Spatiotemporal variations of halogenated polycyclic aromatic
hydrocarbons in sediments of Pearl River Estuary: Occurrence, sources, and potential ecological risks
- Introduction
Halogenated polycyclic aromatic hydrocarbons (HPAHs), as a group of the derivatives of polycyclic aromatic hydrocarbons (PAHs) (Sun et al., 2013), are ubiquitous in various environmental matrices, e.g., air, water, fly ash, soil, sediment, and living organisms (Ohura et al., 2008; Wang et al., 2016a; Nishimura et al., 2017; Horii et al., 2009). HPAHs are considered the by-products of anthropogenic activities, and their main pollution sources include incomplete combustion of fossil fuel (Jin et al., 2017), burning of e-waste (Ma et al., 2009), automobile emissions (Gao et al., 2018), and industrial emissions (Jin et al., 2018). Moreover, photoreactions and thermal reactions may be the potential ways for
HPAHs to in situ produce in the environment (Ohura et al., 2013). Due to high lipophilicity, HPAHs can be readily accumulated in organisms and amplified through the food chain (Myers et al., 2014; Wickrama- Arachchige et al., 2020). HPAHs are recognized as carcinogens, terato- gens and mutagens (Fu et al., 1999). In terms of aryl hydrocarbon re- ceptor (AhR)-mediated signaling pathway, HPAHs are even more toxic than parent counterparts (Fu et al., 1999; Ohura et al., 2007), and AhR activities are closely dependent on the halogen atom number of HPAHs (Huang et al., 2018). Therefore, the potential ecological risks of HPAHs in the environment have attracted great attention.
Estuaries that are located at the transitional zone of river and marine play vital roles in the transport of contaminants (Abell et al., 2010; Yan
- Corresponding author at: Guangdong Provincial Key Laboratory of Marine Resources and Coastal Engineering, School of Marine Sciences, Sun Yat-sen University, Zhuhai, 519082, China.
E-mail address: chenbw5@mail.sysu.edu.cn (B. Chen).
1 These authors contributed equally to this work.
https://doi.org/10.1016/j.marpolbul.2025.118722
Received 21 May 2025; Received in revised form 12 August 2025; Accepted 12 September 2025
Available online 17 September 2025
0025-326X/© 2025 Elsevier Ltd. All rights are reserved, including those for text and data mining, AI training, and similar technologies.
et al., 2019). PAHs and substituted PAHs can enter the estuaries and coastal areas through riverine input and atmospheric deposition (Yuan et al., 2015; Guo et al., 2014), and subsequently are redistributed among different environmental compartments (Yang et al., 2008). PAHs tend to adsorb and accumulate in the sediments due to the high hydrophobicity, and resuspension process can also lead to the release of PAHs from the sediments to overlying water (Yang et al., 2008; Bery et al., 2024). Therefore, examining PAHs and substituted PAHs in the sediments is important for assessing their aquatic ecological risks in the estuary. Nevertheless, the information regarding HPAHs in the sediments is fairly limited (Xie et al., 2021), which might largely impair our comprehensive estimation on the pollution status and threats of HPAHs.
The Pearl River Estuary (PRE) is a typical tide-dominated estuary that connects the Pearl River to the South China Sea (Niu et al., 2021). With rapid urbanization and industrialization of the Pearl River Delta (Komprda et al., 2009), an enormous quantity of contaminants are dis- charged into the PRE in recent decades (Yuan et al., 2015; Yuan et al., 2017; Chen et al., 2013). Previous studies showed that the levels of HPAHs in the PRE sediments were about 1–2 orders of magnitude higher than those in other estuaries or coastal regions, e.g., Tokyo Bay, Saginaw River watershed, and Bui Dau (Yuan et al., 2020). However, spatio- temporal changes in the levels and sources of HPAHs in the PRE have not been well understood, which could be closely associated with anthro- pogenic inputs and hydrological parameters (e.g., suspended particulate matters) (Yuan et al., 2020; Liu et al., 2017). Knowledge regarding spatiotemporal distribution HPAHs in the PRE could be essential to ac- curate assessment of their potential ecological risks, as well as instructing the pollution controls implemented by the local governments.
In this study, a total of 30 surface sediment samples were collected at 15 sampling sites in the PRE during the wet and dry seasons, respectively (two samples per site, one per season), and the spatiotemporal distri- bution patterns of 17 target HPAHs were characterized. Next, the posi- tive matrix factorization (PMF) model was used to identify potential anthropogenic sources of HPAHs and evaluate their relative contribu- tions. At last, the ecological risks of HPAHs in the PRE sediments were estimated using the risk quotient (RQ) and toxicity equivalent quotient (TEQ), which were widely applied in the ecological risk assessment of PAHs (Tsai et al., 2009; Zhao et al., 2021).
- Experimental section
2.1. Chemicals
In this study, 17 HPAHs and their corresponding parent PAHs were analyzed. The 9-bromofluorene (9-Br-Flu), 2,7-dibromofluorene (2,7- DiBr-Flu), 1-bromopyrene (1-Br-Pyr), 9,10-dichloroanthracene (9,10- DiCl-Ant), 9,10-dibromoanthracene (9,10-DiBr-Ant), 7-bromobenzo[a] anthracene (7-Br-BaA) and 9-bromophenanthrene (9-Br-Phe) and in- ternal standards phenanthrene-d10 (Phe-d10), anthracene-d10 (Ant-d10), fluoranthene-d10(Flt-d10), pyrene-d10(Pyr-d10) and benzo[a]pyrene-d12 (Bap-d12) were purchased from CNW (CNW Technologies GmbH, Ger- many). The 2-bromoanthracene (2-Br-Ant), 9-bromoanthracene (9-Br- Ant), and 2-chloroanthracene (2-Cl-Ant) were purchased from TCI (Tokyo Chemical Industry Co. Ltd., Japan). The 9-chloranthracene (9-Cl- Ant) was purchased from ALFA (Alfa Aesar, UK), and 2,7-dichlorofluor- ene (2,7-DiCl-Flu) was purchased from Adamas (Adamas-beta, China). The 3-chlorofluoranthene (3-Cl-Flt), 1-chloropyrene (1-Cl-Pyr), and 6- chlorobenzo[a]pyrene (6-Cl-Bap) were purchased from Chiron (Chiron AS, Norway). The 9-chlorophenanthrene (9-Cl-Phe) and 2-bromofluor- ene (2-Br-Flu), and 7 parent PAHs (anthracene (Ant), fluorene (Flu), phenanthrene (Phe), pyrene (Pyr), fluoranthene (Flt), benzo[a]anthra- cene (BaA) and benzo[a]pyrene (Bap)), were purchased from Aldrich (Sigma-Aldrich, USA).
2.2. Sample collection
In this study, sediment samples were collected from 15 sampling sites in the PRE during the wet season (September) and dry season (December) in 2019, as shown in Fig. 1. All sampling sites are located at four runoff outlets of the PRE at the east side (i.e., Humen, Jiaomen, Hongqilimen, and Hengmen), which had significantly higher PAH contamination levels compared to those at the west side (i.e., Modao- men, Jitimen, Hutiaomen, and Yamen) (Yuan et al., 2015). Sampling sites were chosen according to the international standard ISO 5667-12 (ISO 5667-12 (2017-07), n.d.). The criteria for selecting sampling sites included stable sedimentation conditions, ease of repeated access to the locations and seasonal accessibility, major runoff outlets with high in- puts of contaminants, reflecting distinct gradients in anthropogenic impacts, and adjacent to historical contamination hotspots. The water depth at sampling sites in the PRE ranged from 3.8 to 13.0 m. Surface sediment samples (0–5 cm) were collected using a grab collector. All samples were transported to the laboratory on ice and in the dark within 24 h. Samples were stored at − 80 ◦ C before subsequent pretreatment and instrumental analysis.
2.3. Sample preparation
Stones, gravel and dead branches and leaves were removed from the sediment samples, and then dewatered in a freeze dryer. The dried sediments were ground and passed through an 80-mesh sieve and af- terwards mixed evenly. HPAHs and PAHs were extracted from the sed- iments using an accelerated solvent extractor. The general process was to weigh ~10 g of sediments and add them to the extraction cells. Deuterated internal standards (Phe-d10, Ant-d10, Flt-d10, Pyr-d10 and Bap-d12) were added to achieve a final concentration of 0.1 μg/mL. The extraction conditions were set according to the previous publication (Yuan et al., 2020). In brief, the sediment samples were extracted for 10 min at 120 ◦ C with the extraction solvent, i.e., a mixture of dichloro- methane and acetone (1:1, v/v). The filtered extractant collected from the heated extraction vessel was then concentrated using a rotary evaporator. Finally, the concentrated extractant was adjusted to a final volume of 1.0 mL with acetone.
2.4. Instrumental analysis
The qualitative and quantitative analysis of PAHs and HPAHs was performed using gas chromatography (GC, Agilent 7890, USA) coupled with a mass spectrometry detector (MSD, Agilent 5975, USA). The GC was equipped with a DB-5MS column (60 m × 0.25 mm × 0.25 μm, Agilent, USA). Helium was used as a carrier gas with a flow rate of 1.0 mL/min. The GC temperature program was as follows: the injection temperature (310 ◦ C); the oven was held at 60 ◦ C for 5 min; the tem- perature was then ramped to 110 ◦ C at the rate of 40 ◦ C/min, and then was ramped to 310 ◦ C at 10 ◦ C/min and finally held isothermally at 310 ◦ C for 8 min. This MSD was equipped with a quadrupole mass spectrometer operating in electron impact mode at 70 eV and 230 ◦ C (ion source temperature).
2.5. Quality assurance and control
Qualitative analysis of HPAHs was performed by comparing the retention times of authentic standards with those in the sediment sam- ples, and the concentrations of HPAHs were quantified using the internal standard method. The standard curves were generated based on GC–MS analysis of a series of standard solutions with the known concentrations. Blank samples spiked with internal standards were in parallel processed with each batch of samples to monitor the instrument stability and avoid potential procedural contamination. Each of sediment samples was analyzed in triplicate. Moreover, the recoveries of HPAHs from the whole pretreatment and analytical procedure were evaluated by spiking3 known concentrations of HPAH standards into the sediments. The recoveries for 7 parent PAHs and 17 HPAHs in the PRE sediments ranged from 73.6 % to 136.9 % and from 77.2 % to 135.3 %, respectively. The details of recovery, limits of detection (LODs) and limits of quantitation (LOQs) are shown in Tables S1 and S2 in the Supplementary Information (SI).
