (h−1kslow (h−1Fslow (%) Ffast (%) R2 
Som1 0.29 1.56 × 10−4 75.5 24.5 0.990 
Som2 0.17 5.82 × 10−5 87.0 13.0 0.955 
Som3 0.17 1.45 × 10−4 90.0 10.0 0.999 
Som4 0.22 6.28 × 10−5 91.0 9.00 0.976 
  Kfast (h−1kslow (h−1Fslow (%) Ffast (%) R2 
Som1 0.29 1.56 × 10−4 75.5 24.5 0.990 
Som2 0.17 5.82 × 10−5 87.0 13.0 0.955 
Som3 0.17 1.45 × 10−4 90.0 10.0 0.999 
Som4 0.22 6.28 × 10−5 91.0 9.00 0.976 

First-order kinetic reaction for sediment with different clay contents

The values obtained for the fast and slow desorbing fractions for pyrene using XAD-4 resin are given in Table 3. Figure 2 presents the modeled curves for these samples.

Table 3

Kinetic parameters for the two-fraction model for sediments with different clay contents

  kfast (h−1kslow (h−1Fslow (%) Ffast (%) R2 
Sc1 0.26 7.8 × 10−4 50 50 0.994 
Sc2 0.07 2.0 × 10−3 52 48 0.995 
Sc3 0.07 1.0 × 10−3 62 38 0.996 
Sc4 0.15 9.0 × 10−4 55 45 0.995 
  kfast (h−1kslow (h−1Fslow (%) Ffast (%) R2 
Sc1 0.26 7.8 × 10−4 50 50 0.994 
Sc2 0.07 2.0 × 10−3 52 48 0.995 
Sc3 0.07 1.0 × 10−3 62 38 0.996 
Sc4 0.15 9.0 × 10−4 55 45 0.995 
Figure 2

Desorption kinetics of pyrene from the sediments with different clay contents by XAD4 resin.

Figure 2

Desorption kinetics of pyrene from the sediments with different clay contents by XAD4 resin.

The rate constants of the rapidly desorbing fractions in different sediments (kfast) varied between 0.07 h−1 (for sediment Sc2 and Sc3) and 0.26 h−1 (for Sc1 sediment). The rate constants of the slowly desorbing fractions (kslow) were roughly 1,000 times smaller than kfast, confirming the assumptions of the model. The highest value of the rapidly desorbing fractions was obtained in sediment Sc1, where the percentage of clay was 12.6%. Values of Ffast for those sediments were bigger than the values obtained in the experiment with sediment with different OM contents.

Hydrophobic aromatic compounds such as pyrene have a strong capacity to absorb to minerals and organic soil and sediment components due to their pi-systems. H bonding strengths appear to be lower than the bonding strengths of cation-pi interactions and pi-pi electron donor-acceptor (EDA) interactions and comparable to inner- and outer-sphere complex formation (Keiluweit & Kleber 2009). Increasing the polarizability of the compound or structure increases the donor and acceptor strength involved (Mackay et al. 1997; Mackay & Callcott 1998). Strong permanent quadrupoles are the result of delocalized resonance effects, which occur throughout the fused π-systems for highly polarizable joint aromatic rings as PAHs. Hence, the compounds become forceful π-donors as the number of associated rings rises.

Cation-π interactions

Electrostatically, this can be conceptualized as the interaction of a positively charged ion and the negative electrostatic potential surface of the ring. It is conceived that these types of association are supported by dispersive and hydrophobic forces. In biochemical processes the significance of these interactions, with bonding strengths comparable to cation–water complexes, has been widely recognized. The first hypothesis of direct cation-π bonding between PAHs and mineral surface cations was an attempt to model surface interactions of certain PAHs. Thus to modern theory, PAH adsorption to minerals is extensively driven by hydrophobic interactions. However, Zhu et al. (2004) made a contradictory observation, whereby adsorption of phenanthrene to Ag+-exchanged montmorillonite was 10 times stronger compared with 1,2,4,5-tetrachlorobenzene (TCB) sorption, despite very similar Kow values. This observation was interpreted to indicate the existence of specific interactions between the sorbate and mineral surface. Zhu et al. (2004) examined this phenomenon under environmental (i.e., at least partly aqueous) conditions. Obvious distribution coefficients, log Kd, of different PAHs were higher to minerals saturated with weakly hydrated cations (Cs+, Ba2+, and especially Ag+) comparatively to minerals saturated with strongly hydrated cations. They also observed more upward-curving isotherms for PAHs with stronger π-donating properties, especially compared to the non-π-donor TCB. The strong effect of the type of cation and the π-donating properties of the solute suggested cation-π interactions (Zhu et al. 2004).

A considerable number of studies support the fact that models for PAH-mineral interactions should combine nonspecific effects of entropic forces with more specific cation-π interactions. Information on cation-π interactions in natural environments is missing, although there is a reliable platform to hypothesize that cation-π interactions occur in the cited contributions. Therefore studies should focus on the experimental substantiation of the mechanism with sorbent systems representative of terrestrial environments. Additional caution should be given to the determination of bonding enthalpies, with respect to the high potential of bonding strength.

π-π EDA interactions are particular, noncovalent forces of attraction between π-donor and -acceptor molecules (Meyer et al. 2003). π-π EDA interactions have been suggested as relevant for the adsorption of π-donating contaminants such as PAHs to π-accepting sites in organic soil components. Zhu et al. (2004) and Wijnja et al. (2004) systematically examined the nature of the π-π EDA interactions between PAHs and model humic materials. Molecular π-π EDA complexes may be characterized by a parallel-planar but slightly offset orientation of the π-donor and -acceptor systems, consentaneous to results of their examinations. This understanding enables the two complementary quadrupoles to interact, with electrostatic forces providing the majority of the contribution (Hunter & Sanders 1990). To a lesser level, attractive contributions such as VdW, CT, and solvent-driven interactions may be involved (Hunter et al. 2001). Lower interactions are shown by non-π-donors of similar hydrophobicity (in terms of Kow and Sw) (Wijnja et al. 2004). The interactions are generally promoted by inhibiting the potential for hydrophobic effects, either by reducing the water content of the solvent or raising acceptor polarity (Wijnja et al. 2004). Previously studies present that π-π EDA interactions are increased in apolar, nonaqueous solvents. Increasingly polar (i.e., less hydrophobic) acceptor molecules have reduced capacities to serve as a template for hydrophobic effects. Arias-Estevez et al. (2007) reported that the key factors controlling the distribution of PAHs in colloidal dispersions of humic substances in water, and therefore their stability-persistence and possible bioavailability in the natural aqueous environments, were proved to be the hydrophobicity and the water solubility of PAHs. Study by Xu et al. (2006) demonstrates, however, that π-π EDA interactions tightly require previous trapping of the aromatic sorbate in proximity to the π-acceptor domains via partitioning processes. Hence, Xu et al. (2006) propose a ‘host-guest interaction’ as the adsorption mechanism, in which molecules partition into the hydrophobic sphere, attach through dispersive forces, and then enlist in π-π EDA interactions. Multiple bonding mechanisms are expected to operate simultaneously for aqueous environmental systems. The estimation of actual binding strengths of π-π interactions is further complicated by the fact that host-guest interactions require several steps (Wijnja et al. 2004; Zhu et al. 2004). Detectable upfield shifts for protons (1H NMR) are the result of the parallel orientation of the two π-systems in π-π interactions, which leads to characteristic ring current effects on nuclei (Viel et al. 2002). A typical CT absorption band can be detected in the UV/vis signal if CT occurs, that is, a certain grade of electron-transfer between donor and acceptor. This previously mentioned evidence points to π-π EDA interactions as a significant binding power between compounds with aromatic π-donor or -acceptor systems and their functional doublet existing in natural organic matter. Based on this discussion, we can conclude that the natures of interaction on the molecular level have great importance to the fate of pyrene in sediment. Also, there are some interactions in the sediment that affect the behavior of pollutants. It is a complex system, so further continuation of studies is needed.

CONCLUSION

The bioavailability of organic compounds is conditioned by the type and intensity of interactions in the sediment/water system, and when it comes to hydrophobic organic compounds, sorption to sediment OM is the most important interaction. OM and clay play important roles in the processes that influence the behavior of organic pollutants in complex aquatic sediment systems. Different types of interaction exist at the molecular level, so it is not easy to explain these phenomena. Determination of the potential for desorption of organic compounds with the sediment is critical for assessing the risk of expressing negative ecotoxic effects in aquatic environments and to assess potentially contaminated samples. Our study demonstrated that the OM content and Ffast can be correlated and that increasing the OM content reduces the bioavailable fraction. However, no correlation between clay contents and rate constants of the fast desorbing fractions was observed during the experiments with sediments with different clay contents.

ACKNOWLEDGEMENTS

The authors gratefully acknowledge the support of the Provincial Secretariat for Higher Education and Scientific Research, Autonomous Province of Vojvodina (Project No. 114-451-2263/2016-02).

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