Pollution Characteristics and Risk Assessment of Chlorinated Paraffins in Seawater and Kelp from Kelp Mariculture Areas of the Shandong Peninsula

Abstract

Chlorinated paraffins (CPs) are persistent, bioaccumulative, and toxic. In marine environments, most studies have focused on short-chain CPs (SCCPs) in animals, while medium-and long-chain CPs (MCCPs and LCCPs) in plants have been neglected. In this study, samples collected from kelp mariculture zones in different seasons were analyzed for the CPs’ contamination characteristics and spatiotemporal distributions in seawater and contamination profiles, bioaccumulation behavior, and dietary exposure risks in kelp. In seawater, the total concentration ranges of SCCPs, MCCPs, and LCCPs were 25.44–245.75, 8.24–27.19, and not detected at 3.26 ng/L, respectively. Spatially, the CP concentrations were influenced by industrial discharge, riverine inputs, and dilution effects, and were significantly higher in nearshore water than in offshore areas (p < 0.05). The concentrations were significantly higher in February than in May, which was attributed to emissions from winter heating and reduced vessel activity during a fishing moratorium. In kelp, the total concentration ranges of SCCPs, MCCPs, and LCCPs were 5.4–210.9, 0.007–0.87, and 0.0–4.45 ng/g wet weight, respectively. Kelp exhibited significant growth-stage-dependent bioaccumulation of CPs, with higher CP concentrations and bioaccumulation factors in its tender stage (February) than during its mature stage (May). Congener analysis revealed similar composition patterns between seawater and kelp. According to a dietary risk assessment (hazard quotient < 0.01), the potential health risks associated with kelp consumption are low.

1. Introduction

Chlorinated paraffins (CPs) are synthetic mixtures of polychlorinated n-alkanes and comprise thousands of isomers, enantiomers, diastereomers, and homologs [1,2]. According to their carbon chain length, CPs are classified into short-chain chlorinated paraffins (SCCPs, C_10–13), medium-chain chlorinated paraffins (MCCPs, C_14–17), and long-chain chlorinated paraffins (LCCPs, C_18–30). Because of their excellent chemical and thermal stability, CPs are widely used as plasticizers, lubricant additives, coolants, flame retardants, metal working fluids, and sealants [2]. During the production, storage, transportation, use, and disposal of these products, CPs are continuously released into the environment. Furthermore, CPs may enter the environment through pathways such as waste incineration [2], wastewater treatment [3] and electronic waste dismantling. To date, CPs have been widely detected in air [4], water [5], soil [6], sediments [7], aquatic and terrestrial wildlife [8,9], and human breast milk [10], and are recognized for their long-range transport potential [11], persistence [12], bioaccumulation [13,14], and toxicity [15]. SCCPs were listed in Annex A of the Stockholm Convention on Persistent Organic Pollutants in 2017 [16]. MCCPs were reviewed for inclusion in Annex A of the Stockholm Convention in 2025 and will be subject to further deliberation at a future meeting of the Persistent Organic Pollutants Review Committee [17].

The environmental and health risks associated with CPs are of particular concern in coastal marine ecosystems. Some studies have indicated that the coastal waters of the Yellow Sea in Shandong, China, are contaminated with SCCPs and MCCPs [18,19]. Notably, this region serves as a major kelp production area in China. Data from 2018 indicate that Shandong Province contributed to 35.7% of the national kelp yield and accounted for 40.9% of the cultivation area, which ranked it as the second-largest kelp production region in China [20]. High concentrations of CPs in coastal waters may pose a threat to the healthy cultivation of kelp and compromise the safety of kelp for human consumption. However, systematic research on the pollution status of CPs in kelp mariculture areas is scarce. In particular, there is a lack of contamination data covering different aquaculture seasons and encompassing all SCCP, MCCP, and LCCP homologs. Given that current knowledge is predominantly derived from studies on marine animals [21,22], systematically conducting comprehensive monitoring and risk assessment of CPs, including SCCPs, MCCPs, and LCCPs, is important for ensuring sustainable development of the kelp mariculture industry and safeguarding human health.

In marine fauna, CPs exhibit potential for bioaccumulation and biomagnification [23,24,25] because of their high octanol–water partition coefficients and resistance to metabolic degradation [26,27]. The bioaccumulation factor (BAF) of a compound in a specific species is influenced by the species size, activity level, and lipid content distribution [28]. For instance, pelagic fish in Liaodong Bay exhibit higher BAFs (2.34) than demersal fish (2.08) [29]. As a macroalga, kelp spans various water layers and may exhibit specific CP bioaccumulation patterns across different growth seasons and algal tissues. Data on the bioaccumulation of CPs, particularly MCCPs and LCCPs, in kelp remain limited. Furthermore, the current understanding of bioaccumulation is predominantly derived from studies on marine animals [30], and the mechanisms and extent of CP accumulation in marine plants remain poorly understood. Because human activities are more intensive in kelp mariculture areas than in non-cultivation zones, the occurrence characteristics of CPs in kelp mariculture areas will be complex, and these compounds potentially pose a threat to the safety of kelp for human consumption [31]. Therefore, comprehensive research is required to further elucidate the contamination levels and bioaccumulation patterns of SCCPs, MCCPs, and LCCPs in marine plants.

In this study, sampling sites were established in both nearshore and offshore areas within a key kelp mariculture region of the Yellow Sea. CP contamination was investigated in kelp and seawater samples collected in 2 months in different seasons. The aims of this study were to (1) elucidate the contamination levels and spatiotemporal distribution characteristics of CPs in seawater from mariculture areas; (2) investigate the contamination levels, bioaccumulation patterns, and potential impacts of CPs in kelp across different cultivation periods; and (3) assess the dietary risks associated with kelp consumption.

2. Materials and Methods

2.1. Instruments and Reagents

The research used an automatic solid-phase extraction (SPE) instrument (Fotector-08HT, RayKol Group, Xiamen, China); a nitrogen evaporator (N-EVAPTM 112, Organomation Associates, West Berlin, MA, USA); a Milli-Q Gradient ultrapure water purification system (Millipore, France); and an ultrasonic cleaner (KQ-600E, Kunshan Ultrasonic Instruments Co., Ltd, Kunshan, Jiangsu, China). Standard solutions of SCCPs, MCCPs, and LCCPs were purchased from Ehrenstorfer GmbH (Augsburg, Germany). The SCCP mixtures (C_10–13) had chlorine contents of 51.5%, 55.5% (98% purity), and 63.0%; the MCCP mixtures (C_14–17) had chlorine contents of 42%, 52%, and 57% (98% purity); and the LCCP mixtures (C_18–30) had chlorine contents of 36% and 49% (98% purity). All standards had a concentration of 1.00 ng/μL. Pesticide residue-grade dichloromethane and n-hexane were purchased from J.T. Baker (Merck KGaA, Germany). Pesticide residue-grade methanol and acetonitrile were obtained from CNW (Düsseldorf, Germany).

2.2. Study Area and Sample Collection

To comprehensively investigate the contamination status of CPs in kelp mariculture areas along the Yellow Sea coast of Shandong, a total of 10 sampling sites were established (Figure 1). These sites were selected to collect both nearshore and offshore seawater samples from each region. The samples were collected near Jiming Island (sites S1 and S2), Xincheng Mountain (sites S3 and S4), Xunshan (sites S5 and S6), Moye Island (sites S7 and S8), and Quanhai (sites S9 and S10). Seawater samples were collected in February and May 2024 using a 5 L water sampler (GU12−NISKIN, HYDRO−BIOS, Altenholz, Germany) at approximately 50 cm below the sea surface. At each sampling site, 3 L of water was collected and transferred into 1 L amber glass bottles, which were subsequently stored at −20 °C until required for analysis.

Fresh kelp samples were collected from sites S1, S3, S5, S7, and S9 in February and May 2024 (Figure 1). The collected kelp samples were rinsed with seawater. The holdfasts were then separated from the fronds, wrapped in aluminum foil, and frozen at −20 °C until required for laboratory analysis.

2.3. Sample Extraction and Cleanup

The pretreatment and extraction of CPs from seawater were adapted from previously reported methods [32,33]. Briefly, seawater samples were first filtered through glass fiber filters (0.45 μm pore size; Shanghai Xingya, China). Next, an automated SPE (GX-274 ASPEC, Gilson Company, Middleton, WI, USA) system was used to pass 500 mL (n = 3) aliquots of the filtrate through hydrophilic–lipophilic balance (HLB) cartridges (polystyrene–divinylbenzene, 200 mg, 6 mL; Waters Corporation, Milford, MA, USA) to enrich the CPs. Before extraction, the HLB cartridges were conditioned sequentially with 5 mL of methanol and then 5 mL of ultrapure water.

The HLB cartridges were eluted with 6 mL of liquid chromatography-grade acetonitrile (CNW Technologies, Düsseldorf, Germany), and the eluents were collected in 15 mL centrifuge tubes (Nunc, Thermo Fisher Scientific, Waltham, MA, USA). The extracts were evaporated to near dryness under a gentle stream of nitrogen gas at 40 °C. The residues were reconstituted in acetonitrile to a final volume of 1 mL, filtered through a 0.22 μm polyvinylidene fluoride membrane (Thermo Fisher Scientific), and prepared for instrumental analysis.

The root and tip samples of the kelp were processed separately. Each sample was cut into small pieces and homogenized using a high-speed grinder. Next, 2 g of the homogenized sample was mixed and ground with 6 g of anhydrous sodium sulfate to remove moisture. The sample was then extracted into dichloromethane/hexane (1:1, v/v) by ultrasonication for 30 min. After centrifugation, the supernatant was transferred to a 15 mL centrifuge (Virya, Shanghai, China) tube. For each sample, the extraction was repeated, and the supernatants were combined. The extract was then purified by passing it through a Florisil column (Agilent Technologies, Santa Clara, CA, USA). The collected eluate was evaporated to near-dryness under a gentle stream of nitrogen gas at 40 °C and reconstituted in acetonitrile to a final volume of 1 mL. Primary-Secondary Amine was added, and the mixture was vortexed and then ultrasonicated for 5 min. After centrifugation, the supernatant was filtered through a 0.22 μm polyvinylidene fluoride membrane for instrumental analysis.

2.4. Instrumental Analysis

Instrumental analysis was performed according to previously reported methods [32,33]. An ultra-high performance liquid chromatography–quadrupole–Orbitrap mass spectrometer (UHPLC-Q-Orbitrap MS, Thermo Fisher Scientific) was used to analyze the CP concentrations in the pretreated samples. Chromatographic separation of the CPs was performed using a Waters HSS PFP column (2.1 × 100 mm, 1.8 μm). The injection volume was 5 μL. The mobile phase was a mixture of ultrapure water and acetonitrile, and the mobile phase flow rate was 0.4 mL/min. To improve the mass spectrometry response, tetramethylammonium chloride was infused via a peristaltic pump through a T-union and mixed with the CP before the sample entered the ion source. A heated electrospray ionization source was used with the following parameters: spray voltage, 2.5 kV; capillary temperature, 320 °C; sheath gas flow rate, 35 L/min; and auxiliary gas flow rate, 10 L/min. Detailed acquisition parameters for the CPs are provided in Table S3.

2.5. Quality Assurance and Control

During sampling, triplicate samples were collected for each target. To avoid contamination, all glassware was rinsed sequentially with ultrapure water and methanol before use. Consumables such as filter membranes, aluminum foil, and scalpels were pre-combusted at 550 °C for 5 h. Additionally, all solutions were filtered through glass fiber membranes before use. During the instrumental analysis, triplicate blank controls were included to monitor potential exogenous CP contamination. The analytical results indicated that the CP concentrations in the blank controls were either close to or below the limit of detection. For each batch of samples, procedural blanks and spiked samples were analyzed to determine their recovery rates, which were then used to correct the recovery rates in the sample data. The recovery range for seawater was 55.1–60.5% and that for kelp was 47.2–75.6%.

2.6. Statistical Analysis

In this study, the results are expressed as the mean ± standard deviation. The Kolmogorov–Smirnov test and Levene’s test were used to assess the normality and homogeneity of the variance of the data. Following verification of these assumptions, one-way analysis of variance was performed to evaluate differences among environmental media or organisms in different regions. In this study, the significance level for differences and correlations was set at p < 0.05. Data analysis and plotting were performed using SPSS 26.0 (IBM Corporation, Armonk, NY, USA), ArcGIS 10.6 (Environmental Systems Research Institute, Inc., Redlands, CA, USA), and Origin 2024 (Origin Lab Corporation, Northampton, MA, USA).

2.7. Calculation of BAFs

The BAF is defined as the ratio of the CP concentration in an organism (Cbiota, ng/g wet weight [w.w.]) to the average concentration in the water (Cwater, ng/L), as shown in Equation (1). The calculated values were used to characterize the bioaccumulation potential of SCCPs.

$$\mathrm{B}\mathrm{A}\mathrm{F}={\displaystyle \frac{{\mathrm{C}}_{\mathrm{b}\mathrm{i}\mathrm{o}\mathrm{t}\mathrm{a}}}{{\mathrm{C}}_{\mathrm{w}\mathrm{a}\mathrm{t}\mathrm{e}\mathrm{r}}}}\times 1000$$

(1)

2.8. Dietary Risk

Dietary intake is a major pathway for human exposure to CPs. The estimated daily intakes [EDIs, ng/(kg·day)] of SCCPs, MCCPs, and LCCPs from kelp consumption for local residents were calculated using Equation (2).

$$\mathbf{E}\mathbf{D}\mathbf{I}={\displaystyle \frac{(\mathit{c}\times \mathit{M})}{\mathbf{B}\mathbf{W}}}$$

(2)

where c is the wet weight concentration of SCCPs, MCCPs, or LCCPs in edible kelp (ng/g w.w.) and M is the daily kelp consumption rate. Because of a lack of specific data on kelp intake, the per capita average aquatic product (laver, kelp, and marine fish) consumption rate for urban residents (7.7 g/d) [34] was used for the daily kelp intake. The body weight (BW) is the average body weight for local residents (kg), which was set at 69.6 kg for adult males and 59.0 kg for adult females [34].

The hazard quotient (HQ) was calculated from the EDI and tolerable daily intake (TDI) using Equation (3) and the result was used to assess comprehensive health risks.

$$\mathrm{H}\mathrm{Q}={\displaystyle \frac{EDI}{\mathrm{T}\mathrm{D}\mathrm{I}}}$$

(3)

The HQ is used to assess the integrated health risk. An HQ value of >1 indicates that the current pollutant concentration is potentially risky. The UK Committee on Toxicity proposed a TDI of 30 μg/(kg·d) for SCCPs and 4 μg/(kg·d) for MCCPs [35]. Given that a TDI for LCCPs has not yet been established, and considering that SCCPs exhibit higher toxicity than LCCPs [36], the TDI for SCCPs was used as a surrogate for that of LCCPs.

3. Results and Discussion

3.1. Pollution Characteristics of CPs in Seawater from Kelp Mariculture Areas

The total concentrations of the SCCPs, MCCPs, and LCCPs in seawater were 25.44–245.75, 8.24–27.19, and not detected at 3.26 ng/L, respectively (Figure 2A and Figure 3A, Table S1). The mean concentrations of the SCCPs, MCCPs, and LCCPs in seawater were 108.2 ± 62.7, 14.6 ± 5.8, and 0.40 ± 0.74 ng/L, respectively. SCCPs and MCCPs were detected in all of the seawater samples. The total SCCP concentrations were higher than those reported in seawater from Liaodong Bay, China, in 2011 (7.7 ng/L) [37]. However, they were lower than the concentrations reported in samples from Laizhou Bay in 2022 (362.23 ng/L) [29], the Huangpu River in Shanghai in 2016 (448 ng/L) [38], and the Weihai coastal area in 2017 (525.2 ng/L) [18]. The MCCPs concentrations in the present study were higher than those detected in seawater from Liaodong Bay, China (8.7 ng/L) [39], and in Lake Ontario in 2004 (0.9 pg/L) [40]. Research on LCCP concentrations in seawater is currently limited; however, the average total LCCP concentration in seawater from the Haima cold seep in the South China Sea was 0.3 ng/L [41], which is comparable to our findings. These comparisons show that the average concentrations of SCCPs and MCCPs in seawater in kelp mariculture areas are of the same order of magnitude as those reported in other regions.

The total concentrations of SCCPs in the seawater samples collected in February were generally higher than those in the seawater samples collected in May, and the differences were statistically significant (p < 0.05). This seasonal variation trend may be associated with an annual fishing moratorium that is in place in the coastal waters of Shandong from May to September [42]. During the open fishing season, many fishing vessels operate at sea, which contribute to SCCPs through antifouling coatings on their hulls and fuel use [43,44]. Previous studies have confirmed that vessels and their exhaust emissions are major sources of SCCPs in the marine environment [45]. Additionally, commencement of the central heating season in February in the Shandong region [46] will contribute to increased emissions of pollutants, and previous studies have demonstrated that CPs can be deposited into seawater via air–water exchange [47,48].

3.2. Pollution Characteristics of Kelp

SCCPs, MCCPs, and LCCPs were detected in all of the kelp samples (Figure 3B). The total concentration range of SCCPs in kelp was 5.93–210.86 ng/g w.w., with a mean value of 82.67 ± 60.90 ng/g w.w. (Figure 2B). The average concentrations of SCCPs were 117.87 ± 67.56 ng/g w.w. in February and 47.48 ± 19.37 ng/g w.w. in May (Figure 3B), and the difference between the 2 months was statistically significant (p < 0.05). This difference may be attributed to two factors: first, the bioaccumulation of CPs in kelp is directly constrained by the contaminant concentrations in the aquaculture environment, and second, as an organism with a specific growth cycle, kelp exhibits variation in its overall accumulation capacity across different developmental stages. As mentioned above, the concentration of SCCPs in seawater was higher in February than in May, and this same pattern was observed in kelp. We attributed this pattern in the kelp to the characteristics of kelp growth as it transitioned from the tender stage (February) to the mature stage (May). Kelp growth is driven by continuous differentiation of stem cells into new algal cells at the base of the blade, where the meristem is located. This growth exhibits distinct temporal characteristics, with the growth rate for the body length initially increasing and then decreasing [49]. When kelp enters the mature stage (May), tissue that formed at the tender stage (February) and would have accumulated high concentrations of SCCPs may have already decomposed in the seawater. Furthermore, as kelp enters the mature stage, its growth rate slows or even becomes negative, and its rate of nutrient uptake from the environment decreases. We concluded that the decomposition of kelp and the reduced uptake of nutrients in mature-stage kelp contributed to the lower SCCP concentrations in kelp in May than in February. We also analyzed the SCCP concentrations in the roots and tails of the kelp and found no significant differences between these two parts (p > 0.05).

In this study, the total concentration ranges of MCCPs and LCCPs in kelp were 0.007–0.87 and not detected at 4.45 ng/g w.w., respectively. The corresponding mean concentrations were 0.36 ± 0.30 and 2.03 ± 2.28 ng/g w.w. The average total concentrations in February were 0.42 ± 0.37 ng/g w.w. for MCCPs and 2.01 ± 2.83 ng/g w.w. for LCCPs, while those in May were 0.30 ± 0.20 ng/g w.w. for MCCPs and 2.01 ± 1.54 ng/g w.w. for LCCPs. Similarly to SCCPs, the total concentrations of MCCPs and LCCPs decreased as the kelp matured. Because of the current scarcity of data on the presence of SCCPs, MCCPs, and LCCPs in marine macroalgae, it was not possible to compare the CP concentrations in kelp from this study with concentrations in other regions.

In summary, in the main kelp mariculture areas along the Yellow Sea coast of Shandong, China, the mean total concentrations of SCCPs, MCCPs, and LCCPs were 82.67 ± 60.90, 0.36 ± 0.30, and 2.03 ± 2.28 ng/g w.w., respectively. These results show that kelp suffers from higher SCCP than MCCP and L, which is consistent with the concentrations of the pollutants observed in the seawater (Figure 2A). Furthermore, SCCPs are the predominant CPs in the main kelp mariculture areas along the Yellow Sea coast, which is primarily attributed to their high production volumes, extensive use, and high volatility, facilitating their transport into the marine environment.

3.3. Bioaccumulation of CPs in Kelp

BAFs were used to characterize the bioaccumulation potential of SCCPs in kelp. Kelp and seawater samples collected in Quanhai (sites S9 and S10) and Xunshan (sites S5 and S6) in February and May were selected for this analysis (Figure 4A–D). The log BAF ranges for the total concentrations of SCCPs, MCCPs, and LCCPs were 2.16–3.04, 0.85–1.99, and 2.83–5.09, respectively (

Table 1. Estimated daily intakes (EDIs) and hazard quotients (HQs) of CPs for dietary risk assessment.

	EDI [ng(kg·d)] Males	Females	HQ Males	Females
SCCPs	3.04874 × 10−4	3.59648 × 10−4	<0.01	<0.01
MCCPs	1.33583 × 10−6	1.57583 × 10−6	<0.01	<0.01
LCCPs	7.47943 × 10−6	8.82319 × 10−6	<0.01	<0.01

).

Significant differences were observed in the BAFs for the total concentrations of SCCPs and LCCPs in kelp between February and May (p < 0.05, Figure 4A,C). This was attributed to the different stages of growth of the kelp, with the kelp in February in the tender stage (a rapid-growth phase) [50] and the kelp in May in the mature stage (a slow-growth phase). The log BAF values were calculated for different CP congeners and revealed that kelp exhibited a strong bioaccumulation capacity for C₁₈Cl₈, C₁₈Cl₉, and C₁₉Cl₇ (Figure 4D). Furthermore, the bioaccumulation capacity of kelp for LCCPs was significantly higher than that for SCCPs or MCCPs. This difference is likely because LCCPs exhibit greater persistence to biodegradation [51] and may bioaccumulate more than SCCPs and MCCPs [52].

3.4. Congener Profiles of CPs in Seawater and Kelp

The distributions of CP carbon congeners in seawater and kelp were evaluated. For SCCPs, C₁₁-SCCP was the dominant congener in seawater (Figure 5A), while C₁₀ and C₁₁ were the dominant congeners in kelp (Figure 5B). The proportions of individual carbon congeners in both seawater and kelp showed no significant changes from February to May. Compared with previous studies, the SCCP congener profiles in seawater from this study were similar to those in Tokyo Bay (C₁₁-SCCP dominant) [53], but differed from those in the Southern Bohai Sea (C_10–11-SCCPs dominant) [33] and the East China Sea (C_10–11-SCCPs dominant) [54]. A comparison of the carbon congener abundance across different kelp growth stages revealed no significant variations.

In contrast to the main carbon congener proportions of SCCPs in seawater and kelp (Figure 5B), for MCCPs, seawater and kelp exhibited similar distribution patterns (C₁₄-MCCPs dominant). The average contributions of C₁₄-MCCPs to seawater were 70% in February and 67% in May, and those in kelp were 57% in February and 62% in May (Figure 5A,B). The congener profile of MCCPs in seawater differed from that of commercial CP-52 [54]. For LCCPs, seawater and kelp exhibited similar congener abundance profiles, with C₁₈-LCCPs dominant.

The CP chlorine congener distributions in seawater and kelp were also evaluated. For SCCPs, seawater and kelp exhibited similar characteristics, with Cl_5–7-SCCPs dominant (Figure 5C,D). The distribution of chlorine congeners in seawater was consistent with that in the East China Sea [54]. For kelp, the distribution of chlorine congeners showed no variation across different growth stages. For MCCPs, Cl_5–7-MCCPs were the dominant chlorine congeners in seawater, with a contribution of >70%. In kelp, Cl_6–8-MCCPs were dominant (Figure 5D) and also had a contribution of >70%. For LCCPs, both seawater and kelp were primarily composed of Cl_7–9 congeners (Figure 5C,D). From the CP congener concentration distribution profiles in seawater and kelp (Figure 5A–D), C₁₈Cl_8–9 was identified as the most abundant congener in both matrices. Seawater and kelp exhibited similar CP composition patterns, as reflected in the distribution trends of their carbon and chlorine congeners (Figure 5). In kelp, the concentrations of individual congeners decreased with increases in the carbon chain length. This result contrasts with previous findings that indicated SCCPs with longer carbon chains tended to accumulate in organisms [23,55]. This discrepancy may be attributed to the shift in the study subject from marine animals to marine plants. Marine animals are typically abundant in storage lipids, predominantly in the form of triglycerides [56], which serve as primary reservoirs for hydrophobic CPs. By contrast, macroalgae such as kelp mainly contain membrane lipids (e.g., glycolipids and phospholipids) and lack substantial lipid storage reserves [57]; this characteristic may limit the bioaccumulation capacity of CPs within algal tissues. The analysis of chlorine homologs in seawater gave an abundance order of Cl₆ (24.65%) > Cl₇ (23.42%) > Cl₈ (20.18%) > Cl₉ (17.15%) > Cl₅ (14.60%). This order may be attributed to the effects of hydrophobicity and molecular weight, as CPs with shorter carbon chains and lower degrees of chlorination have higher water solubility [29].

3.5. Dietary Risk Assessment

Dietary intake is the primary pathway for human exposure to CPs [58], and their associated health risks are primarily assessed using the EDI. In coastal cities, where marine food constitutes a large portion of the diet, it represents a major source of CPs intake for humans. The average concentrations of CPs in kelp from the Yellow Sea aquaculture area were used to estimate the total daily intakes of CPs for local adults (Table 1). The results were well below the TDI, and the HQ was <0.01. This indicates that CP concentrations in kelp from this region do not pose a dietary exposure risk.

Previous studies have reported EDI ranges of 1.14–3.70 ng/(kg·d) for SCCPs and 0.65–3.24 ng/(kg·d) for MCCPs in fish from Argentina [59]; 50.34 to 84.89 ng/(kg·d) for SCCPs in fish from Laizhou Bay, China [29]; and 47–190 ng/(kg·d) for SCCPs in meat, 16–29 ng/(kg·d) for SCCPs in eggs, 23–170 ng/(kg·d) for MCCPs in meat, and 2.4–8.9 ng/(kg·d) for MCCPs in eggs from South China [60]. By comparison, the EDI for kelp in this study was significantly lower than the reported values for meat and eggs. This difference may be attributed to the tendency of CPs to accumulate more readily in organisms with higher lipid contents. To support this assumption, experiments have confirmed that there is a significant positive linear correlation between CP concentrations and the lipid content of organisms [61,62]. Furthermore, CP concentrations in most animal-derived foods are generally higher than those in plant-derived foods [60].

In conclusion, the CPs in kelp from the mariculture areas along the coast of the Yellow Sea in Shandong do not currently pose a dietary exposure risk. It should be noted that the calculation of CP intake in this study considered the combined contributions of kelp, laver, and marine fish; therefore, the results may be overestimated. Although CPs are persistent and bioaccumulative, kelp is a seasonal crop with a limited bioaccumulation capacity for these compounds. Consequently, given current contamination levels in the marine environment, the dietary exposure risk associated with kelp consumption is expected to remain relatively stable in the future.

4. Conclusions

CP contamination was detected in both seawater and kelp within kelp mariculture areas along the coast of the Yellow Sea in Shandong, China. Furthermore, kelp bioaccumulated SCCPs, MCCPs, and LCCPs. In seawater, CP concentrations in February were significantly higher than those in May. This may be attributed to the fishing ban implemented in May and the resulting reduction in vessel traffic. Spatially, the CPs were primarily associated with human activities. Higher concentrations were observed in nearshore seawater than offshore seawater, and this difference was attributed to the influence of riverine inputs. For kelp, significant differences were observed in the CP concentrations and bioaccumulation capacities between the tender stage (February), which is characterized by rapid growth and high nutrient uptake, and the mature stage, which is characterized by slow growth rates. Analysis of CPs bioaccumulation in kelp revealed that, in terms of carbon chain length, kelp exhibited a relatively high bioaccumulation capacity for LCCPs; however, SCCPs remained the predominant constituents in the algal tissues, with CP concentrations displaying a decreasing trend as the carbon chain length increased. Regarding the degree of chlorination, Cl₆ and Cl₇ homologs were dominant in seawater, which was attributed to the solubilities of these homologs. The bioaccumulation patterns of homologs with different degrees of chlorination in kelp were generally consistent with those in seawater, which indicated that the bioaccumulation behavior was influenced by the CPs’ profile in the environment. A risk assessment indicated that, at current pollution levels, the consumption of kelp poses no dietary exposure risk to humans.