The exceptional sorptive ability of carbon nanomaterials (CNMs) for hydrophobic organic contaminants (HOCs) is driven by their characteristically large reactive surface areas and highly hydrophobic nature. Given these properties, it is possible for CNMs to impact on the persistence, mobility and bioavailability of contaminants within soils, either favourably through sorption and sequestration, hence reducing their bioavailability, or unfavourably through increasing contaminant dispersal. This review considers the complex and dynamic nature of both soil and CNM physicochemical properties to determine their fate and behaviour, together with their interaction with contaminants and the soil microflora. It is argued that assessment of CNMs within soil should be conducted on a case-by-case basis and further work to assess the long-term stability and toxicity of sorbed contaminants, as well as the toxicity of CNMs themselves, is required before their sorptive abilities can be applied to remedy environmental issues.
With the continued upscaling of carbon nanomaterial (CNM) production (Nowack and Bucheli, 2007) as well as the diverse array of consumer (Sharma and Ahuja, 2008), medical (Peretz and Regev, 2012) and industrial applications in which they are increasingly becoming incorporated, widespread environmental release of these physically and chemically unique macromolecules has become inevitable (Köhler et al., 2008). Once released, soils are likely to be a primary repository (Mueller and Nowack, 2008; Gottschalk et al., 2009), with the quantities anticipated to increase on an annual basis (Gottschalk et al., 2009). In spite of this, studies focused on CNMs within soils are scarce, and many areas of uncertainty remain. Understanding the interactions between CNMs, soils and components therein is therefore an urgent and essential aspect of any risk assessment process.
In their pristine form, CNMs are broadly characterised by their large reactive surface areas, highly hydrophobic characteristics and high degree of biogeochemical recalcitrance. They are known to be toxic to various soil microbiota (Riding et al., 2012a, b), and possess a high affinity for the sorption of a range of hydrophobic organic compounds (HOCs), such as polycyclic aromatic hydrocarbons (PAHs), and polychlorinated biphenyls (PCBs) (Pan and Xing, 2010). As both PAHs and PCBs are important classes of hydrophobic, toxic organic compounds, which are both abundant and persistent in soils (Stokes et al., 2005), the potential for CNMs to modify the availability and mobility of HOCs, either favourably through sorption and sequestration, or unfavourably through increasing contaminant dispersal, is currently unknown. Presently, there is only limited and occasionally contradictory information regarding the implications of contaminants while sorbed to CNMs, as well as the fate and behaviour of CNMs in uncontaminated soils. Exploring these issues in light of the emerging nature of CNMs as xenobiotic soil components is therefore essential.
This review seeks to answer three key questions. (i) What factors influence the behaviour and fate of CNMs within the soil environment? (ii) To what extent can CNMs influence the sorption, desorption and mobility of contaminants in soils? (iii) What are the impacts of CNMs on soil microorganisms and the biodegradation of contaminants in soils?
Within the environment, some CNMs can occur naturally or have close naturally occurring relatives due to various environmental events (Heymann et al., 1994; Chijiwa et al., 1999; Velasco-Santos et al., 2003; Esquivel and Murr, 2004). However, concentrations occurring naturally are likely to be relatively small (0.1–0.2 parts per million) (Heymann et al., 1994; Chijiwa et al., 1999). Therefore, when referring to CNMs, this review explicitly focuses on those that are anthropogenic in origin.
The properties of CNMs vary dramatically between the different methods of
production, functionalization status and cleaning/purification methods
employed (Nowack and Bucheli, 2007). Hence, determining their environmental
behaviour is all the more challenging, and generalisation of the
characteristics of CNMs is not possible, with each type requiring careful
characterisation (Nowack and Bucheli, 2007). Of the many different forms of
CNMs available, this review focuses specifically on carbon nanotubes (CNTs)
and C
To date, CNTs are arguably the most promising of all nanomaterials produced
(Giles, 2006). In their pristine form, CNTs are extremely hydrophobic and
consist of graphene sheets rolled into nanoscale diameter cylinders, the
ends of which may contain spherical fullerene cappings (Mauter and
Elimelech, 2008). One single-rolled graphite sheet is called a single-walled
carbon nanotube (SWCNT), while several SWCNTs nested together in a
concentric fashion comprise a multi-walled carbon nanotube (MWCNT) (Pan and
Xing, 2008). They consist of sp
Fullerenes are spherically arranged carbon atoms resembling a geodesic dome.
The size of the fullerene dome can vary depending on the number and
spherical configuration of carbon atoms. C
Unlike most organic chemicals with well-defined structures, the diversity of particle sizes, lengths, diameters, charges, surface areas, coatings, molecular weight, impurities and aggregation states are not necessarily constant. These are often tailored to the intended end use of the particles and can be modified by the environmental compartments in which they reside, which limits their detection and characterisation in soils and other complex environmental matrices through chromatographic techniques (Petersen et al., 2011). A summary of methods used to detect CNMs within soils and sediments is presented in Table 1.
Recent methods used to detect CNMs in complex environmental matrices.
In addition, as the lifecycles of CNM-containing products are likely to vary greatly, the means by which these materials enter the soil environment are also likely to be highly variable (Pan and Xing, 2012). An excellent review of different CNM exposure scenarios, for both humans and the environment, is provided by Köhler et al. (2008). Further complicating their detection is the emerging nature of manufactured CNMs as soil xenobiotic components, and hence their presently low concentrations, together with their interaction with naturally occurring nanomaterials and other environmental components, which leads to particles with sizes and compositions that significantly differ from their native forms (Nowack and Bucheli, 2007; Darlington et al., 2009; Chen et al., 2011). As such, careful consideration of multiple environmental variables is required to determine their impact on CNM fate and behaviour.
Once released into the soil, the fate and behaviour of CNMs is governed by their interactions with various components within the environment. Derjaguin–Landau–Verwey–Overbeek (DLVO) interactions, such as electrostatic interactions and van der Waals (vdW) forces, and non-DLVO interactions, such as hydrogen bonding and steric hindrance, ultimately determine the mobility, aggregation and adhesion of CNMs within soils. These forces may operate in concert to various extents, with the predominating force controlled by factors such as the properties and quantity of soil organic matter (SOM), characteristics of inorganic matter, and the type and quantity of clays, together with the properties of CNMs themselves. Each of these factors are heavily influenced by variables that are not necessarily constant over time, such as pH and ionic strength.
Soil organic matter plays a substantial role in both the fate and behaviour of CNMs through alterations in the dominance of the various DLVO and non-DLVO interactions. SOM (which consists primarily of decomposed plant and animal remains (Lee et al., 1981)) is an all-encompassing term describing organic matter (OM) dispersed ubiquitously throughout the soil environment, and is composed of a heterogeneous mixture of lipids, carbohydrates, carboxylic acids, humic substances, hydrophilic acids, proteins, carbohydrates, hydrocarbons and amino acids. However, the mechanism by which SOM maintains the CNM stability in suspension is still under investigation and debate (Dinesh et al., 2012). Specifically, the aim of this section is therefore to provide an overview of recent investigations in which the manner of solid SOM, dissolved organic matter (DOM), humic acid (HA) and tannic acid (TA) influence the behaviour of CNMs.
Adsorption of molecular DOM onto CNMs occurs through either aromatic ring
sorption or binding of aliphatic chains via
Typically, frequently occurring cations within the environment (K
When considering the fate and behaviour of CNMs in soils, solid peat may
have a different impact to that of molecular DOM under environmentally
relevant ionic conditions. This could be caused by the alteration of
particle-phase distributions due to the direct sorption of CNMs, as well as
the possibility of DOM or cations being release from the soil particles
themselves (Zhang et al., 2011a). In the absence of sodium ions, Zhang et
al. (2011a) found no adsorption of MWCNTs to solid peat, indicating a
limited affinity of DOM-MWCNT composites towards the solid phase relative to
the aqueous phase, possibly due to electrostatic repulsion and
hydrophilicity of DOM-coated nanotubes. With the addition of Na
Consideration of the
The properties of the humic substances determine the extent to which DLVO and non-DLVO interactions influence particle behaviour. TA (Chibowski et al., 1998) and HA sorbed to CNTs enhance stabilisation in water through reducing vdW forces between particles and increasing steric repulsion (Terashima and Nagao, 2007; Ren et al., 2011). However, Qu et al. (2012) identified that high molecular weight (HMW) HAs were more effective in promoting suspension stability due to stronger steric repulsion than that of low molecular weight (LMW) HAs. Similarly, HAs containing large quantities surfactive domains, such as those which are strongly hydrophilic and lipophilic, promote the dispersal of CNTs in solutions, while those containing carbohydrates and predominantly hydrophilic domains resulted in limited dispersal (Chappell et al., 2009).
The composition of SOM in relation to ionic strength and pH dictates the behaviour of CNMs within soils. Presently, however, insufficient data regarding the relative impact of different SOM fractions and combinations on DLVO and non-DLVO forces in soils is lacking, reducing the ability to estimate how CNMs may behave based on analysis of soil OM content. Furthermore, in addition to the organic fraction of soils and coating of CNMs, the role of the inorganic fraction in determining particle behaviour must also be considered.
In addition to the organic fraction, CNM stability in saturated soil–water
suspensions is strongly influenced by the impact of the inorganic fraction,
and is largely neglected within the present literature. Han et al. (2008)
studied the impact of kaolinite and montmorillonite clay minerals with
particle sizes of around 2
Furthermore, the charge characteristics of soils can also influence the behaviour and fate of CNMs. Broadly, all soils can be divided into two groups: permanent-charge (P-C) and variable-charge (V-C) soils (Sollins et al., 1988). In P-C soils, the substitution of ions with lower valence for ions with higher valence results in the alteration of crystal lattice structures within layer-silicate clays (illite, smectite, chlorite and kaolin) and a permanent charge deficit, which persists irrespective of variations in the composition of soil solutions and pH (Sollins et al., 1988). In V-C soils, protonation and deprotonation of surface hydroxyl groups results in the positive charge and hence anion exchange capacity (AEC), whereas deprotonation results in cation exchange capacity (CEC) (Sollins et al., 1988). The structure of V-C soils is also modified in response to increasing pH, resulting in increased repulsion and more limited aggregation (Sollins et al., 1988). Both P-C and V-C surfaces are present in all soil types; however, only one charge system typically dominates, dictated largely by soil mineralogy (Sollins et al., 1988). While V-C soils occur more frequently in tropical regions due to the typical mineralogical composition which forms under humid, warm conditions, they do not occur ubiquitously, and many areas with predominantly P-C characteristics occur (Sanchez, 1976; Sollins et al., 1988). Hence, while V-C soils represent a small fraction of global soil types, interactions between nanoparticles and soils are likely to be much more dynamic relative to those with a P-C. Despite this, almost all investigations have predominantly focused on P-C soils, restricting the applicability of CNM fate and behaviour investigations.
The behaviour of CNMs in V-C soils has been assessed by Zhang et al. (2012b), who investigated the stability of MWCNTs suspended in water containing either kaolinite, smectite or shale over a range of sodium concentrations. Without additional Na
In addition to the influence of soil type and properties, the properties of
CNMs themselves vary greatly depending on an array of parameters. As
commercial applications of CNMs will likely employ surface functional groups
and a variety of different preparation techniques, nanoparticle properties
and behaviour within the environment will become increasingly complex (Turco
et al., 2011). For example, the physicochemical properties of pristine
Typically, agglomeration of CNMs in the presence of divalent (Ca
In other investigations, surface immobilisation of macromolecules, such as
HAs at environmentally relevant concentrations, has increased the solubility
of C
Comparatively determining the relative importance of CNM functionalization
and ionic strength on CNTs and
Functionalization status is therefore a fundamental consideration to the behaviour of CNMs, resulting in distinct characteristics, which significantly modify behaviour in relation to their unfunctionalised counterparts. However, key questions as to the behaviour of CNMs within the environment remain unaddressed; for example, how does the repeated exposure of CNMs to weathering cycles within the soil influence their fate and behaviour?
The ability of natural colloids to assist in the transport of organic
contaminants has been well documented and reviewed (De Jonge et al., 2004;
Sen and Khilar, 2006; Li et al., 2013). Typically, hydrophobic compounds
such as PCBs and PAHs have limited environmental mobility due to strong
sorption to SOM. Kan and Tomson (1990), however, demonstrated that high
concentrations of colloidal materials such as DOM may enhance the transport
of hydrophobic compounds such as phenanthrene and naphthalene by a factor of
1000 or greater, with possible implications for the spread of
contamination and groundwater quality (De Jonge et al., 2004). Although CNMs
may be tailored to suit specific requirements, their behaviour is not
necessarily different to colloids naturally occurring in the environment
(Colvin, 2003; Lead and Wilkinson, 2006). To determine the relevance of
natural nanoparticle-facilitated transport of contaminants in porous media
such as soils, Kretzschmar et al. (1999) identified four key factors that
will be used as a framework for this section:
sufficiently high concentration of nanoparticles mobility of the nanoparticles carrying sorbed HOCs sorbate toxicity even when present in trace quantities the ratio of sorption to desorption relative to the timescale of particle
mobility.
The sorption affinity of CNMs for common environmental contaminants such as
PAHs, known to pose significant risks to both the environment and human
health due to their toxic properties (Menzie et al., 1992; Shaw and Connell,
1994; Cebulska-Wasilewska et al., 2007), has been reported as over 3
orders of magnitude greater than that of natural soil/sediments (Yang et
al., 2006b). The potential for these emerging materials to become widespread
in the soil environment, particularly those with a strongly hydrophobic
nature and large reactive surface area such as CNMs, raises questions and
concerns about the environmental consequences of their release (Pan and
Xing, 2010).
Understanding the adsorption and desorption of HOCs to CNMs in soils is critical to the environmental risk assessment processes, as well as determining their potential applications as environmental adsorbents (Yang et al., 2006a). As the fundamentals of CNM-HOC sorption have been extensively reviewed, the reader is referred to a review by Ren et al. (2011) for a comprehensive overview. This section addresses the manner in which soils may alter the HOC sorption/desorption properties of CNMs, focusing specifically on two conflicting effects: (i) CNM dispersal by DOM (increasing the surface area and hence the number of adsorption sites; Hyung et al., 2006; Lin and Xing, 2008), versus (ii) the formation of CNM-DOM coatings (blocking and/or competing for adsorption sites reducing the number available for organic contaminants (Chen et al., 2008; Wang and Keller, 2009; Cui et al., 2011; Wang et al., 2011; Zhang et al., 2011)). The relative importance of these two phenomena is poorly understood in relation to their sorption and desorption of organic contaminants (Zhang et al., 2011; Pan and Xing, 2012), and it is highly dependent on both the nanoparticle properties, as well as the nature of SOM and the sorbate (Wang et al., 2009; Zhang et al., 2011; Lerman et al., 2013).
In assessing the impact of OM on CNM sorption in the environment, further complications arise as contaminants are able to sorb to both the CNM and CNM-OM coating (X. Wang et al., 2008). Hyung and Kim (2008) identified that SOM adsorption to nanotubes was highly variable depending on the type of SOM, occurring proportionally to its aromatic carbon content. This has implications for determining the ability of CNMs to sorb organic compounds, yet most investigations fail to consider the role of different OM fractions in CNT-pollutant interactions (Lerman et al., 2013).
X. Wang et al. (2008) assessed the extent to which HAs and peptone altered the
sorption of phenanthrene, naphthalene or 1-naphthol onto MWCNTs (outer
diameter of 40 nm), by fitting sorption data with Freundlich and Polanyi
models. Their results showed that each type of DOM resulted in nonlinear
sorption isotherms to the MWCNTs, following the order peptone
Within a soil environment, Li (2012) identified the sorption behaviour of
naphthalene, phenanthrene and fluorine in a sandy loam soil, silt loam soil
and Ottawa sand was unaffected following amendment of MWCNTs at
concentrations of 2 mg g
An excellent study by Towell et al. (2011) assessed the extent to which
HPCD extraction of HOCs with different physicochemical properties desorbed
from soils amended with CNMs at concentrations between 0.05 and 0.5 %
(substantially larger than that employed by Li, 2012). At concentrations
While sorption of HOCs to CNMs in soils can occur, the extent of sorption and desorption is dependent on the type of OM and concentration of CNMs. With a view into the manner in which the properties outlined above potentially facilitate the transport of contaminants sorbed to CNMs in soils, studies in which mobility has been directly investigated will also be discussed.
“Worst case scenario” processes by which CNMs may facilitate the transport of HOCs. Top-left panel: (A) HOC equilibrates with CNM and is (B) transported. The top-right panel shows the processes by which CNMs may be transported. The centre-right panel (1) shows the transport and rapid desorption of HOCs from CNMs. Equilibrium is achieved between the liquid phase, CNM and matrix. The bottom-right panel (2) shows slow desorption kinetics, with no desorption from the CNM (Hofmann and von der Kammer, 2009). Re-printed with permission from Elsevier © 2014.
Once sorbed to freely suspended CNMs within the soil matrix, the mobility of
HOCs is potentially increased; however, very few studies have focused on
determining the impact of CNMs on contaminant movement in soils. An overview
of the basic principal of CNM-facilitated HOC transport is presented in
Fig. 1. Using column leach tests, Li (2012) examined the behaviour of
phenanthrene, fluorine, naphthalene and pyrene in a saturated sandy loam
soil amended with MWCNTs, functionalised MWCNTs (f-MWCNTs) and
functionalised SWCNTs (f-SWCNTs) at a concentration of 5 mg kg
To determine the extent to which CNMs facilitated the movement of
contaminants relative to various types of DOM, Zhang et al. (2011b) used
saturated, sandy soil columns contaminated with either PCBs or phenanthrene
to comparatively assess the mobilising ability of
Different processes of
Using a different approach, Hofmann and von der Kammer (2009) computer
modelled the extent to which CNMs could result in the movement of HOCs in
soils under various scenario-based conditions, to determine when relevant
CNM transport of sorbed HOCs might occur. Worst-case scenarios were adopted,
assuming fully mobile CNMs within the porous medium, over a range of
realistic yet high CNM concentrations (100 mg L
Parameterisation of the ratio of desorption to sorption and particle
transportation is achieved by the Damköhler number (
Simulation of diffusion-limited desorption using of pore
water velocities (va) between 1 m 50 d
Hofmann and von der Kammer (2009) calculated Damköhler numbers for CNM
aggregates of different sizes and partitioning coefficients according to the
rate constant data shown in Fig. 2, and based on different flow velocities
of 1 m in 50 d (fast flow)–1 m in 10 years (slow movement). It was inferred
that the CNM-contaminant transport mechanisms are strongly dependent on the
size of CNM agglomerates together with the distribution coefficients (log
From the above discussion, it can be concluded that each of the four factors identified by Kretzschmar et al. (1999) for significant transport of contaminants by CNMs have been met. However, more work examining the subsurface transport of CNMs through well-defined soils of various types (such as clays, peats and silts) and CNMs with a variety of functional groups, sizes and sorbed compounds in both saturated and unsaturated conditions are required (Jaisi and Elimelech, 2009; Petersen et al., 2011). Of studies that are available, variation in experimental conditions between the investigations renders comparisons of the efficiency of contaminant mobility between CNM types tentative until standardised comparative testing is conducted. Additionally, the molecular weights and sizes of CNMs may not be constant during their transport within the soil environment, due to their physical, chemical or biological interaction with soil components, which will likely influence their aggregation status, shape, surface charge (Pan and Xing, 2012), and possibly also their ability to sorb and mobilise contaminants over long timescales. Furthermore, definitive data of the desorption kinetics of HOCs from CNMs in soils are essential to understanding their ability to transport contaminants (Ibaraki and Sudicky, 1995; Choi and Yavuz Corapcioglu, 1997; Corapcioglu et al., 1999; Bold et al., 2003; Hofmann and von der Kammer, 2009), with slow desorption identified as a critical requirement (Roy and Dzombak, 1998). The lack of experimentally derived desorption kinetic data from a range of soil types and conditions makes determining the extent to which HOC sorption is strong enough and desorption slow enough to allow CNMs to transport sorbed HOCs, and the associated implications of transport, difficult to predict (Qu et al., 2012).
As soils represent one of the ultimate sinks for nanomaterials (Nowack and Bucheli, 2007), terrestrial microorganisms, which are a large component of soils, may be significantly affected (Navarro et al., 2008). The reader is directed to an excellent review by Holden et al. (2014), which evaluates the possible exposure concentrations of anthropogenic nanomaterials in a range of environmental compartments, and assesses their relevance. However, understanding the impact of CNMs on the soil microbial community is a subject still in its infancy (Dinesh et al., 2012). The extent to which CNMs interact with microflora will (in part) determine the extent of possible disruptions to bio-geochemical processes within soils that they may cause (Neal, 2008). This section discusses recent literature related to the modification of CNM fate and behaviour by microbiota, the toxicity of CNMs in soils and the possible implications for the biodegradation of contaminants.
The influence of microbial populations on the physical and chemical state of
nanoparticles must be considered when discussing the ultimate fate of
nanomaterials (Aruguete and Hochella, 2010). Degradation of C
Ultimately, Turco et al. (2011) suggested that the fate of C
The toxicity of CNMs is dependent upon the bioaccessibility of nanoparticles to bacteria, and retention of some the nanoparticles' reactivity (Neal, 2008). Currently, little literature is available related to the toxicity of CNMs within soils (Dinesh et al., 2012). Hence, the discussion presented here provides a theoretical estimation of the specific microbial communities that may be more vulnerable to soilborne CNMs, followed by an overview of recent CNM-amended soil toxicity findings published within the literature.
Soil conditions will ultimately dictate the extent to which CNMs are able to interact with terrestrial microflora. Based on the discussion earlier relating to the fate and behaviour of CNMs in soils, in addition to information regarding cell properties (Mehmannavaz et al., 2001), it may be possible to tentatively speculate as to the bioavailability or bioaccessibility of CNMs to different microbial populations. When assessing nanotoxicity, consideration must be given to both the likelihood of a nanoparticle coming into contact with microbial cells and the initial concentration added to soils in order to provide an accurate means of estimating the particle availability (Dinesh et al., 2012). A strong interplay exists between the dispersal status of nanoparticles and their bioaccessibility to specific soil microbial populations (Turco et al., 2011). As bacteria frequently adhere to surfaces in the soil environment, attached cells within biofilms constitute a large proportion of the bacterial community in the subsurface environment (Neal, 2008). Neal (2008) therefore proposed that the study of nanotoxicity towards biofilm communities is a more appropriate measure of toxicity in environmental systems than planktonic cells. However, it is conceivable that given appropriate DLVO and non-DLVO forces between CNMs, microorganisms and the soil matrix, CNMs could also become available to planktonic cells. One example of which may be that CNM-SOM coatings could result in easier access to the cell surface relative to uncoated particles due to the similarities in solubility between the cell membrane and surfactant; however, the coating itself may attenuate the toxicity due to a lack of physical contact between the CNM and a microbial cell (Lubick, 2008). Further work into the conditions under which CNMs will be available to different microbial communities in soils is needed.
The extent to which soils with different properties determine the toxicity
of some CNMs was directly investigated by Chung et al. (2011). The impact of
MWCNTs at 50, 500 and 5000
Other investigations of nanotoxicity within soil using
In a similar investigation, Tong et al. (2007) assessed the role aggregation
status plays in determining nanotoxicity within soils. The impact of either
Despite differences in experimental setups between the studies by Johansen
et al. (2008) and Tong et al. (2007), from the data presented, it is not
possible to rule out the bioaccessibility and toxicity of C
Sorption of contaminants is a fundamental mechanism in the regulation of organic compound bioavailability (Lou et al., 2011). Given their strong sorptive capability, the addition of CNMs to soil may result in the sequestration of organic contaminants, reducing their extractability and bioaccessibility, operating in a similar manner to hard or black carbon (Chen et al., 2007). However, the extent to which the processes identified in Sect. 4.1 impact upon the bioaccessibility of contaminants and biodegradation has not received much research within soils.
The conditions under which CNMs enter the soil are also critical to
determining their impact upon contaminant bioaccessibility. Zhou et al. (2013) incubated
These results show that CNT interactions with contaminants within the soil
environment reduced the number of available sorption sites, with their
sorptive ability further reduced by CNM aggregation and interaction with
soil components such as humic substances, DOM, peptone and TA, which
potentially coat CNTs modifying surface polarity, reducing surface area and
hence reducing HOC sorption capacity as discussed in Sect. 4.1 (X. Wang et
al., 2008; Cui et al., 2011; Zhou et al., 2013). As the adsorption of
Similar results were obtained by Cui et al. (2011). Sediments (20 g) were
first amended with either biochar (100 mg), charcoal (20 mg) or SWCNTs (20 mg), then spiked with phenanthrene (0.50 mg kg
In addition to the impact of soil types on the impact of CNMs on organic
contaminant sorption, properties of the organic chemicals within soils are
also influential in dictating their interaction with different types of
CNMs. Towell et al. (2011) assessed the impact of fullerene soot (FS),
SWCNTs and MWCNTs at 0, 0.05, 0.1 and 0.5 % concentrations, on the HPCD
extractability (proven as an indicator of PAH bioaccessibility to soil
microflora (Reid et al., 2000; Doick et al., 2005; Stokes et al., 2005; Rhodes et
al., 2008b) and mineralisation of
When considering the fraction of contaminants sorbed to CNMs within these investigations, and the resulting reduced bioavailability, two schools of thought may be adopted: (i) over time the nondegradable, bound fraction may innocuously degrade (Gevao et al., 2000a), or (ii) the bound fraction is potentially remobilised over long timescales with potential environmental implications (Gevao et al., 2000b). This draws on the discussion by Semple et al. (2013), in which the significance of distinguishing between bioavailability and bioaccessibility is significant, particularly when dealing with environmental “super sorbents” such as CNMs with reference to remediation of contaminated land and risk assessment. Semple et al. (2004) defined bioavailability as “that which is freely available to cross an organism's cellular membrane from the medium the organism inhabits at a given time”, and is considered as a rate of substrate delivery to cells. While bioaccessibility encompasses this fraction, it additionally extends to those which are potentially available over time, but are currently chemically or physically removed from the microorganism (Semple et al., 2004). In other words, it provides a definition of the total extent of substrate that will be available to cells. Arguably, bioaccessibility is of relatively greater importance when considering the fate and behaviour of CNM-sorbed contaminants, due to the larger temporal range and lack of implied immediacy. However, under some environmental conditions, microbial colonisation of CNM agglomerates can occur, with potential implications for the bioaccessibility of the bound contaminant fraction.
While the toxicity of CNMs in soil is dependent on their bioaccessibility in addition to retention of reactivity, if agglomerates of CNMs are present with a reduced cytotoxic nature, it is conceivable that interstitial gaps in the agglomerate with mesopore dimensions will result in their increased suitability for the sorption of microorganisms (Agnihotri et al., 2005; Upadhyayula and Gadhamshetty, 2010). When this is related to the previous discussion of CNM contaminant sorption and the implications for biodegradation, it is possible to reconsider the lack of bioaccessibility of CNM-sorbed contaminants reported in some studies, and consider their potential to increase contaminant bioaccessibility in certain situations. Properties of particular importance when considering CNMs for such applications include (i) structures with high porosities readily colonisable by microorganisms, (ii) potential ability to encourage biofilm formation through offering a buffering capacity, and (iii) the ability to adsorb high concentrations of contaminants from bulk solution yet regulate the microbial biodegradation through desorption (Abu-Salah et al., 1996).
Biofilms are groups of well-organised, adjoining cells encapsulated within a
matrix of insoluble, extracellular polymetric substances (EPS) (Morikawa,
2006). EPS encapsulation supports cell substance and growth through the
trapping, binding and dissemination of external nutrients by charged
polysaccharide groups (Cheng et al., 2007), and offers greater protection
against external stresses within the environment relative to those residing
in a planktonic state (Pang et al., 2005). Materials that allow a high
degree of bacterial colonisation and possibly biofilm formation are
potentially suited to facilitating biodegradation (Upadhyayula and
Gadhamshetty, 2010), which is typically most effective when microorganisms
are in a biofilm state as opposed to planktonic, due to greater bioavailability,
protection and adaptability to toxic conditions and hence more rapid
pollutant degradation (Singh and Cameotra, 2004; Singh et al., 2006).
Furthermore, bacterial colonisation may stabilise nanoparticle aggregates
as polysaccharides, such as those generated by bacteria, have been observed
to significantly increase the aggregation of C
Upadhyayula and Gadhamshetty (2010) conducted hypothetical calculations to
determine the quantity of cells that an agglomerate of CNTs could
potentially sorb. The dimensions of a typical bacterium such as
When the potential for biofilm development on CNMs is considered in relation
to their HOC sorptive ability and aggregation within soils, it has been
suggested that CNMs may be useful for enhancing biodegradation of organic
pollutants that cannot be easily concentrated. With CNM aggregates behaving
as an organic chemical collector and accumulator, biofilm development on
CNMs potentially increases the bioavailability/bioaccessibility of the
contaminant (Yang et al., 2006b). Given adequate reversibility of organic
compound adsorption and limited desorption hysteresis, sorption of bacterial
cells to the surface of CNM aggregates may shorten the diffusion distance,
facilitating the utilisation of the sorbed organic compound by the bacteria.
This is well illustrated by Yan et al. (2004), who studied the removal
efficiency of microcystin (MC) toxins from solution by
Xia et al. (2013) studied the bioavailability and desorption (Tenax TA) of
Very little information is available on how CNMs act within soil matrices,
especially in relation to their adsorption to organic fractions, organic
pollutants and their subsequent toxicity (Dinesh et al., 2012). With an
angelus sorbents such as black carbon (BC), elevated mineralisation of a
phenanthrene substrate has been observed as a direct result of BC addition
to soil, which was tentatively attributed to microbial sorption and
utilisation of phenanthrene from the sorbed phase (Rhodes et al., 2008a, 2012). Only one study has identified an increase in
contaminant mineralisation in soils following the addition of CNMs. Xia et
al. (2010) studied phenanthrene biodegradation and desorption
characteristics (using XAD-2) in 21–40 d aged MWCNT-amended soils relative
to soils amended with wood char and black carbon. Following each ageing
interval,
Given the discussion above, it is possible to consider an additional factor to those proposed by Kretzschmar et al. (1999) in Sect. 4, to determine the significance of contaminant-facilitated transport by CNMs. If the CNM-sorbed contaminant is available to the cells through utilisation from the sorbed phase, the importance of desorption of sorbed compounds from CNMs during transport is reduced. It is therefore proposed that incorporation of a fifth factor, “the bioavailability and bioaccessibility of CNM sorbed contaminants to microorganisms from the solid phase” may be appropriate, inferring that bioaccessibility through desorption investigations may lead to incorrect assumptions. However, substantially more work is required to identify the exact mechanism involved in these findings, and the specific conditions under which contaminant and microbial sorption to CNMs could potentially result in toxicity from the CNM itself, from the sorbed contaminant or both (Nowack and Bucheli, 2007). It is also possible that under some environmental conditions, rapid desorption or excessive bioavailability of sorbed contaminants may shock load sorbed bacteria and prove toxic (Upadhyayula and Gadhamshetty, 2010). Biodegradation of contaminants sorbed to CNMs therefore still requires substantial investigation into specific combinations of pollutants and microorganisms (Upadhyayula and Gadhamshetty, 2010), to determine whether the bioaccessibility of sorbed contaminants is either increased or decreased, and if the addition of CNMs will increase the mobility of contaminants in the environment. The general paucity of knowledge regarding the duration for which contaminants will remain sorbed to CNMs requires addressing to determine the long-term stability of contaminants sorbed to different nanoparticle types. Furthermore, the extent to which CNMs influence the transformation residues of HOCs in soils such as bound residues formed during organic pollution degradation in soil is unknown (Barriuso et al., 2008; Shan et al., 2011; Zhou et al., 2013).
The complex and dynamic nature of both soil environments and CNM physicochemical properties generates enormous uncertainty in attempting to predict their behaviour and impact on contaminant sorption, sequestration and transport as well as microbial interactions. This review argues that the fate and behaviour of CNMs in soils is influenced by multiple parameters such as the type and quantity of SOM, the type of clay particles present, the dominant charge characteristics of the matrix as dictated by the soil inorganic fraction, and the properties of the CNM, each of which is heavily influenced by pH and ionic strength. In addition, to a small extent, biological activity has been shown to modify the carbon nanomaterial fate. However, presently no research has been conducted into the manner in which these factors interact and collaboratively influence the fate and behaviour of CNMs in real environmental scenarios; therefore additional research is required.
The extent to which CNMs are able to modify the behaviour of contaminants in soils and facilitate their transport is dependent on the CNM concentration, the properties of SOM, molecular weight of the HOC and the interaction of the CNM with the HOC before its addition to soils. When present in sufficient concentrations, CNMs have the ability to facilitate the transport of co-existing contaminants such as PAHs to a greater extent than naturally occurring colloids such as DOM, the extent of which is dependent on the physicochemical properties of the contaminant, CNM functionalization status, aggregation size and method of preparation. Further work derived from experimental research is needed to address the lack of data relating to the transport of CNMs through soils of different properties. Additionally, CNM-HOC desorption kinetics within soils require defining, as this presently limits our understanding of the significance of CNM-facilitated transport.
Finally, CNMs are undoubtedly efficient sorbents for a range of HOCs. However, while a reduction in the bioaccessibility of contaminants in soils following the addition of CNMs has been demonstrated (Towell et al., 2011), further research is required before their sorbtive abilities can be applied to the remediation of contaminated soils. Specifically, information regarding the stability of sorbed contaminants, as well as their potential to increase contaminant mobilisation together with other secondary effects, are as yet too poorly understood to fully anticipate the possible environmental impact of CNMs. To determine the behaviour of CNMs within soils, it is concluded that no one set of environmental or CNM characteristics can be viewed in isolation. Hence, given the diverse array of variables, it is argued that risk assessment of CNMs within the soil environment should be conducted on a case-by-case basis. A detailed analysis of other environmental compartments in which CNMs can potentially accumulate, such as sediments, should also be considered.
Edited by: R. Zornoza