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Published in final edited form as: Chemosphere. 2024 May 24;360:142384. doi: 10.1016/j.chemosphere.2024.142384

Environmental Significance of PAH Photoproduct Formation: TiO2 Nanoparticle Influence, Altered Bioavailability, and Potential Photochemical Mechanisms

Lindsey St Mary a,b, Lisandra S D Trine b, Courtney Roper b,c, Jackson Wiley b, Luca Craciunescu d, Lia Sotorrios d, Martin Paterson d, Staci L Massey Simonich b, Martin McCoustra d, Theodore B Henry a
PMCID: PMC11321274  NIHMSID: NIHMS2000745  PMID: 38797205

Abstract

Interactions between polycyclic aromatic hydrocarbons (PAHs) and titanium dioxide (TiO2) nanoparticles (NPs) can produce unforeseen photoproducts in the aqueous phase. Both PAHs and TiO2-NPs are well-studied and highly persistent environmental pollutants, but the consequences of PAH-TiO2-NP interactions are rarely explored. We investigated PAH photoproduct formation over time for benzo[a]pyrene (BaP), fluoranthene (FLT), and pyrene (PYR) in the presence of ultraviolet A (UVA) using a combination of analytical and computational methods including, identification of PAH photoproducts, assessment of expression profiles for gene indicators of PAH metabolism, and computational evaluation of the reaction mechanisms through which certain photoproducts might be formed. Chemical analyses identified diverse photoproducts, but all PAHs shared a primary photoproduct, 9,10-phenanthraquinone (9,10-PQ), regardless of TiO2-NP presence. The computed reaction mechanisms revealed the roles photodissociation and singlet oxygen chemistry likely play in PAH mediated photochemical processes that result in the congruent production of 9,10-PQ within this study. Our investigation of PAH photoproduct formation has provided substantial evidence of the many, diverse and congruent, photoproducts formed from physicochemically distinct PAHs and how TiO2-NPs influence bioavailability and time-related formation of PAH photoproducts.

Graphical Abstract

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Introduction

Polycyclic aromatic hydrocarbons (PAHs) and titanium dioxide (TiO2) nanoparticle (NP) are photoactive environmental contaminants that typically deposit in aquatic environments (1, 2, 3). Higher molecular weight (HMW) PAHs (4–7 aromatic rings) such as benzo[a]pyrene (BaP), fluoranthene (FLT), and pyrene (PYR), form through the incomplete combustion of carbonaceous materials (4, 5). BaP, FLT, and PYR typically enter the aquatic environment through urban water run-off and atmospheric deposition, while TiO2-NPs are often released through the use of personal care and consumer products like sunscreen, makeup, paints, and electronics (5, 6). TiO2-NPs are the most widely produced NP due to their numerous applications which is enhanced by their modification versatility (7, 8, 9, 10, 11, 12). TiO2-NPs serve as potent photocatalysts due to the production of ·OH radicals upon photoactivation, this is particularly valuable for remediating heavily polluted sites, as well as in water treatment applications (7, 8, 10, 11, 12, 13). Additionally, their photocatalytic abilities can be further improved through various synthesis methods (11). The low aqueous solubility of HMW PAHs enhance adsorption to NP surfaces, and with the continued release of both PAHs and TiO2-NPs this increases the likelihood of interactions in the aquatic environment (14, 15, 16). PAH interactions with TiO2-NPs can influence the environmental fate, transport, phototransformation, and bioavailability of PAHs and their photoproducts in the aquatic environment. There has been considerable effort to assess toxicity of TiO2-NPs and PAHs individually, which has resulted in legislation that restricts the amount allowed in consumer products and enforces monitoring activities. However, substituted PAHs that may result from phototransformation, like oxygenated PAHs (OPAHs) and hydroxylated PAHs (OHPAHs), are not monitored in the environment nor considered during remediation processes (17, 18). Furthermore, OPAHs and OHPAHs have been identified at aquatic remediation sites, in the atmosphere and groundwater, as well as in aquatic species has led to toxicity assessments which have revealed that OPAHs and OHPAHs exhibit significant toxicity (12, 19, 20, 21, 22, 23, 24, 25, 26).

Research on PAH photoproduct formation is limited, and photoproducts are typically considered only as parent-derived or oxidized intermediates progression towards mineralization during PAH photodegradation, if they’re considered at all.(15, 27, 28, 29, 30). However, this approach overlooks the formation of diverse photoproducts that might elicit the greatest impact on human health and the environment (21, 22, 23, 31, 32, 33). Furthermore, there is a lack of information about photoproducts that form after photocatalytic processes end. Most studies on PAH photocatalysis and phototoxicity concentrate on processes that occur only during photocatalytic events. Employing a multidisciplinary approach, involving chemical analyses of various PAH derivatives and analytical biology to evaluate changes in bioavailability, will improve our understanding of photochemical processes and their implications for environmental and human health.

Zebrafish (Danio rerio) have been used extensively to assess toxicity and explore molecular mechanisms of PAH metabolism because of this, expression of many genes encoding for enzymes responsible for PAH metabolism are well-characterized (21, 22, 34, 35). Previous studies have utilized this genetic specificity in the zebrafish to explore altered bioavailability of PAHs and their subsequent by-products, particularly in the presence of TiO2-NPs (31, 36, 37). The use of the zebrafish as an indicator of PAH bioavailability through the analyses of gene profiles, combined with analytical chemistry and computational techniques, can provide novel insight into the interplay between PAH photoproduct formation and TiO2-NP influence on the bioavailability of subsequent reaction products throughout in the aqueous phase.

The overall aim of this study was to investigate changes in BaP, FLT, and PYR bioavailability and photoproduct formation upon interaction with TiO2-NPs following UVA irradiation in the aqueous phase. BaP, FLT, and PYR were chosen due to their distinct physiochemical characteristics (i.e., BaP contains a bay region, FLT contains a 5-membered aromatic ring, and PYR has a clustered ring arrangement) (Figure S1). Specific objectives were to: 1) identify diverse photoproducts that formed through GC-MS analyses; 2) measure time-related expression profiles of specific genes (cyp1a, ephx1, ephx2, and sod1) in larval zebrafish exposed to UVA irradiated BaP- FLT-and PYR- TiO2-NP reactions; 3) computationally evaluate potential mechanisms of phototransformation for 9,10-PQ; and 4) integrate the results of the different approaches to enhance our understanding of PAH photoproduct formation.

Materials and Methods

Chemicals, Nanoparticles, and Materials

Stock of BaP (>96%, SKU-B1760), FLT (>99%, SKU-F4418), and PYR (99%, SKU-571245) were purchased from Sigma Aldrich. The TiO2-NPs JRCNM01005a (NM105) used in all experiments were obtained from the EU nanoparticle repository. TiO2-NP physical properties and overview of all materials, PAHs, and reagents are described in the Supporting Information.

Solution Preparation and Collection

Primary stocks of 400 mg/L BaP, FLT, and PYR were made in dimethyl sulfoxide (DMSO) due to the low aqueous solubility of PAHs and TiO2-NP stock was dispersed in reverse osmosis (RO) water at 500 mg/L (38). Primary stocks were then used to make working solutions of 20 μg/L BaP, 70 μg/L FLT, 80 μg/L PYR alone and in combination with 2 mg/L TiO2-NP in OECD media (Organization for Economic Co-operation and Development-OECD guideline 210) on the day of the experiment. Final concentrations of DMSO in both exposure and chemistry solutions were no higher than 0.2% to avoid both hydroxyl radical quenching and toxicity in the zebrafish exposure solutions (39, 40, 41) . Samples meant for chemical analyses did not contain zebrafish larvae in order to specifically assess photoproducts resulting from photochemical processes only. UVA irradiated samples were irradiated for 40 minutes at an intensity of 7.26 mW/cm2. PAH and TiO2-NP concentrations were chosen according to concentrations found in diverse water sources (Table S1). The experimental layout and timeline for exposure and chemistry collections, as well as solution preparations and concentrations for all PAHs and TiO2-NPs can be found in Supporting Information (Figure S2).

Identification and Quantification of PAH Photoproducts

PAH and TiO2-NP chemistry solutions were prepared in OECD media at the same concentrations as those used in larval zebrafish exposures but did not contain zebrafish.

Sample Extraction, Preparation, and Chemical Analysis

Solid phase extraction (SPE) of the specific PAH and PAH/TiO2-NP samples was done using Isolute ENV+ cartridges under vacuum and subsequently prepared using previously published methods (42, 43). Extracted PAH and PAH/TiO2-NP samples were prepared using a previously published method (43). Gas chromatography-mass spectrometry (GC-MS) analyses for OPAHs and OHPAHs were performed using an Agilent 7890B GC system coupled to an Agilent 5977A mass spectrometer (MS) detector with electron impact (EI) ionization. OPAH and OHPAHs were analyzed with optimized methods previously published (42, 43). All analyses were run in selected ion monitoring (SIM) mode. GC-MS analyses for all analytes were performed in triplicate using an Agilent DB-5MS (30 m × 0.25 mm I.D. × 0.25 μm film thickness) capillary column and blank corrected. Detailed extraction, preparation (including derivatization), chemical analysis methods, chemicals screened, and information on the internal standards/surrogates used are found in Supplementary Information (Table S3, S6).

Gene Expression Analysis

Zebrafish Husbandry and Exposures

Wildtype WIK strain zebrafish (Danio rerio) embryos were collected from the Heriot-Watt University zebrafish facility, and the same parental lineage was used throughout all exposure studies to avoid familial differences. The WIK adult line was gifted by Professor Charles Tyler at University of Exeter and housed at the Heriot-Watt University zebrafish facility in 19-L tanks filled to 14 L with OECD, on 80 L recirculating systems in a room maintained at 28°C with a 14 h light/10 h dark cycle.

All 72-hour post fertilization (hpf) zebrafish exposures followed OECD guidelines for the testing of chemicals using the Fish Embryo Toxicity (FET) Test. Exposures began at 09:30, in 20 mL glass vials, to ensure all larvae were at the same developmental stage to reduce developmental-specific differences in gene expression profiles due to the rapid larval development of the zebrafish. Each exposure group (n=15), including 0.1% DMSO controls, were collected at 1, 2, 4, and 6 hours post exposure (hpe), in triplicate (n=3); totaling 45 animals per exposure group. All larval collections were placed on ice for 20 min and immediately stored at −80°C.

RNA Isolation, cDNA Synthesis, and qPCR

Total RNA was isolated from pooled larval samples (n=15) using the RNeasy MiniKit using a previously described method and DNase treated to eliminate any genomic DNA contamination of the sample (37). Total RNA was then eluted into 30 μL of sterile RNase/DNase free water, and RNA quantity and quality assessed. All samples with 280/260 ratio between 2.0–2.2 were diluted to a final concentration of 100 ng/μL and cDNA was synthesized from 2 μg total RNA.

Quantitative PCR (qPCR) was performed with cDNA diluted 1 in 10 with nuclease-free water in 20 μL reactions consisting of 300 nM custom designed gene-specific primers ordered from integrated DNA technologies (IDT) and SYBR Green Mastermix. The conditions were as follows: 95°C for 2 min; 40 cycles of 95°C for 15 s, primer-specific annealing temp for 1 min; 95°C for 15 s; 60°C for 1 min; 95°C for 15 s. All samples were analyzed in triplicate with no-template controls included in each run and dissociation analysis to verify primer pair specificity. Primer parameters for the reference gene (β-actin) and target genes (cyp1a, ephx1, ephx2, sod1) can be found in Supporting Information along with the statistical analyses and results of this study (Table S2, S7-S9).

Computational Exploration of Phototransformation

All calculations were carried out using the Gaussian 16 Revision C.02 and Molpro Version 2022.2 program packages (44, 45, 46, 47). Structures were obtained with the B3LYP functional and the 6–31G** basis set in a polarizable continuum model (PCM) parametrized for water to mimic solvent effects (48, 49, 50, 51). Relative free energies, if not denoted otherwise, come from CASPT2/6–31G** calculations in vacuum with an added solvent and free energy correction taken from B3LYP/6–31G** calculations in PCM, in order to properly describe the electronic structure of the structures in question (52, 53).

Results and Discussion

PAH Photoproduct Formation

The identification of 1-hydroxynaphthalene as the primary OHPAH photoproduct of both BaP and FLT combined with the absence of hydroxylated parent-derived molecules, suggests that upon UVA absorption, the naphthalenic rings of FLT and the lower bay region rings of BaP were photodissociated, and subsequently hydroxylated (15, 54, 55, 56). The identification of smaller, physicochemically different OHPAH and OPAH photoproducts across all PAHS investigated, supports that photodissociation upon UVA absorption likely occurred which has only been observed in laboratory astrochemical studies (1, 57). Additionally, 1-hydroxypyrene was the primary OHPAH photoproduct of PYR. No other hydroxylated PYR derivatives were measured in this study, so it is possible that other hydroxypyrene isomers were formed, but unidentified. Previous studies have found the clustered ring arrangement of PYR requires more energetic photons to dissociate, so complete photodissociation of PYR upon UVA absorption was unlikely in the current study (56). Subsequently, most hydroxylation events likely occurred directly on the PYR molecule rather than photodissociated molecules (unlike BaP and FLT), but the identification of physicochemically diverse photoproducts indicates photodissociation and photorecombination events did occur sometime during or following irradiation (Figure 3). Smaller photoproducts of BaP, FLT, and PYR have been identified previously, but the identification of such divergent photoproducts suggest complex PAH-specific photodissociative and photorecombination reactions continued following the end of irradiation.

Figure 3.

Figure 3.

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Concentrations (μg/L) photoproducts identified following UVA irradiation (7.26 mW/cm2) of 80 μg/L PYR and 80 μg/L PYR+ 2 mg/L TiO2-NP samples at 1, 2, 4, and 6 hours post irradiation (hpi). A) The hydroxylated PAH (OHPAH) photoproducts and B) the oxygenated PAH (OPAH) photoproducts. Bars show average ± SE, (n=3) of triplicate runs of the same sample on the instrument.

Consistently, all samples revealed the presence of photoproducts with unique ring arrangements, and/or larger molecular structures, compared to the original PPAHs investigated. No products larger than BaP were targeted for measurement in our investigations, so it is possible that photoproducts larger than 5-ring molecules were present, but unidentified. Evolution of PAHs into larger, more complex molecules has been discussed extensively in astronomical investigations of PAH evolution in the interstellar medium and in soot formation (58, 59, 60, 61). To the author’s knowledge, this is the first instance where photoproducts larger than the initial BaP, FLT, and PYR molecules have been identified following UVA irradiation in the aqueous phase. Current knowledge of PAH photocatalysis focuses on the degradation of PAHs into smaller molecules and subsequent mineralization in the environment (15, 54, 55). Alternatively, if recombination of photodissociated molecules or photoproducts results in the formation of more stable molecules (increased aromatic sextets) then it is possible that recombination into larger products would be physicochemically favored over degradation into smaller, less stable molecules (1). This may also pertain to aromatic ring rearrangements in the PAH molecule, which was observed in all the PAHs investigated in this study. Photoproducts that were either isomers or larger than the original PAH molecules, maintained constant concentrations over the time points investigated (i.e., 4-hydroxychrysene in PYR samples, Figure 3). These results imply that some photodissociated molecules and photoproducts recombine and/or rearrange, likely in order to yield more stable molecules that are subsequently hydroxylated and oxygenated which may persist in an aquatic environment.

PAHs in the presence of TiO2-NPs contained more dihydroxy photoproducts and/or showed opposite time-related concentration patterns of photoproducts (Figures 1–3). No dihydroxy photoproducts larger than dihydroxynaphthalene were targeted in the chemical screening, so these results do not conclude that only dihydroxynaphthalenes were formed. In the presence of TiO2-NPs, more isomers of dihydroxynaphthalene were identified in BaP samples, while PYR samples displayed increases in 1,5-dihydroxynaphthalene and 1-hydroxypyrene concentrations. Extensive studies of TiO2-NP mediated PAH photocatalysis have confirmed that generation of ·OH radicals by TiO2-NPs contributes to increased hydroxylation of PAHs during UVA irradiation (7, 15, 30, 54, 62). In-line with previous studies, we also suggest that the generation of ·OH radicals by TiO2-NPs contributed to increased hydroxylation of the PPAHs investigated as well as their subsequent photoproducts. It is important to note that a new study has also found that PAHs generate ·OH and O2·- radicals which participated in subsequent photochemical reactions, so it is possible this could have occurred within our system (Sarmiento, D., 2023; PAH Transformation). Within our defined time course, the highest 1-hydroxynapthalene and 9,10-PQ concentrations (35–160 μg/L and 270 μg/L respectively) were found in FLT and TiO2-NP solutions, that continued to increase until 6 hpi. At the initial 1-hour timepoint, PAH samples with TiO2-NPs showed higher concentrations of photoproducts which subsequently reduced by 6 hpi. This suggests that TiO2-NPs photocatalytically contributed to PAH degradation, thus enhancing photocatalytic efficiency. Combining these findings with prior studies suggests that TiO2-NP generated OH radicals likely played a role in enhancing hydroxylation efficiency in our system. However, UVA irradiation of the PAHs was sufficient to catalyze comparable photoproduct formation.

Figure 1.

Figure 1.

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Concentrations (μg/L) of benzo[a]pyrene (BaP) photoproducts identified following UVA irradiation (7.26 mW/cm2) of 20 μg/L BaP and 20 μg/L BaP+ 2 mg/L TiO2-NP samples at 1, 2, 4, and 6 hours post irradiation (hpi). A) The hydroxylated PAH (OHPAH) photoproducts and B) the oxygenated PAH (OPAH) photoproducts. Bars show average ± SE, (n=3) of triplicate runs of the same sample on the instrument.

The primary OPAH photoproduct of all three PAHs was 9,10-PQ, regardless of TiO2-NP presence (Figures 1–3). Targeted screening of PAH photoproducts was limited to those with chemical standards commercially available (Table S6). With that, additional photoproducts were likely present in the samples, but were not identified in this study. To the authors’ knowledge, this is the first instance where a single common OPAH photoproduct of BaP, FLT, and PYR, has been identified following UVA irradiation (63, 64). Additionally, 9,10-PQ has been previously identified as the primary photoproduct of anthracene (ANT) and phenanthrene (PHE), low molecular weight (LMW) PAHs (32). Currently there are 16 priority PAHs monitored in various environmental matrices by the U.S. Environmental Protection Agency (EPA), which also includes measurements before and after remediation activities (Titaley et al., 2019, Hussain, et al; PAHs in Environment). The identification of a common photoproduct, like 9,10-PQ, among structurally diverse PAHs could serve as an ideal indicator of PAH pollution and benefit environmental monitoring activities, particularly since 9,10-PQ is bioactive, stable, and environmentally mobile (62, 65, 66, 67). Further exploration of additional PAHs is necessary to substantiate their phototransformation into 9,10-PQ.

Gene Expression Profiling

The presence of TiO2-NPs altered gene expression profiles of all the PAHs investigated (Figure 4). Expression of cyp1a was significantly reduced in the presence of TiO2-NPs in both irradiated and non-irradiated FLT exposures while BaP and PYR exposures demonstrated NP-related induction of cyp1a in a time-dependent manner, where induction increased progressively from 1 to 6 hpe (Figure S3). This observation indicates a reduction in the bioavailability of FLT due to adsorption of FLT to the TiO2-NP surface, demonstrated in previous studies, while BaP and PYR may not interact with the NP surface as strongly thus remaining more bioavailable (68). In line with the current study, previous investigations have also shown that individual PAHs possess distinct adsorptive and photocatalytic capabilities that directly alter bioavailability which may be reflected in gene expression analyses (15, 32, 36, 67). The significantly lower induction of cyp1a identified in irradiated FLT exposures compared to non-irradiated exposures, both in the presence and absence of TiO2-NPs, could indicate photodegradation of FLT was favored over biodegradation as cyp1a is responsible for encoding enzymes that metabolize PPAH molecules. Additionally, the induction of ephx2 at 4pe in irradiated PYR and TiO2-NP exposures coincides with the highest induction of cyp1a, which could have been caused by the presence of bioavailable photodissociated PPAHs and PYR photoproducts (Figure 3, 4, S3). These data suggest that the presence of TiO2-NPs catalyze production of bioavailable BaP and PYR photoproducts, but strong FLT-NP interactions likely sequestered the molecule and its subsequent photoproducts.

Figure 4.

Figure 4.

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Average (± SE, n=3) log fold change of cyp1a, ephx1, ephx2, and sod1 in zebrafish larvae following exposure to UVA irradiated (7.26 mW/cm2) and non-irradiated 20 μg/L BaP and 20 μg/L BaP+ 2 mg/L TiO2-NP; 70 μg/L FLT, 70 μg/L FLT+ 2 mg/L TiO2-NP; 80 μg/L PYR and 80 μg/L PYR+ 2 mg/L TiO2-NP at 1, 2, 4, and 6 hours post exposure (hpe). Average log fold changes were calculated using the ΔΔCT method and normalized to β-actin at each time point. Statistically significant differences were identified by two-way ANOVA, TukeyHSD, p-values can be found in SI Tables 7-9.

Irradiation of BaP and FLT exposures diminished expression of genes that were highly expressed in irradiated exposures, whereas irradiated PYR exposures displayed higher inductions of cyp1a, ephx1, and sod1 compared to non-irradiated exposures at 4 hpe. The PAH-specific, irradiation-related differences observed are likely attributed to the differing bioavailability of the PPAHs and their photoproducts. It is likely that PYR photoproducts were more readily bioavailable due to their persistence as well as the higher number of total photoproducts formed compared to those of BaP and FLT (Figure 3). The significant reduction of cyp1a expression in irradiated BaP and FLT exposures (in both the absence and presence of TiO2-NPs) combined with previous knowledge that cyp1a specifically encodes enzymes responsible for metabolizing PPAHs indicated that phototransformation of the PPAHs must have occurred. Compared to an earlier study investigating anthracene (ANT) and phenanthrene (PHE) expression profiles of the same genes targeted in this study, the HMW PAH exposures within this study induced greater responses compared to ANT or PHE, which supports previous studies suggesting HMW PAHs are more readily biodegraded than the LMW PAHs (32, 55). Overall, these data demonstrate the susceptibility of all three PAHs to biodegradation in the absence of UVA irradiation as well as reduced bioavailability elicited by TiO2-NP presence, particularly in FLT exposures.

Proposed Pathways for Phototransformation

Photolysis of ANT and subsequent recombination to produce PHE was calculated in order to address initial findings of ANT phototransformation to 9,10-PQ as seen in St. Mary et al., 2021(Figure 5). The proposed pathway shows that ANT could dissociate photolytically into two triradical fragments as the UVA irradiation used in the experiment is of sufficiently high energy to facilitate this. Two of the triradicals can then recombine to form a PHE molecule. This recombination is likely preferred as PHE is ~4 kcal/mol more resonance-stabilized than ANT. This has been shown previously and confirmed by our calculations (Table S4, 69). The advantage of this possible pathway is that the triradical fragments could be formed out of different PAHs, not only ANT, and as the recombination to PHE is thermodynamically favored, this further supports our calculated transformation of PHE to 9,10-PQ as a common photoproduct after singlet oxygen addition (70). The presence of singlet oxygen within the system can be justified by PAHs acting as wellknown photosensitizers (71, 72). The detailed addition pathway is shown in Figure S5.

Figure 5.

Figure 5.

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Proposed phototransformation process of anthracene to phenanthrene in the absence of TiO2-NPs and subsequent addition of singlet oxygen to form 9,10-PQ.

It is important to note that given the complexity involved, the proposed mechanisms are but one of many possible pathways envisaged, albeit one that follows known photochemistry in related systems. Alternative considerations including carbon capture, photolytic fragmentation and mechanisms involving excited states of the PAHs are further possibilities. We believe that our computations give an initial insight into the possible photocatalytic processes that take place during PAH phototransformation and can lead to further study of other competing pathways.

Conclusions

We have found numerous, diverse photoproducts of PAHs, which undermines the depiction of the chemical profiles monitored in the environment currently. Furthermore, many of the identified photoproducts exhibit high toxicity (21, 22, 24). The occurrence of PAH photodissociation and photorecombination events was evidenced by the identification of photoproducts that were smaller, larger, and physicochemically distinct from the initial PPAH molecules. In our system, hydroxylation and oxygenation reactions continued through 1, 2, 4, and 6 hpi. The identification of higher photoproduct concentrations than the original PAH concentrations has not been determined, but it is possible this was caused by interactions between the air-water interface, interactions of PAHs with dissolved oxygen within the E3 media, or a combination of both (Table S5, 73). Further investigations are required to explore photoproduct concentrations beyond our specified time range and determine the extent of ongoing PAH phototransformation following irradiation.

The presence of TiO2-NPs changed the bioavailability of PAHs in a manner specific to each PAH, but the bioavailability of subsequent photoproducts did not consistently mirror that of the PPAH. These results illustrate the complex interplay of PAH-specific adsorption and bioavailability in the aqueous phase. Furthermore, it is important to add that agglomeration and transport of TiO2-NPs in the aquatic environment is likely influenced by the presence of sediment, and other particulate matter (74, 75). This warrants additional research to quantify TiO2-NP adsorption affinity, aggregation, and interactions with other particulates to gain a clearer understanding of the processes discussed within this manuscript and their relevance to aquatic environmental dynamics.

The consistent generation of 9,10-PQ across diverse PAHs could make it a valuable indicator for assessing 1) PAH contamination, 2) the progress of remediation processes, and 3) the safety of polluted sites (32). The proposed reaction mechanisms regarding phototransformation into the primary photoproduct, 9,10-PQ, can be used to aide in the elucidation of how other photoproducts identified in this study were formed which warrants further investigations. It’s crucial to investigate additional PAHs to confirm 9,10-PQ as a common photoproduct, paying particular attention to PAHs frequently found in the environment. We have found that solely monitoring PAH molecules may not accurately reflect the safety or extent of potential PAH contamination or degradation, in the aqueous phase. Overall, these findings emphasize the importance of adopting comprehensive analytical approaches that capture the full spectrum of photoproducts formed during PAH catalysis. This is particularly important during environmental remediation processes to ensure aquatic environments deemed as safe, are assuredly safe for the public as well as wildlife. The data provided by this study is meant to serve as a foundation for future, much needed investigations, to achieve a more comprehensive understanding of PAH photoproduct formation in the aquatic environment.

Supplementary Material

MMC6
MMC5
MMC4
MMC2
MMC1
MMC3

Figure 2.

Figure 2.

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Concentrations (μg/L) of fluoranthene (FLT) photoproducts identified following UVA irradiation (7.26 mW/cm2) of 70 μg/L FLT and 70 μg/L FLT+ 2 mg/L TiO2-NP samples at 1, 2, 4, and 6 hours post irradiation (hpi). A) The hydroxylated PAH (OHPAH) photoproducts and B) the oxygenated PAH (OPAH) photoproducts. Bars show average ± SE, (n=3) of triplicate runs of the same sample on the instrument.

Highlights.

  • Diverse photoproducts of unique physiochemistry are formed from individual PAHs

  • Gene expression profiling and analytical chemistry showed PAH-specific bioavailability

  • 9,10-phenanthraquinone was the primary photoproduct of disparate PAHs

  • Photorecombination to form 9,10-phenanthraquinone is thermodynamically favored

Acknowledgements

Special thanks to Dr. Charles Tyler for gifting zebrafish from his facility at the University of Exeter. This work was supported by the European Union (EU) and Horizon 2020 awarded under the Marie-Sklodowska-Curie action to the EUROPAH consortium, grant number 722346. This publication was made possible in part by Grant Numbers AGS-1411214 from the National Science Foundation (NSF), and P42-ES016465 and P30-ES00210, from National Institute of Environmental Health Sciences (NIEHS), National Institute of Health (NIH). Its contents are the sole responsibility of the authors and do not represent the official view of the NIEHS or NIH. Martin Paterson gratefully acknowledges funding from the Engineering and Physical Sciences Research Council UK (EPSRC) through grants EP/V006746 and EP/T021675, and support from the Leverhulme Trust (Grant No. RPG-2020-208). Luca Craciunescu acknowledges a James-Watt scholarship from Heriot-Watt University. Graphical abstract created with Biorender.com.

Footnotes

Supplementary Information

Details on experimental layout, materials, methods; solution preparations; concentrations of anthracene and phenanthrene found in water sources; RNA isolation, cDNA synthesis, and qPCR reagents used; SPE cartridges and solvents; GC-MS system, sample extractions, sample preparation, and derivatizing agents for OHPAH screening; primer parameters for primers used in qPCR analyses and primer optimization; zebrafish husbandry and exposures; summary of statistical and data analyses of gene expression; data analyses of GC-MS; details of computational methods used; heat maps of p-values calculated for gene expression analyses; heat map comparing concentration differences of photoproducts in the presence of TiO2-NPs.

Declaration of interests

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

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Associated Data

This section collects any data citations, data availability statements, or supplementary materials included in this article.

Supplementary Materials

MMC6
NIHMS2000745-supplement-MMC6.xlsx (141.1KB, xlsx)
MMC5
NIHMS2000745-supplement-MMC5.xlsx (138.2KB, xlsx)
MMC4
NIHMS2000745-supplement-MMC4.xlsx (141.1KB, xlsx)
MMC2
NIHMS2000745-supplement-MMC2.xlsx (38.7KB, xlsx)
MMC1
NIHMS2000745-supplement-MMC1.docx (1.5MB, docx)
MMC3
NIHMS2000745-supplement-MMC3.xlsx (12.2KB, xlsx)

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