Chemical transformation of silver nanoparticles in aquatic environments: Mechanism, morphology and toxicity

Chemosphere 191 (2018) 324e334

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Review

Chemical transformation of silver nanoparticles in aquatic environments: Mechanism, morphology and toxicity Weicheng Zhang a, b, Bangding Xiao b, Tao Fang b, * a

Key Laboratory of Ecological Security for Water Source Region of Mid-line of South-to-North Diversion Project of Henan Province, College of Agricultural Engineering, Nanyang Normal University, Nanyang, 473061, China b Institute of Hydrobiology, Chinese Academy of Sciences, Wuhan, 430072, China

h i g h l i g h t s

g r a p h i c a l a b s t r a c t


Importance of morphology and toxicity changes of Ag NPs in whole water environment is highlighted.
Correlations between water environmental factors and these four chemical transformations of Ag NPs are discussed.
Various environmental transformations of Ag NPs should be accounted for in a full picture to understand their human health and environmental risk assessment.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 15 July 2017 Received in revised form 16 September 2017 Accepted 2 October 2017 Available online 3 October 2017

Silver nanoparticles (Ag NPs) have been inevitably introduced into ecological environment during their extensive applications in daily human life. Thermodynamically, Ag NPs are unstable and transform into other species under various aqueous conditions. Ag NPs and their transformation products pose potential threats to environment and humans. However, the complex environmental conditions and transformations of Ag NPs complicate their human health and environmental risk assessment. To bridge the knowledge gap, four essential environmental transformations, oxidative dissolution, sulfidation, chlorination and photoreduction, of Ag NPs are reviewed herein. The mechanism, morphology and size change, as well as the toxicity of Ag NPs during these transformations under certain aqueous conditions are detailed. In particular, these environmental transformations have shown strong correlations that are discussed. The transformation, fate, bioavailability, morphology and toxicity of Ag NPs are critical factors and should be considered in a complete human health and environmental risk assessment of Ag NPs. The fluctuation of these factors in the realistic environment is also vital and should be considered. © 2017 Elsevier Ltd. All rights reserved.

Handling Editor: Tamara S. Galloway Keywords: Silver nanoparticles Chemical transformation mechanisms Morphology Toxicity

Contents 1. 2.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 325 Chemical environmental transformations of Ag NPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 326

* Corresponding author. No. 7 Donghu South Road, Wuchang District, Wuhan, Hubei Province, China. E-mail address: [email protected] (T. Fang). https://doi.org/10.1016/j.chemosphere.2017.10.016 0045-6535/© 2017 Elsevier Ltd. All rights reserved.

W. Zhang et al. / Chemosphere 191 (2018) 324e334

325

2.1.

3. 4.

Oxidative dissolution of Ag NPs and the toxic impact . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 326 2.1.1. Mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 326 2.1.2. Shape and size change . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 327 2.1.3. Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 327 2.2. Sulfidation of Ag NPs and the toxic impact . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 328 2.2.1. Mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 328 2.2.2. Shape and size change . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 329 2.2.3. Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 329 2.3. Chlorination of Ag NPs and their toxic impact . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 330 2.3.1. Mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 330 2.3.2. Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 330 2.4. Reduction of Agþ to Ag NPs and toxic impact . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 330 2.4.1. Mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 330 2.4.2. Shape and size change . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 332 2.4.3. Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 332 Correlation of environmental transformations of Ag NPs and their common influencing factors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 332 Environmental significances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 333 Notes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 333 Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 333 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 333

1. Introduction Silver nanoparticles (Ag NPs) are known for their antibacterial capability and are one of the most widely produced and commercially applied nanoparticles (Tolaymat et al., 2010). Ag NPs have garnered public attention in environmental safety and toxicity domains, because they have been introduced into the aquatic environment during production, storage, and application, and most of Ag NPs (>50%) could be released from productions (such as clothes) during their first wash (Benn and Westerhoff, 2008; Geranio et al., 2009; Kittler et al., 2010). Hoque and him colleagues collected Ag NPs existing information from a Canada wastewater treatment plants (WWTPs). The amounts of Ag NPs (9.3 nm) could up to 1.90 mg/L in wastewater (Hoque et al., 2012). The potential threats of Ag NPs have been shown in a range of organisms, such as bacteria, algae, fungi, invertebrates, plants and fish (Marambio-Jones and Hoek, 2010; Fabrega et al., 2011; Sharifi et al., 2012; Bondarenko et al., 2013; Le Ouay and Stellacci, 2015; Soenen et al., 2015). Many excellent studies had previously reviewed the toxicity (Johnston et al., 2010; Marambio-Jones and Hoek, 2010; Fabrega et al., 2011; Levard et al., 2012; Sharifi et al., 2012; Bondarenko et al., 2013; Le Ouay and Stellacci, 2015; Soenen et al., 2015), physical, biological and chemical transformations (Marambio-Jones and Hoek, 2010; Fabrega et al., 2011; Levard et al., 2012; Lowry et al., 2012b; Virender, 2013; Yu et al., 2013; Dwivedi et al., 2015; Zhang et al., 2016a), and biouptake (Rai et al., 2009; Johnston et al., 2010; Mahmoudi et al., 2011; Sharifi et al., 2012) of Ag NPs. However, a knowledge gap remains in the health and environmental risk assessment of Ag NPs. Predicting or assessing environmental risks of Ag NPs remain difficultly because they present distinct toxicity to particular organisms (Garner et al., 2015) and Ag NPs trend to transfer to diverse species in different aquatic environments. Importantly, Ag NPs and their chemical transformation products normally coexist in aquatic environment and jointly influence environment and human health. Thus, the risk and toxicity of Ag NPs to aqueous environments could not solely explained by their nano properties, such as size and surface activity, but greatly depend on their chemical transformation products. In addition to their fate, toxicity and bioavailability, the environmental transformation processes and products of Ag NPs are primarily and critically important for determination of the human

health and environmental risks posed by Ag NPs. Ag NPs have various transformation pathways, as shown in Scheme 1, and various transformation products under certain aqueous conditions. Generally, Ag NPs are unstable under oxygenrich aqueous conditions, part of the Ag NPs tend to dissolve to Agþ. Therefore, Ag NPs and Agþ usually coexist under oxygen-rich aqueous conditions, and they cooperate to produce hazardous effects to aquatic organisms. Interestingly, many investigations had found toxicities of Ag NPs are primarily contributable to released Agþ that only account for less than 10% of total Ag NPs mass (Fabrega et al., 2009; Peretyazhko et al., 2014). The Agþ can be photoreduced to small Ag NPs by sunlight with dissolved organic matters (DOMs). This process will alter the equilibrium between Ag NPs and Agþ, and impact fate and toxicity of Ag NPs. Notably, sunlight shows two different effects on release of Agþ from Ag NPs (Yu et al., 2016): it increases Agþ release by accelerating the oxidation of Ag NPs(Grillet et al., 2013) and decreases the Agþ release by promoting the aggregation of Ag NPs (Cheng et al., 2011; Shi et al., 2012, 2013). Both Ag NPs and Agþ have a high affinity toward sulfur that naturally exists in anaerobic environments to form Ag2S solid. Due to low solubility, the newly formed Ag2S is more stable and less toxic than Ag NPs or Agþ; thus, the sulfidation processes of Ag NPs are considered a natural antidote for the

Scheme 1.

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toxicity of Ag NPs (Levard et al., 2013a). In seawater or in Cl-rich water, Ag NPs are initially oxidatively dissolved, and then, they form AgCl NPs and AgCl(x1) compound. Similar to sulfidation, the x chlorination of Ag NPs to AgCl decreases the toxic impact to organisms. However, while the sulfidation and chlorination can mitigate the toxicity of Ag NPs in most cases, the transformation products, Ag2S and AgCl or AgCl(x1) , are still bioavailable and x toxic to some organisms (e.g., plants and Caenorhabditis elegans (Doolette et al., 2015; Starnes et al., 2015; Wang et al., 2015, 2016; Collin et al., 2016)). Moreover, an increased hazardous toxicity of Ag NPs can also be induced during sulfidation and/or chlorination processes, because of free Agþ released as an intermediate product has potential toxic effects. The fate and transformation of Ag NPs present various patterns in real water environments depending on certain conditions (Levard et al., 2012; Lowry et al., 2012a, 2012b). It has been shown that oxidative transformation of Ag NPs to release Agþ dominates the reaction in oxygen-rich water solutions (Yu et al., 2016), which is followed by simultaneous dissolution of Ag NPs and reduction of Agþ until reaching equilibrium (Peretyazhko et al., 2014). The reductive transformation of Agþ to form Ag NPs is a predominant process under sunlight irradiation in the presence of DOMs (Yu et al., 2016). The transformation of Ag NPs shifts from dissolution to sulfidation with a switch from a lack of sulfide source to a rich sulfide source water environment. AgCl formation via chlorination replaces the dissolution of Ag NPs when Ag NPs shift from fresh water to seawater. Although Ag2S is more stable than AgCl and is the dominant Ag species under anaerobic conditions, AgCl or other species (AgCl(x1) ) remain (Lombi et al., 2013; Sekine et al., 2015; x Wang et al., 2016). Importantly, these environmental transformations of Ag NPs have strong correlations and share many environmental influencing factors. For example, oxidative dissolution of Ag NPs is the primary step for others processes, except for direct sulfidation mechanism. Sunlight irradiation and pH influence both dissolution and photoreduction processes of Ag NPs. DOMs has been shown to suppress oxidation of Ag NPs but promotes reduction of Agþ (Yu et al., 2016). In addition, the shape and size, which can affect the toxicity of Ag NPs, can be altered during these environmental transformations. Chemical transformation products of Ag NPs showed different morphologies or sizes and various toxicological responses with initial Ag NPs. These environmental transformations and their correlations should be considered in a full risk assessment of Ag NPs. To purpose of understanding co-effect of Ag NPs and their chemical environmental transformation products in water environment, this review only takes into account chemical transformations of Ag NPs. Physical transformations, such as homo- and hetero-aggregation, are also very important for Ag NPs’ environmental risk, but they will not be detailed in this review. This review details the mechanisms of four major chemical environmental transformations (oxidative dissolution, photoreduction, sulfidation and chlorination) and corresponding shape (and size) and toxicity change of Ag NPs under certain aqueous conditions. Finally, the correlation of these transformations and their common environmental influencing factors are discussed.

protons are required in Ag NPs dissolution (equation (1)). Thermodynamic studies indicated that Ag NPs tend to dissolve completely in oxic water environments (Liu and Hurt, 2010); however, rate information about the dissolution was not given. According to the data and as reviewed previously (Liu et al., 2010; Zhang et al., 2016a), the dissolution of Ag NPs falls into two stages (Scheme 2). The first stage is a fast dissolution stage, which is normally described by first-order kinetics (see equation (2), where t is the reaction time (h), k is the pseudo-first-order reaction rate constant (1/h), and CAgNPs(t) is the concentration of Ag NPs at time t (mg/L)). The second stage is a slow dissolution stage, which is characterised by “a zero reaction rate” (almost no apparent silver ion release) (Zhang et al., 2016a). The fast dissolution stage can be explained by the fast dissolution of a Ag2O layer and the desorption of Agþ adhered to the surface of Ag NPs. Therefore, the fast dissolution stage is oxygen independent but is still related to protons, as shown is Scheme 2. Once the outer layer of Ag2O dissolves completely, further oxidation of metallic Ag to Ag2O might occur, and the dissolution of Ag NPs continues. Comparably, the explanations for the slow dissolution stage can be considered as follows. (I) The oxidation of Ag NPs surfaces is a slow and rate-determining step for dissolution (Liu and Hurt, 2010). In oxic water, Ag NPs could be oxidised to generate Ag2O that adheres to the surface of Ag NPs to form a core (Ag NPs)-shell (Ag2O) structure. The oxidative process of Ag NPs is slow, although the formed Ag2O dissolves quickly. Additionally, the Ag2O shell can still affect the Ag NPs dissolution by protecting it from further oxidation or by reducing its available surface for dissolution. (II) Particle size and concentration of Ag NPs limit the release kinetics. Ma et al. (2014) found dissolution rate of Ag NPs ranged from 1% for the large particles (80 nm) to up to 60% for the smallest particles (5 nm) during over three months. Moreover, our previous research detected that high concentration of Ag NPs (10 mg/L) presented lower dissolution rate of Agþ (<10%) compared to low concentration of Ag NPs (1 mg/L; > 20%) in 48 h experimental duration (Zhang et al., 2016b). Typically, Agþ release is highly dependent on surface area of Ag NPs. Thus, larger size or high concentration (trend to homoaggregate) Ag NPs lose fewer Agþ ions than the smaller or low concentration Ag NPs (Dobias and Bernier-Latmani, 2013), due to lack of surface area. (III) The released Agþ could be resorbed by the Ag NPs and inhibit the oxidation or block the dissolution sites (Liu et al., 2010). Indeed, the released Agþ can adhere to Ag NPs surfaces and limit the diffusivity of O2 and protons (Peretyazhko et al., 2014), and newly formed large Ag NPs,

2. Chemical environmental transformations of Ag NPs 2.1. Oxidative dissolution of Ag NPs and the toxic impact 2.1.1. Mechanisms Oxidation is a primary process for Ag NPs transformation, such as to form AgCl, Ag2S, Ag NPs and silver complexes during aquatic environment movement (except a direct sulfidation pathway). Chemically, both oxidation agent (dissolved oxygen, DO) and

Scheme 2.

W. Zhang et al. / Chemosphere 191 (2018) 324e334

with less available surface, slow the dissolution rate compared to that of the initial small Ag NPs. 4Ag þ O2 þ 4Hþ / 4Ag þ 2H2O



(1)

dCAgNPsðtÞ ¼ k CAgNPsðtÞ dt

(2)

Ag þ O2 / Agþ þ intermediate superoxide (O2) (rate limiting step), (3)


In aqueous environments, the most important factors affecting Ag NPs dissolution are oxidation agent and protons. Indeed, no Agþ release found is from Ag NPs under anoxic aqueous conditions (Xiu et al., 2012). Chemically, the dissolved oxygen serving as an oxidiser can be replaced by hydrogen peroxide (H2O2), which can promote the dissolution rate (Liu and Hurt, 2010). Under acidic conditions, the Ag NPs dissolved quickly and sometimes completely. However, the dissolution of Ag NPs is hindered under alkaline conditions, suggesting the critical role of protons. In addition, particle size also governs the dissolution of Ag NPs, as demonstrated by small Ag NPs dissolving quickly and more completely than large Ag NPs. As discussed above, the dissolution is surface-area modified. Thus, any factors that affect the surface area of Ag NPs can moderate the extent or rate of dissolution. Other environmental factors, such as sunlight, temperature, HS, Cl, ionic strength and NOMs, can also greatly influence the dissolution extent and rate (Liu and Hurt, 2010; Liu et al., 2010; Levard et al., 2012; Yu et al., 2016), but these factors are not detailed herein. 2.1.2. Shape and size change Due to the simultaneous oxidative dissolution and re-formation of Ag NPs, two typical shape changes are detected during the dissolution of Ag NPs. The first change is the formation of a core (Ag NPs)-shell (Ag2O) structure during the oxidative process (see Scheme 2). In this process, the initial Ag NPs surface can be oxidised to Ag2O by an oxidiser in water, such as O2 and H2O2. The second change is the re-formation of new large size Ag NPs. Ag NPs (6, 9 and 13 nm) showed increasing sizes after dissolution, likely through the Ostwald ripening mechanism, but with no morphology change (spherical shape) (Peretyazhko et al., 2014). In contrast to the photoreduction process (discussed below), small and newly formed Ag NPs were observed during the dissolution. Additionally, larger particles and worm-like structures of Ag NPs were also observed, which can be mechanistically explained by the simultaneous dissolution and reduction of Ag NPs (Yu et al., 2016). 2.1.3. Toxicity The toxicity of Ag NPs could be derived from ion-related and/or particle-related toxicity mechanisms, and both mechanisms can be adjusted by the dissolution of Ag NPs. Considering the ion-related toxicity mechanism, oxidative dissolution could greatly increase the toxicological response of Ag NPs in organisms. For example, small Ag NPs present a higher extent and faster rate of dissolution than large Ag NPs (Dobias and Bernier-Latmani, 2013), resulting in a large amount of free Agþ in solution that induces more serious damage to target organisms (Liu and Hurt, 2010; Zhang et al., 2016a). Importantly, bacterial membranes normally present negative charges that show a great affinity for cationic silver (I) but not for Ag NPs. Therefore, Agþ can be easily adsorbed on bacterial membranes and cause a more serious toxicological response. In addition, large Ag NPs can persist in the environment; thus, a Trojan-horse mechanism converts Ag NPs into free Agþ pools, to continuously provide released Agþ, that can release free Agþ

327

continuously (Hsiao et al., 2015). Typically, the toxicity mechanism of Agþ involves binding with proteins/peptides and/or DNA to inhibit intracellular signal transduction. Importantly, oxidative stress could also be possible depending on the detected superoxide intermediates during the oxidative dissolution processes (see equation (3)). For example, Liu and Hurt found that H2O2 was generated slowly during oxidation if Ag NPs were in an airsaturated aqueous solution (Liu and Hurt, 2010). Di He et al., sug
gested that O2 is more likely formed in the oxidative process of Ag NPs with oxygen (He et al., 2011). When considering the particlerelated toxicity mechanism, small Ag NPs cause severe damage because of their large surface area to volume ratio and high chemical activity. Within a series of differently sized citrate-coated Ag NPs, smaller Ag NPs resulted in higher Ag ion dissolution and toxicity (Ivask et al., 2014). However, larger Ag NPs, with a lower reactivity capability on the surface, can persist for prolonged periods in the environment compared to that of small Ag NPs, and thus they can act as mobile Agþ pool inducing potential ecological hazards. The newly formed large Ag NPs remain for a prolonged period in the environment compared with the initial small Ag NPs. Indeed, several investigations have found that Ag NPs (≈ 50 nm) persist in aqueous environments for many months (Dobias and Bernier-Latmani, 2013; Gorka and Liu, 2016). Considering that the dissolutionere-formation of Ag NPs occurred concurrently, the concentrations of released free Agþ and residual Ag NPs are dynamic during the dissolution processes. Correspondingly, the ionrelated and particle-related toxicological responses of Ag NPs are also dynamic until them reach to dissolution equilibrium. Theoretically, particle-related toxicities are dominant at the initial Ag NPs exposure because of the lack of free Agþ in solution. The ionrelated toxicological response could begin to dominate when the exposure is continual because of the increasing Agþ concentration and the decreasing Ag NPs concentration (see Fig. 1). For example, our latest study indicated that the toxicity of Ag NPs to Escherichia coli is primarily attributable to particle-related toxicity (59%) at the 6 h time point but to ion-related toxicity (57%) at the 48 h time point (Zhang et al., 2016b). Moreover, the toxicity of Ag NPs to Saccharomyces cerevisiae had a different pattern(Zhang et al., 2016b). Further investigations are required to fully understand the impact of dissolution on the toxicity of Ag NPs. In summary, the process of Ag NPs dissolution involves a fast

Fig. 1. The relative percentages of released Agþ and retained Ag NPs in double-distilled water at specific times and the corresponding relative ion-related toxicity and particlerelated toxicity of Ag NPs to Escherichia coli.

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dissolution mechanism upon initial exposure and a slow dissolution mechanism that appears several hours later. Among the many environmental influencing factors, the DO and pH are most important for Ag NPs oxidative dissolution. Oxidative dissolution strongly mediates the toxicity of Ag NPs by dynamically modifying ion-related and particle-related toxicities. As an essential and primary transformation process, the oxidative dissolution of Ag NPs is vital for assessment of the environmental risk of Ag NPs.

2.2. Sulfidation of Ag NPs and the toxic impact 2.2.1. Mechanisms Two sulfidation mechanisms of Ag NPs were proposed, direct and indirect (Scheme 3), which are both dependent on sulfide concentrations (Liu et al., 2011). The first mechanism, particles (Ag NPs) forming particles (Ag2S NPs), typically occurs at the high sulfide concentrations existing in anaerobic environments. Theoretically, high concentration of sulfide with dissolved oxygen cannot coexist, implying that dissolved oxygen should be not a favourable electron acceptor. However, dissolved oxygen is still needed in the direct sulfidation mechanism (Liu et al., 2011), although without an oxidative dissolution process, as shown in equations (4) and (5) (Liu et al., 2011). Indeed, the very good stability of Ag2S NPs with DO in the light or dark could partially support this fact (Li et al., 2016a).

4Ag þ O2 þ 2H2 S/2Ag2 S þ 2H2 O

(4)

4Ag þ O2 þ 2HS /2Ag2 S þ 2OH

(5)

In contrast, the indirect sulfidation mechanism occurs at low sulfide concentrations, with Agþ initially dissolving from Ag NPs and then precipitating with sulfide (HS, H2S). Given that oxidative dissolution is the first process and precipitating process is fairly rapid, DO is essential in indirect sulfidation mechanism, and the Agþ released should be a rate-determining step (equation 6a and b).

4Ag þ O2 þ 2H2 O/4Ag þ 4HO 2Ag

þ

þ

ðslow and rate determining stepÞ

þ HS þ HO /Ag2 S þ H2 O

Scheme 3.

ðfast stepÞ

(6a) (6b)

Different outcomes of the sulfidation of Ag NPs have been observed in natural environments and in the laboratory. Sulfides exist in the natural environment at lower concentrations (pM to mM range) (Kaegi et al., 2013) than those used in laboratory experiments (1e10 mM) (Reinsch et al., 2012; Baalousha et al., 2015). Therefore, dominance of the indirect sulfidation mechanism is possible. However, sulfidation via the direct mechanism was observed in a wastewater treatment plant (WWTP) (Kent et al., 2014) even under a low S/Ag ratio (Baalousha et al., 2015), without the involvement of dissolution and re-formation processes. With different sources of sulfide, the extent of sulfidation of Ag NPs occurred much more slowly in mesocosms (H2S or HS) than in the laboratory (Na2S), as reflected in the transformation of Ag NPs to Ag2S remaining incomplete 18 months later (Lowry et al., 2012a). Correspondingly, the sulfidation rates of Ag NPs in the laboratory have ranged from minutes to days (Choi et al., 2009; Thalmann et al., 2014, 2016). Theoretically, sulfidation of Ag NPs should be the only process to exclusively occur in anaerobic environments, where the sulfide source is rich, such as the anaerobic zones of the WWTP (Kim et al., 2010; Kaegi et al., 2013; Kent et al., 2014). Interestingly, Ag NPs sulfidation had been observed under oxic conditions, without free bisulfides, by reaction with the metal sulfides ZnS and CuS (Thalmann et al., 2014). Different with in anaerobic environments, sulfidation of Ag NPs in oxic conditions requires both a sulfide source and an oxidation agent. Where the dissolution-precipitation sulfidation pathway was found, it followed pseudo first-order formation kinetics. This dissolutionprecipitation pathway indicates that oxidative dissolution is the first step. However, the oxidative dissolution is not for releasing S2 or HS, but for releasing Agþ. Indeed, the Ag2S particles formed through Agþ displeasing Cu2þ and Zn2þ. The DO is still very important for the Ag NPs sulfidation in oxic water, because of the releasing Agþ is slow and a rate determining step. Due to the excessive amounts of CuS and ZnS, the oxidative dissolution is a rate-determining step, and complete sulfidation of Ag NPs can be expected without free sulphides (Thalmann et al., 2014). Clearly, laboratory studies are meaningful for realistic environmental studies, but they cannot fully present the pattern of the natural environment. Several factors affect the reaction rate or the extent of Ag NPs sulfidation, including particle size, DOMs, and the HS/Ag ratio, among others. Increasing extents or reaction rates of Ag NPs sulfidation with decreased particle size have been observed (Liu et al., 2011; Kaegi et al., 2013; Thalmann et al., 2014, 2016; Collin et al., 2016). For example, 20 and 40 nm Ag NPs were found to be completely sulfidized (with 1000 mg/L HA, 15 min reaction duration), whereas only 70 and 80% of 100 and 200 nm Ag NPs were sulfidized, respectively (Thalmann et al., 2016). This finding can be explained by the dependence of the extent or reaction rate on the specific surface area. A small particle size results in a large specific surface area, although small Ag NPs tend to aggregate strongly. However, the sulfidation of large nanosilver particles (200 nme2 mm) could proceed to completion (Liu et al., 2011; Thalmann et al., 2016). Opposite patterns were reported for the effect of DOMs on the sulfidation of Ag NPs. One pattern is that humic acid (slightly) increased the sulfidation extent or rate of Ag NPs (Liu et al., 2011; Thalmann et al., 2016). The explanation for this finding is the replacement of the surface coating by DOMs, leading to an extensive Ag NPs surface area available for the sulfidation reaction (Levard et al., 2011; Thalmann et al., 2016). Moreover, DOMs attached to the surface of Ag NPs can serve as a transport channel to bring HS in close contact with Ag NPs (Thalmann et al., 2016). The other pattern is that Suwannee River Fulvic Acid (SRFA) slightly decreased the sulfidation rate of Ag NPs (Baalousha et al., 2015), possibly because the DOMs prevents sulfidation reaction of

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Ag NPs by blocking the surface of Ag NPs, which would be the same as dissolution inhibition, especially for indirect sulfidation mechanism. These two patterns can be partially supported by the influence of DOMs on Ag2S NPs stability, while the presence of NOM either inhibited, had no effect on, or promoted Agþ release from Ag2S NPs (Collin et al., 2016; Li et al., 2016a). Extent or reaction rates of Ag NPs sulfidation are strongly dependent on the HS/Ag ratio, as shown by an increasing sulfide to Ag NPs ratio increasing the formation of Ag2S NPs (Levard et al., 2011; Liu et al., 2011; Reinsch et al., 2012; Thalmann et al., 2014). The phenomenon can be supported by sulfidation of Ag NPs following pseudo first-order kinetics (Liu et al., 2011; Thalmann et al., 2014). Interestingly, a systematic study comparing four different models found that the sulfidation rate of Ag NPs can be best described employing a diffusion-limited solid-state reaction model (equation (7)) but not the pseudo first-order rate model (Thalmann et al., 2016). Clearly, there are other factors that can affect the sulfidation of Ag NPs, such as pH, particle structure, type of sulfide (natural vs laboratory vs metallic sulfide) and DO (Liu et al., 2011; Lowry et al., 2012a; Reinsch et al., 2012; Thalmann et al., 2014).

F¼

rffiffiffiffiffiffiffiffiffiffiffiffiffi !3 k 1  t r2

(7)

F is the metallic Ag fraction (determined by XAS), r is the radius of the Ag NPs (determined by TEM), k is the rate constant and t is the time (Thalmann et al., 2016).

2.2.2. Shape and size change Multiple morphologies of Ag2S NPs were detected, which essentially relate to different sulfidation mechanisms of Ag NPs. In a direct sulfidation mechanism, a typical core (Ag NPs)-shell (Ag2S NPs)-type structure was observed and was followed by residual Ag NPs converting completely to Ag2S NPs during the prolonged reaction (Thalmann et al., 2016). Normally, the shape change of Ag2S NPs is not uniform in sulfidation processes. Ag2S NPs crystallites form as discontinuous, irregular protrusions and amorphous formations on the silver surface (Liu et al., 2011; Kent et al., 2014). Reasonably, the particle height of Ag2S increases compared to that of Ag NPs, which creates a passivation layer on Ag2S surface (Kent et al., 2014). Particularly, excess sulfides could adhere to surface of Ag2S NPs under S-rich aquatic environments, resulting in an increased height of Ag2S (Kim et al., 2010). Furthermore, Ag2S nanocrystals have been observed to have an ellipsoidal shape, formed into very small, loosely packed aggregations (Kim et al., 2010). In an indirect sulfidation mechanism, Ag2S formed as a remarkable chain-like fractal structure or nanobridges between Ag NPs (Levard et al., 2011, 2012). Smaller Ag2S particles have typically been detected in a precipitation process (Thalmann et al., 2014; Li et al., 2016a), which differs from the height growth of Ag2S particles detected in a direct formation mechanism. Interconversion of these morphologies occurs, although the morphologies of Ag2S are strongly related to formation mechanisms. For example, in the presence of HA, Ag NPs initially sulfidized to form core-shell particle structures that further converted into hollow Ag2S particles after 45 min (Thalmann et al., 2016). As the sulfidation rate increases, the Ag2S structures transform from nanobridge structures to amorphous and unshaped structures (Levard et al., 2011). Other factors, such as HA, can also affect the morphology formation of Ag2S. Without HA, Ag NPs sulfidized from one side to form asymmetrical structures, whereas the symmetrical core-shell structures occurred and were further transformed into hollow Ag2S in the presence of HA (Thalmann et al., 2016).

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2.2.3. Toxicity Sulfidation transformation limits the toxicity of Ag NPs towards a range of organisms (Choi et al., 2009; Levard et al., 2012, 2013a; Reinsch et al., 2012; Devi et al., 2015; Doolette et al., 2015; Starnes et al., 2015; Stegemeier et al., 2015; Collin et al., 2016). For example, sulfidation had been considered a natural antidote for Ag NPs, reflected by even partial sulfidation can dramatically decrease the toxicity of Ag NPs (Levard et al., 2013a). Sulfidation has resulted in a significant alleviation of changes in liver oxidative stress, detoxification enzymes and brain acetylcholinesterase activities that were affected by Ag NPs exposure (Devi et al., 2015). When small aliquots of sulfide were added to the exposure, the toxicity of Ag NPs towards nitrifying bacteria was inhibited by up to 80% (Choi et al., 2009). The antidote mechanisms of Ag NPs through sulfidation can be summarised as follows. (I) Sulfidation represses the ion-related toxicity of Ag NPs by decreasing amount of existing free Agþ and/or further Agþ release. The released Agþ is widely believed to, at least partly, contribute to Ag NPs toxicity. Ag2S NPs are thermodynamically more stable than are Ag NPs, and even a small fraction of sulfidation of Ag NPs can significantly reduce the free Agþ (Levard et al., 2013a). Therefore, in the case of existing free Agþ, sulfidation directly decreases absolute concentration of Agþ, resulting in mitigating toxicity. For the case of further released Agþ, new formed Ag2S can block the surface of Ag NPs, which in turn lead to decrease relative concentration of Agþ and decrease the toxicity. For example, over 10 times more free Agþ was released from pristine Ag NPs than those released from aged Ag2S NPs (Starnes et al., 2015). (II) Sulfidation suppresses the particle-related toxicity by modifying particle parameters, such as the shape, particle size, surface area or surface charge, of Ag NPs. As mentioned above, smaller particle sizes or higher heights of Ag2S NPs could be formed, which can lead to different effects on toxicity. Moreover, strong aggregations have been found during sulfidation (Levard et al., 2011; Liu et al., 2011; Reinsch et al., 2012), and these aggregations normally decrease the toxicity of Ag NPs (Kvítek et al., 2008). As the shape changed or was lost during sulfidation, the surface charge of Ag NPs varied (Levard et al., 2011). The toxicity of Ag NPs to bacteria has been shown to be strongly dependent on the surface charge (El Badawy et al., 2011). Moreover, surface charge plays an important role in determining how effective the interaction is between NPs and negatively charged bacterial surfaces (Choi and Hu, 2008; Khan et al., 2011). The modification of surface charge is also speculated to closely correlate with the toxicity of Ag2S NPs, although no systematic investigation has been reported. Sulfidation as a natural antidote to the toxicity of metallic nanoparticles (NPs) may be overestimated. Remarkably, sulfidation did not reduce the toxicity of CuO NPs to medaka (Oryzias latipes) embryos (Li et al., 2015) in which free Cu2þ released from CuS NPs was higher than those from CuO NPs, leading to a significant increase in oxidative stress in fish embryos. The re-releasing of Agþ from Ag2S NPs is possible, when Ag2S NPs are oxidised by Fe(III) to from Fe(II) in water under sunlight conditions (Li et al., 2016b). Thus, although oxidations of Ag2S to release Agþ have not been reported, sulfidation serving as a natural antidote should be carefully considered in environmental safety and risk assessment. The toxicity potencies of Ag NPs can remain during or after sulfidation. Different Ag NPs toxicity potencies can be exhibited during the two sulfidation mechanisms. In direct sulfidation, the sulfidation is not uniform; it is asymmetrical and starts from one side of the Ag NPs (Thalmann et al., 2016), implying that part of the unsulfidized Ag NPs have potential toxicity to organisms. The DO can aid Agþ release from Ag NPs via oxidative dissolution, although this process is not predominant. In indirect sulfidation, the released Agþ can be temporarily available for bioup into account during the oxidative dissolution process (Liu et al., 2011). Although sulfidation

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can significantly reduce the toxicity of Ag NPs, sulfidized Ag NPs show toxic influences to a range of organisms and present various toxicity mechanisms compared with those of Ag NPs (Starnes et al., 2015; Stegemeier et al., 2015; Wang et al., 2015; Collin et al., 2016; Starnes et al., 2016). For example, Ag2S NPs induce a toxic effect through direct internalisation into Caenorhabditis elegans cuticles or accumulation in plant tissues, neither of which depend on the released free Agþ (Starnes et al., 2015, 2016; Stegemeier et al., 2015; Wang et al., 2015). Comparably, Ag NPs induce toxicity depending on both ion-related and particle-related toxic effects. In a comparison with short-term (24 h) exposures, Ag2S NPs only showed toxic effects to plants in a long-term (2 weeks) investigation (Wang et al., 2015), suggesting the importance of long-term safe risk assessment. Importantly, the authors proposed that the internalised or accumulated Ag2S NPs may act as a Trojan horse continuously releasing Agþ (Wang et al., 2015). Notably, the photoinduced transformation of Ag2S NPs in the presence of Fe3þ elevates free Agþ and results in the reconstitution of smaller Ag2S NPs in the environment (Li et al., 2016a). In summary, direct and indirect mechanisms are involved in Ag NPs sulfidation transformations, and they yield different morphologies and potential toxicities of Ag2S NPs. The aquatic toxicity of Ag NPs in the environment can be reduced by sulfidation through modification of the ion-related and/or particle-related toxicities. However, sulfidation as an antidote to the toxicity of Ag NPs could be overestimated under certain environmental conditions, such as under Fe3þ-rich (in dark or light) aqueous conditions. Sulfidised Ag NPs are still bioavailable and show toxic effects to organisms, which present distinct toxicity mechanisms compared with those of Ag NPs. 2.3. Chlorination of Ag NPs and their toxic impact 2.3.1. Mechanisms In the environment, the chlorination of Ag NPs is reported via an indirect mechanism, indicating that oxidative dissolution of Ag NPs to release free Agþ is a first step, followed by the precipitation of Agþ with Cl. The indirect mechanism of chlorination was supported by low concentrations of chloride (0.1 mg/L) mitigating the toxicity of Agþ but not the toxicity of Ag NPs (Xiu et al., 2011). Chlorination is strongly affected by Cl/Ag ratio and redox conditions (Levard et al., 2012). Under aerobic (oxidizing) conditions, silver chloride species are predicted to form, while under anaerobic (reducing) conditions, precipitation of Ag2S is supposed. Typically, nanoscale AgCl can form on the surface of Ag NPs with a core (Ag NPs)-shell (AgCl NPs)-type structure at a low Cl/Ag ratio (Li et al., 2010), in which the shell (AgCl) normally inhibits the further dissolution of Ag NPs. In contrast, the shell (AgCl) converts to dissoluble species of AgCl(x1) at a high Cl/Ag ratio, in which x further dissolution of Ag NPs could be enhanced. Thermodynamically, in the freshwater system (0.01 M Cl; low Cl/Ag ratio), AgCl and Ag0 are possible phases, whereas in the seawater system (0.5 M 0 Cl; high Cl/Ag ratios), AgCl 2 and Ag are the most likely species, as 3 well as AgCl2 and AgCl possibly. According to the MINEQL 3 4 model, silver nitrate in freshwater is present as 60% AgCl0, 34% Agþ, and 5% AgCl2 (Meyer et al., 2010; Luo et al., 2016). To data, a Cl/Ag ratio of 2675 seems to be a threshold to distinguish low and high Cl/ Ag ratios (Levard et al., 2013b). For Cl/Ag  2675, the dominant species is solid AgCl during the chlorination, resulted limitation of dissolution of Ag NPs. For Cl/Ag  2675, the dominant species are AgCl(x1) species (see Scheme 4), resulted promotion of dissolux tion of Ag NPs. Previously, a dissolutionerecrystallisation mechanism was proposed to explain the shape change of Ag NPs when in contact with chloride (Yang et al., 2007). An in-plane Ag NPs shape change

from triangular to circular was detected when the Ag NPs were exposed to 10 nM Cl, and steepened sidewalls as well as a height increase of 6e12 nm were found when using atomic force microscopy (Kent and Vikesland, 2012). As mentioned above, the Cl/Ag ratio significantly affects the formation of AgCl NPs; the shape change could also be strongly driven by the Cl/Ag ratio. Typically, the core (Ag NPs)-shell (AgCl NPs)-type structures are found at low Cl/Ag ratios, suggesting that the process occurred on the surface of Ag NPs and that released Agþ more likely adhered to Ag NPs surfaces. Due to the low solubility of AgCl, further chlorination could be blocked. In addition, chain-like structures have also been observed (Li et al., 2010), with a similar structure to that found to form when Ag NPs react with HS via the indirect mechanism. As the Cl/Ag ratio increased, the formed solid shell (AgCl) adsorbs excess Cl and becomes dissoluble, resulting in the dominant formation of AgCl(x1) escaping from the surface. x 2.3.2. Toxicity Chlorination strongly decreases the toxicity of Ag NPs. This phenomenon can be explained by chlorination significantly decreasing the dissolved bioavailable silver species Agþ (Levard et al., 2013a). The formed AgCl(s) decreased not only the Agþ but also further Agþ release. Although the presence of excess chloride may result in high concentration of dissoluble Ag species (AgCl(x1) ) rather than AgCl(s), the toxicity of Ag has been shown x to decrease compared to without of chloride (Leblanc et al., 1984; Lee et al., 2005). Indeed, elevated Cl concentrations resulted in increased Ag accumulation in plant tissues (over a two week exposure), specifically in shoot tissues, with 11.5 times higher Ag concentrations (Wang et al., 2016). This level of higher Ag was not phytotoxic, with no significant dependence on the mass of shoot tissue. Interestingly, different pH conditions (from 5.4 to 7.1) showed no influence on the uptake of AgCl(x1) by shoot tissues x (Wang et al., 2016). 2.4. Reduction of Agþ to Ag NPs and toxic impact 2.4.1. Mechanisms Numerous studies have reported that Agþ can be chemically reduced to Ag NPs under high-energy irradiation. In the environment, free Agþ can be reduced by DOMs to form Ag NPs under (simulated) sunlight irradiation. However, solar or simulated light exposure by itself can accelerate the formation of Agþ by speeding up the oxidative dissolution of Ag NPs (via the formation of Ag2O (Li € mer et al., 2016)) in water, resulting in no re-formed et al., 2014; Ro Ag NPs (Yin et al., 2012). DOMs was extensively found, which can improve the stability of Ag NPs and block Agþ release by coating the surface of Ag NPs. However, DOMs can reduce Agþ to form Ag NPs at a low rate in the dark (Sal'nikov et al., 2009; Yin et al., 2012). The reductive formation rate of Ag NPs with DOMs increased significantly under sunlight irradiation. Considering that sunlight and DOMs normally coexist in real environments, the reductive reformation of Agþ to Ag NPs is high significance. The reductive formation mechanism of Agþ to Ag NPs is reduction-nucleationgrowth and fusion (see Scheme 5 and equation (8)) (Yin et al., 2015). Wen-Che Hou et al. proposed that the reduction of Agþ by DOMs to form new Ag NPs was a pseudo-first order reaction (Hou et al., 2013). UV

Agþ þАgn þ FA!Agnþ1 þ FAðOXÞ

(8)

In a reduction process, the functional groups in DOMs (such as phenol, quinone, ketone and hydroxyl groups) were photo
irradiated by sunlight, producing superoxides (e.g., O2, H2O2 and triplet NOMs). These active superoxides acted as electron donors to

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331

Scheme 4.

Scheme 5.

reduce free Agþ acting as an electron acceptor to small Ag NPs. Free Agþ can be reduced to Ag NPs by these superoxides via a chargingdischarging mechanism (He et al., 2011). Free radical and dissolved O2 significantly enhanced the formation rate of Ag NPs, thereby supporting this view (He et al., 2011; Jones et al., 2011; Yin et al., 2012; Adegboyega et al., 2013). Additionally, a dissolved O2independent formation mechanism has also been found (Hou et al., 2013). Although both light exposure and NOMs were required for Ag NPs re-formation, the varying concentration of DO did not significantly change Ag NPs formation rates. Importantly, the reduction of free Agþ to Ag NPs by Fe(II) without of sunlight irradiation was reported lately (Li et al., 2016b). Re-formed small Ag NPs served as nucleuses with double “faces” in the nucleation process. The one “face” is that free Agþ adsorbs onto the surface of Ag NPs (core) (Li and Lenhart, 2012; Adegboyega et al., 2013; Yu et al., 2014) and is further catalysed by DOMs. The other “face” is that organic free radicals produced under sunlight irradiation transfer electrons to the Ag NPs, serving as an electron

pool for the further growth of Ag NPs (Henglein, 1979; He et al., 2011; Jones et al., 2011). A greater ability of Ag NPs to accept and store electrons results in a faster reduction of Agþ to form Ag NPs. The presence of Ag NPs can significantly enhance the reduction rate of Agþ (Jones et al., 2011), and the rate increased by at least 4 orders of magnitude compared to that in the absence of Ag NPs (He et al., 2011). In the growth and/or fusion processes, changes in the shape and size of the initial Ag NPs normally co-exist because the growth of re-formed Ag NPs and the oxidative dissolution of the initial Ag NPs occur simultaneously (Manoharan et al., 2014; Zou et al., 2015). These processes dominate the morphological changes of Ag NPs and are detailed below. DOMs, irradiation and pH are among the most essential factors driving the reductive re-formation of Ag NPs. Light irradiation activates DOMs and provides reaction energy for the formation; the formation rate increases significantly under sunlight irradiation. Therefore, more intense light (l > 300 nm) and long exposure times can greatly enhance the photoreduction re-formation rate of Ag NPs (Hou et al., 2013; Yin et al., 2014). DOMs with diverse functional groups show distinct photoreduction rates of Ag NPs. For example, Adegboyega, N. F. et al. reported that different FAs and Suwannee River humic acid (SRHA) followed a decreasing reductive formation of Ag NPs order of NLFA > SRHA > PPFA > SRFA (Adegboyega et al., 2013). Functional groups, such as phenols, in DOMs served as free radical producers regulating photoreduction. Thus, the different nitrogen and sulfur contents and radical characteristics of these DOMs resulted in varying growth rates of the re-formation of Ag NPs. Without these functional groups, Ag NPs were formed under sunlight irradiation alone (Yin et al., 2012). Indeed, the photoreduction formation of Ag NPs was almost completely suppressed by employing DOMs with blocked phenolic groups (Yin et al., 2012). This phenomenon is involved in two possible pathways: direct charge transfer from the phenol group to the metallic ion in a phenol-metal ion complex and the reaction of dissolved oxygen with the phenol group to produce the superoxide anion free radical
(O2), which reduces metallic ions to low-valent metal. Several

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studies have reported that photoreduction formation of Ag NPs is favourable at high pH (Yin et al., 2012, 2014; Yu et al., 2014; Zhang et al., 2015). Wen-Che Hou, et al. proposed that the pH-dependence of Ag NPs re-formation can be explained by the higher binding ability of Agþ to humic acid at high pH than that at low pH, which promotes Agþ reduction (Hou et al., 2013). The re-formed Ag NPs should be more thermodynamically stable at higher pH than at lower pH, and it is presumed that the equilibrium of photoreduction is biased in favour of Ag NPs formation. In addition, other factors, such as redox potentials and dissolved oxygen and temperature, can also affect the reductive formation of Ag NPs (Yin et al., 2012, 2014; Adegboyega et al., 2013). 2.4.2. Shape and size change Multiple morphologies were observed in the reductive formation of Ag NPs and were strongly driven by formation processes. During the reduction process, small Ag NPs were frequently observed due to the reduction of ions to particles. For example, very small spherical (or semispherical) particles were found in a mixture of DOMs and Ag NPs (Yin et al., 2014; Zou et al., 2015). Additionally, particle sizes have been found to be reduced from 26 to 18 nm and from 15 to 2e6 nm during photoreduction (Yu et al., 2014, 2016). Notably, the detected small Ag NPs could be the initial particles that release free Agþ during the primary irradiation. Various morphologies appear in the processes of nucleation, growth and fusion. Summarily, spherical colloids were the dominant structure in the growth stage. With prolonged irradiation, triangular nanoprisms appeared, which later disappeared gradually while large irregular aggregations appeared (Yin et al., 2012; Zou et al., 2015). These structural changes were not observed in the dark. Interestingly, chain-like or cross-linking structures were detected (Cheng et al., 2011; Li and Lenhart, 2012; Yin et al., 2012, 2015), and they were also detected in the indirect sulfidation of Ag NPs. Because the both transformations have free Agþ in common. 2.4.3. Toxicity Considering that free Agþ is believed to cause the toxic effect of Ag NPs to a range of organisms, the photoreduction re-formation of Agþ to Ag NPs could partially decrease toxicity of Ag NPs. By decreasing the amount of Agþ, photoreduction of Agþ to Ag NPs has been shown to reduce the toxicity of the initial Ag NPs (Cheng et al., 2011; Zhang et al., 2015). Notably, oxidative dissolution and reductive formation of Ag NPs should occur simultaneously, indicating that the toxicities of Agþ released from Ag NPs remain possible. UV irradiation has been demonstrated to significantly enhance the toxicity of Ag NPs compared to that in the dark, which was explained by accelerated Agþ release and ROS generation (Li et al., 2014), which both exist in the photoreduction mechanism. However, light irradiation weakened toxicity of Ag NPs were also found (Shi et al., 2012, 2013). Considering the particle-related toxic effect, the toxicity of Ag NPs could become severe during photoreduction. This view is supported by the following facts. (I) The amount of Ag NPs increase as the Agþ concentration decreases. (II) Compared with big Ag NPs, the newly formed small Ag NPs could induce more serious toxicity during the primary photoreduction stage. (III) The newly formed bare-Ag NPs could yield enhanced toxicity compared with the initial Ag NPs coating. (IV) The newly formed Ag NPs are more stable than are the initial particles, resulting in a high and long-term toxicological threat. In summary, the reduction of Agþ to form Ag NPs is an important transformation process for the fate and toxicity of Ag NPs. During natural reduction, the shape, surface properties and stability of the initial Ag NPs could be changed, which can significantly affect the bioavailability and toxicity of silver in the environment.

3. Correlation of environmental transformations of Ag NPs and their common influencing factors Thermodynamically, Ag NPs are unstable in complex physiological or natural environments and trend to transform to other dominant silver species under certain conditions. For instance, Ag NPs were not persistent in real environmental compartments containing dissolved oxygen, based on thermodynamic analysis and kinetic measurements (Liu and Hurt, 2010). However, complete dissolution was not observed when employing different river waters and sizes of Ag NPs (Dobias and Bernier-Latmani, 2013). The oxidative dissolution and reduction processes of Ag NPs are coupled in natural water. Thus, generally, Ag NPs and their chemical transformation products jointly exert a hazardous effect to target organisms, as detected. As noted recently, Ag2S NPs could be the dominant silver species, and determine hazardous toxicity under most anaerobic environmental conditions if there are high concentrations of sulphide (Collin et al., 2016; Starnes et al., 2016). Notably, once Ag2S particles are formed under oxic conditions, they may also be dominant specie lying on their stability in oxic waters. The oxidative dissolution, photoreduction, sulfidation and chlorination of Ag NPs are dominant transformations and not independent in aquatic environments. Oxidative dissolution, however, is a vital and primary step for the other transformations (except a direct sulfidation pathway). Correspondingly, free Agþ as an intermediate is retained into the next transformations and possibly influences the uptake, bioaccumulation, and adverse biological or ecological implications (Liu et al., 2011). These environmental transformations of Ag NPs progress simultaneously in complex aquatic environments. An excellent example is oxidative dissolution and reduction or photoreduction of Ag NPs occurring simultaneously, whereby the dissolution and precipitation of Ag NPs compete until reaching equilibrium. Indeed, the required environmental factors, such as DO, sunlight and DOMs, are very common and coexist in aquatic environments. Sulfidation and chlorination of Ag NPs is another example of simultaneous processes. Ag2S and AgCl have been observed, respectively, under anaerobic conditions after Ag NPs introduced, although Ag2S is more stable than is AgCl (Lombi et al., 2013; Sekine et al., 2015; Wang et al., 2016). The processes of dissolution and chlorination of Ag NPs coexist at the intersection of fresh (river) and salt (sea) water. Many environmental factors coexist during the environmental transformations of Ag NPs. Excellent examples of these factors are light, DOMs, pH and DO. Light exposure can accelerate the oxidative dissolution of Ag NPs to Agþ. Interestingly, sunlight irradiation can significantly accelerate the dissolution of Ag NPs (Yu et al., 2016), and free Agþ can be reduced to Ag NPs with DOMs by sunlight irradiation. Interestingly, sunlight can accelerate dissolution of Ag2S NPs to release Agþ by Fe(III), and the released Agþ can be reduced to Ag NPs by Fe(II) without of sunlight irradiation are needed (Li et al., 2016b). The light irradiation had been also found inducing the growth of Ag NPs and sequentially forming bulk agglomeration, and thus decrease the dissolution of Ag NPs (Shi et al., 2012, 2013). Moreover, aggregation of the initial Ag NPs (Yu et al., 2016) and growth and aggregation of the re-formed Ag NPs could occur simultaneously under sunlight irradiation; these morphological changes were barely distinguishable (Li and Lenhart, 2012). DOMs can promote the re-formation of Ag NPs, indicating that Ag NPs are stable in the condition. Under this condition, the amount of Agþ could be decreased through blocking the oxidation sites of Ag NPs (Liu and Hurt, 2010). The pH value creates opposing effects on dissolution and reduction transformations. Acidic conditions promote the oxidative dissolution of Ag NPs, as reflected by the protons playing a critical role in that process. Basic conditions (pH ¼ 8.5) can completely inhibit oxidative dissolution (Yu et al.,

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2016) and accelerate the reduction of Agþ to Ag NPs under sunlight irradiation. This finding could be explained by high pH maintaining the stability of re-formed Ag NPs and the higher binding ability of Agþ to humic acid at a high pH that promotes Agþ reduction (Hou et al., 2013). The dissolved oxygen is a vital factor controlling the dissolution of Ag NPs in aquatic environments. As the dissolution is the primary step for the reduction, sulfidation and chlorination transformations, the DO also plays an important role in these transformations. For example, in the indirect sulfidation transformation, DO governs the oxidative dissolution of Ag NPs, which is a rate-determining step (see equation (6a)). 4. Environmental significances Based on these data, the toxic mechanisms of Ag NPs to aquatic organisms are still not fully understood. Clearly, the toxicological response to Ag NPs in organisms links both silver ions (ion-related toxicity) and silver particles (particle-related toxicity and/or the Trojan horse effect). Additionally, Ag NPs are more likely to transform in physically complex water environments and are likely to remain in a dynamic equilibrium; thus, the transformation products that replace Ag NPs are the predominant species and mechanistically induce toxic effects under certain aquatic conditions. Moreover, some of the transformation products, such as Ag2S and AgCl, possess high stability, resulting in long-term storage and high transportability in the environment. Therefore, Ag NPs and their chemical transformation products are possibly retained for a long time in real environments and do not undergo only a sole transformation. The potential toxicity of Ag NPs in the real environment is more complex than that under certain water or laboratory conditions. In addition to the biouptake, nanoproperties and environmental influencing factors, the environmental transformations of Ag NPs and their chemical transformation products should be seriously considered when deriving a complete assessment of the environmental risk of Ag NPs. Notes The authors declare no competing financial interest. Acknowledgments We gratefully acknowledge financial support from the National Natural Science Foundation of China grant (No.21477159). Thanks to the anonymous reviewers for their comments which greatly improved the manuscript. References Adegboyega, N.F., Sharma, V.K., Siskova, K., Zboril, R., Sohn, M., Schultz, B.J., Banerjee, S., 2013. Interactions of aqueous Agþ with fulvic acids: mechanisms of silver nanoparticle formation and investigation of stability. Environ. Sci. Technol. 47, 757e764. Baalousha, M., Arkill, K.P., Romer, I., Palmer, R.E., Lead, J.R., 2015. Transformations of citrate and Tween coated silver nanoparticles reacted with Na2S. Sci. Total Environ. 502, 344e353. Benn, T.M., Westerhoff, P., 2008. Nanoparticle silver released into water from commercially available sock fabrics. Environ. Sci. Technol. 42, 4133e4139. Bondarenko, O., Juganson, K., Ivask, A., Kasemets, K., Mortimer, M., Kahru, A., 2013. Toxicity of Ag, CuO and ZnO nanoparticles to selected environmentally relevant test organisms and mammalian cells in vitro: a critical review. Arch. Toxicol. 87, 1181e1200. Cheng, Y., Yin, L., Lin, S., Wiesner, M., Bernhardt, E., Liu, J., 2011. Toxicity reduction of polymer-stabilized silver nanoparticles by sunlight. J. Phys. Chem. C 115, 4425e4432. Choi, O., Clevenger, T.E., Deng, B., Surampalli, R.Y., Ross Jr., L., Hu, Z., 2009. Role of sulfide and ligand strength in controlling nanosilver toxicity. Water Res. 43, 1879e1886. Choi, O., Hu, Z., 2008. Size dependent and reactive oxygen species related nanosilver toxicity to nitrifying bacteria. Environ. Sci. Technol. 42, 4583e4588.

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