
==== Front
Eco Environ Health
Eco Environ Health
Eco-Environment & Health
2772-9850
Elsevier

S2772-9850(24)00026-7
10.1016/j.eehl.2024.03.002
Original Research Article
Effects of transient flow conditions on colloid-facilitated release of decabromodiphenyl ether: Implications for contaminant mobility at e-waste recycling sites
Li Yueyue 1
Huo Zebin 1
Ying Yuqin
Duan Lin
Jiang Chuanjia jiangcj@nankai.edu.cn
⁎
Chen Wei
College of Environmental Science and Engineering, Ministry of Education Key Laboratory of Pollution Processes and Environmental Criteria, Tianjin Key Laboratory of Environmental Remediation and Pollution Control, Nankai University, Tianjin 300350, China
⁎ Corresponding author. jiangcj@nankai.edu.cn
1 These authors contributed equally to this work.

26 3 2024
9 2024
26 3 2024
3 3 317324
14 12 2023
7 3 2024
7 3 2024
© 2024 The Author(s)
2024
https://creativecommons.org/licenses/by-nc-nd/4.0/ This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
Polybrominated diphenyl ethers (PBDEs) are ubiquitous contaminants, especially in the soil and groundwater of contaminated sites and landfills. Notably, 2,2′,3,3′,4,4′,5,5′,6,6′-decabromodiphenyl ether (BDE-209), one of the most frequently and abundantly detected PBDE congeners in the environment, has recently been designated as a new pollutant subject to rigorous control in China. Colloid-facilitated transport is a key mechanism for the release of PBDEs from surface soils and their migration in the aquifer, but the effects of hydrodynamic conditions, particularly transient flow, on colloid-facilitated release of PBDEs are not well understood. Herein, we examined the effects of typical transient flow conditions on the release characteristics of colloids and BDE-209 from surface soil collected from an e-waste recycling site by undisturbed soil core leaching tests involving multiple dry–wet cycles (with different drying durations) and freeze–thaw cycles. We observed significant positive correlations between BDE-209 and colloid concentrations in the leachate in both the dry–wet and freeze–thaw leaching experiments, highlighting the critical role of colloids in facilitating BDE-209 release. However, colloids mobilized during the dry–wet cycles contained higher contents of BDE-209 than those in the freeze–thaw cycle tests, and the difference was primarily due to the more intensive disintegration of soil aggregates and elution of newly formed inorganic colloidal particles (mainly primary silicate minerals such as quartz and albite) with low BDE-209 content during the freeze–thaw process. These findings underscore the necessity of considering transient flow conditions when assessing the fate and risks of PBDEs at contaminated sites.

Graphical abstract

Image 1

Highlights

• Positive correlations between 2,2′,3,3′,4,4′,5,5′,6,6′-decabromodiphenyl ether (BDE-209) and colloid concentrations were observed.

• Transient flow conditions can influence releases of colloids and BDE-209 from soil.

• There was less BDE-209 in colloids mobilized by freeze–thaw than by dry–wet cycles.

• Transient flow should be considered in predicting polybrominated diphenyl ether mobility in the subsurface.

Keywords

Hydrophobic organic contaminants
Colloid-facilitated transport
Dry–wet cycle
Freeze–thaw cycle
Undisturbed soil core
==== Body
pmc1 Introduction

Polybrominated diphenyl ethers (PBDEs) are widely used flame retardants in industries and consumer products, e.g., electronics, textiles, and plastics [1]. Due to their toxicities and persistence [2,3], the environmental fate and risks of PBDEs have raised concerns globally [[4], [5], [6]]. Three technical PBDE mixtures, including commercial pentaBDE (mainly tetrabromodiphenyl ethers and pentabromodiphenyl ethers), commercial octaBDE (mainly hexabromodiphenyl ethers and heptabromodiphenyl ethers), and commercial decaBDE (mainly 2,2′,3,3′,4,4′,5,5′,6,6′-decabromodiphenyl ether [BDE-209] and nonabromodiphenyl ethers), have been included in Annex A of the Stockholm Convention as persistent organic pollutants (POPs) [[7], [8], [9]]. Notably, BDE-209 is one of the most frequently detected and abundant PBDE congeners in the environment [10,11]. For example, high concentrations of BDE-209 were reported in soil [[12], [13], [14], [15], [16], [17]], freshwater and marine sediments [[16], [17], [18], [19], [20]], and indoor dusts [[21], [22], [23]] in China. BDE-209 has been banned in many regions of the world, including many countries in Europe and North America [24,25], and recently, it has been designated as a key new pollutant subject to rigorous control in China [26]. PBDEs have been detected worldwide [27,28], with high contamination levels detected in the surface soils of electronic waste (e-waste) recycling sites [29,30] and some landfills [31]. Although PBDEs exhibit strong hydrophobicity and low solubility in water, their congeners have been reported to leach from surface soils into aquifers [32], leading to the contamination of groundwater [[33], [34], [35]]. Soil colloidal particles, characterized by their relatively high mobility in porous media and affinities for various environmental pollutants, can serve as vectors for the transport of contaminants in the subsurface. Consequently, colloid-facilitated transport is a major mechanism for the release of hydrophobic organic pollutants, including PBDEs, from surface soils and their migration in groundwater [[36], [37], [38], [39]].

The extent of colloid-facilitated transport of PBDEs depends on the physicochemical properties of the soil colloids as well as the solution composition and hydrodynamic conditions of the aquifer. It has been reported that the capability of soil colloids to enrich PBDEs depends on the properties of soil, such as the presence of soil organic matter [40,41]. Moreover, according to leaching experiments with undisturbed soil cores collected from an e-waste recycling site (hereafter referred to as e-waste site), the ionic strength of infiltration water can significantly affect the release of colloids and BDE-209 from the surface soil [38]. Besides these factors, hydrodynamic conditions can significantly shape the transport behavior of colloids in porous media, especially in the shallow vadose zone [[42], [43], [44], [45]]. Compared to steady infiltration, transient flow conditions, typically related to rainfall, snowmelt [46], and accidental discharge of wastewater can affect the soil moisture content and the flow rate of soil water, leading to an increased release of colloids [[47], [48], [49], [50], [51]]. During dry–wet cycles, the drying period preceding rainfall can influence colloid release during transient infiltration by affecting the moisture content and capillary pressure of soil [48,52], but prolonged drying can diminish colloid release by reinforcing the soil aggregate structure [52] or hydraulically disconnecting previously saturated pores from the main leaching pathway [48]. Besides the dry–wet process, the freeze–thaw process of soil porewater due to seasonal and even diurnal temperature changes [53], which usually accompanies snowfall and snowmelt, may have additional effects on the release of soil colloids and contaminants. During freezing, the expansive phase transition of liquid water into ice exerts pressure on pore walls, disintegrating soil aggregates and promoting colloid formation and release into pore water during the thaw process [50]. Considering the critical roles of soil colloids in mobilizing PBDEs in the subsurface, it can be hypothesized that transient flow significantly affects the release of colloids as well as PBDEs from surface soil to the aquifer. Understanding the influences of typical transient flow conditions on the colloid-facilitated release of PBDEs is crucial for accurately forecasting the fate and risks of these hydrophobic contaminants at contaminated sites.

Here, we aimed to examine the effects of typical transient flow conditions on the release characteristics of colloids and BDE-209 from surface soil. Leaching tests involving multiple dry–wet cycles (with different drying durations) and freeze–thaw cycles were performed using undisturbed soil cores collected from an e-waste site in Taizhou, China. Our observations revealed significant positive correlations between BDE-209 and colloid concentrations in the leachate during both the dry–wet and freeze–thaw leaching tests, highlighting the critical role of colloids in facilitating BDE-209 release. Furthermore, we found that colloids mobilized in the dry–wet cycles contained higher contents of BDE-209 than those in the freeze–thaw cycles, and the difference is primarily due to the more intensive disintegration of soil aggregates and elution of newly formed inorganic colloidal particles with low BDE-209 content during the freeze–thaw process. These findings underscore the necessity of considering transient flow conditions when assessing the fate and risks of PBDEs at contaminated sites and can aid in the design of hydrological intervention measures to curtail the spreading of hydrophobic POPs into sensitive subsurface environments.

2 Materials and methods

2.1 Materials

Undisturbed soil cores were collected from an e-waste site (28°31′59.36″N, 121°21′49.42″E) in Taizhou, Zhejiang Province, China. The once-active e-waste-dismantling practices at the site had already ceased before the soil core sampling was performed, but the soil was laden with relatively high levels of PBDEs, predominantly BDE-209 (Table S1).

Before extracting the cores, the plants were carefully removed without damaging the topsoil layer. Using a stainless-steel sampler (7.5-cm diameter) designed specifically for a core holder, soil cores from the top 0- to 10-cm layer were collected. Both the sampler and the soil core within were immediately sealed to preserve their original condition. They were then transported to the laboratory and were stored at a temperature of 4 °C for use.

Information on the chemicals and reagents used in the study is provided in Supplementary Material (SM).

2.2 Experimental apparatus

The leaching apparatus (Fig. S1) for the undisturbed soil core leaching test was the same as that used in our previous study [38]. The soil core, together with the sampler, was placed into a core holder for the leaching test. Notably, the base of the core holder was designed with seven effluent ports, each covering a water flow collection area of 2.0 cm2. Simulated infiltration water (with the preparation method given in SM and Table S2) was pumped into the soil core with a designated flow rate of 1.5 cm/h (Darcy velocity).

2.3 Experimental procedure

2.3.1 Dry–wet cycle tests

Prior to the dry–wet cycles, the undisturbed soil core was pre-leached with a simulated infiltration solution (at 0.01-mM ionic strength and pH of 4.4) at a consistent flow rate of 1.5 cm/h for 24 h. This preliminary step was designed to stabilize the moisture content in the soil core and flush out soluble contaminants. Following this step, the leaching was stopped, allowing the core to drain naturally under gravitational force [54]. After 0.7 d of drying, the soil core underwent a 48-h leaching phase at the same flow rate using the aforementioned simulated infiltration. During this phase, effluents were collected from the ports into glass tubes at 12-h intervals for subsequent analyses. This process was repeated, and the soil core experienced five additional dry–wet cycles with varying drying durations (1.5, 2.5, 3.4, 4.6, and 8.8 days) [48].

2.3.2 Freeze–thaw cycle tests

A different soil core was used in the freeze–thaw cycle tests. Prior to the freeze–thaw cycles, the core was pre-leached and gravitationally drained using the same procedure as that described in Section 2.3.1. The core was then subjected to a freezing regime at −20 °C for 48 h, followed by a thawing period lasting 24 h under ambient conditions. Subsequent to the thawing, leaching resumed using the same simulated infiltration at 1.5 cm/h for 48 h. During this leaching phase, effluent samples were acquired at 12-h intervals for analyses. This entire procedure was then repeated, with the soil core undergoing four additional freeze–thaw cycles.

2.4 Analyses of leachates

Effluent samples from active ports were analyzed for the concentration, hydrodynamic diameter (Dh), ζ-potential, mineral composition, and morphology of colloids, as well as the concentration of BDE-209.

The concentration of colloids (mg/L) in the effluent samples was measured using an ultraviolet-visible spectrophotometer (G9821A, Agilent Technologies, USA) [44], with more information given in SM. The Dh and ζ-potential of colloids in the effluent were measured by using a Litesizer 500 instrument (Anton Paar). The mineral composition and morphology of the colloids in the effluent were characterized using an X-ray diffraction (XRD) instrument (Ultima IV, Rigaku, Japan) and a transmission electron microscope (TEM) (JEM-2800, JOEL, Japan). The concentrations of common metals (Al, Fe, Ca, Mg, K, Mn, and Na) in the surface soil and colloids were determined using inductively coupled plasma optical emission spectroscopy (Agilent 730, USA) after microwave digestion. The composition of organic substances in the effluent was qualitatively determined by excitation emission matrix (EEM) fluorescence spectroscopy. More detailed characterization methods are provided in SM.

Following liquid–liquid extraction (see SM for details) [55], BDE-209 concentration was determined on an Agilent 6890N gas chromatograph (GC) with an electron capture detector (ECD) (details given in SM). Moreover, the concentrations of 14 PBDE congeners in the surface soil (Table S1) were determined by gas chromatography–mass spectrometry (GC–MS) (7890A-5975C, Agilent Technology, USA) after Soxhlet extraction [41], with detailed methods provided in SM. Initial evaluations with GC–MS revealed that, among the 14 PBDE congeners present in the soil, only BDE-209 was detectable in the effluent. Therefore, a BDE-209 standard solution was used for the GC–ECD analysis.

Additionally, the pH and conductivity of the effluent were measured using a pH meter (model S210-B) and a conductivity meter (model FE30), both manufactured by Mettler Toledo, USA.

2.5 Quality assurance/quality control

Data with a signal-to-noise ratio exceeding a value of 3 were used for further analysis. The calibration curves yielded correlation coefficients > 0.994. For BDE-209, the method detection limit (MDL) stood at 0.07 ng/L. The recovery ratios for the surrogate standard, BDE-77, spanned between 67% and 106%. Due to the narrow range observed in surrogate standard recovery, the PBDE concentrations were not adjusted based on surrogate recovery. The recovery ratios of BDE-209 in spiked blanks (BDE-209 dissolved in simulated infiltration water, which underwent the same extraction and analysis procedures as the effluent samples) ranged from 77% to 102%.

2.6 Statistical analysis

Statistical analysis was performed using SPSS version 26.0 (IBM, USA). Pearson correlation analysis and linear fitting were performed for the concentrations of colloids and BDE-209.

3 Results and discussion

3.1 Effect of dry–wet cycles on the release of colloids and BDE-209

During the leaching test with different drying durations, we observed distinct preferential flow pathways in the soil core (Fig. 1 and Table S3): water primarily eluted from two of the seven ports, and no leachate was collected from the other five ports (denoted as No. 1 and 2). The water flux from each port was relatively constant during the tests, both exceeding the application rate of the simulated infiltration (i.e., 1.5 cm/h). Notably, port No. 2 was the primary drainage port, with a flux two or three times higher than that of port No. 1. These preferential flow paths could substantially enhance the release and downward migration of colloids and the adsorbed contaminants [48,50,54].Fig. 1 Effluent flux, colloid concentration, Dh, ζ-potential, conductivity, and pH from two active ports during dry–wet cycle tests at different drying durations. Influent Darcy velocity = 1.5 cm/h, ionic strength = 0.01 mM, pH = 4.4.

Fig. 1

The duration of the drying period played a pivotal role in colloid release. As the drying time increased to 2.5 or 3.4 d, an increase in colloid concentration in the leachate was observed; when the drying time further increased, colloid concentration decreased (Fig. 1 and Table S3). The colloid mass eluted from the soil core for different drying durations also followed this trend (Fig. 2), which has been noted in dry–wet leaching experiments by Majdalani et al. [52] and Mohanty et al. [48]. The increase in colloid release with longer drying time below the critical drying duration (CDD) could be attributed to the fact that the drying process can induce differential capillary stress within soil pores, resulting in colloid generation through the cracking of soil pore walls [48,[56], [57], [58]]. Since soil fractures are commonly coated with iron and manganese oxyhydroxides [48,59], higher concentrations of Fe and Mn are expected in the leached colloids than in bulk soil [48]. In this study, the concentrations of Fe and Mn in colloids were 1.41 and 1.37 times that in bulk soil, respectively (Fig. S2), corroborating the hypothesis that dry–wet cycles can generate colloids through the cracking of soil pore walls. However, beyond the CDD, a reduction in colloid release with longer drying time could result from hydraulic disconnection of micropores within the soil matrix from the preferential flow paths (mainly macropores and fractures) after prolonged drying, leading to diminished migratory potential of colloids from the soil matrix [48]. Furthermore, prolonged drying may result in the deposition of minerals and salts, impeding colloid mobilization [48,52,60]. It is noted that the CDD observed in this study (i.e., 2.5–3.4 d) was similar to that reported by Mohanty et al. [48] but was shorter than that (9–11 d) reported by Majdalani et al. [52], likely due to differences in the properties of the soil cores used in the studies. Specifically, the soil cores used in our study and by Mohanty et al. [48] contained abundant macropores or cracks, and the higher water permeability and abundance of preferential flow paths allowed for swifter drainage of water. Conversely, the soil used by Majdalani et al. [52] lacked macropores or cracks, leading to a longer duration for the drying front to reach the soil matrix. Moreover, the low-flux port No. 1 exhibited a CDD (3.4 d) longer than that for the high-flux port No. 2 (2.5 d), further corroborating the dependence of the CDD on the water permeability of the soil.Fig. 2 Effect of antecedent drying duration on the total mass of colloids mobilized through the active ports during dry–wet cycles.

Fig. 2

With the varying drying duration, the average Dh of colloids in the effluent remained relatively stable, ranging from 428 ± 25 nm to 508 ± 50 nm (Fig. 1 and Table S3). However, the ζ-potential was more negative for the drying durations of 2.5 and 3.4 d (Fig. 1 and Table S3), when relatively high colloid concentrations were observed. The similar trends between the ζ-potential and concentration of eluted colloids may be ascribed to the role of electrostatic repulsion in colloid mobilization: when the dried soil is saturated with infiltrating water, the soil colloid particles that bear more negative ζ-potential experience stronger electrostatic repulsion from the soil matrix, making it more likely for these particles to detach from soil pore walls and migrate with the infiltrating water.

The drying duration had a limited effect on the conductivity or pH of the effluent. However, during each dry–wet cycle, the conductivity in the initial leaching stage (i.e., by the first 12 h) was much higher than that in subsequent stages. This trend was due to the deposition of soluble salts on the surface of soil pore walls during drying, and the accumulated salts quickly dissolved upon the onset of leaching, resulting in the initial high conductivity (ranging from 252.0 to 381.0 μS/cm, on average 331.8 ± 29.5 μS/cm). As leaching proceeded, the salts accumulated on the pore walls were depleted, and the conductivity of the effluent depended on the content of ions in the soil matrix exchangeable with the infiltrate, which was relatively stable for a given soil core. By the end of the leaching in each dry–wet cycle, the conductivity was in the narrow range of 70.3–96.0 μS/cm, with an average of 83.9 ± 7.9 μS/cm. Regardless of the drying duration, the pH of the effluent was consistently between 7.4 and 7.8, though the pH of the influent was 4.4. The insensitivity of effluent pH to influent pH during soil core leaching tests has been previously observed and ascribed to the pH buffering capacity of soil [38].

During the dry–wet leaching experiments, BDE-209 concentration in the effluent samples from port No. 2 was in the range from 0.09 to 0.48 ng/L (Table S3). However, BDE-209 was not detected in the effluent from port No. 1, which was attributed to the low colloid mass in the leachate from port No. 1, rendering the BDE-209 mass in the analytical sample below detection. The average BDE-209 concentration during the dry–wet leaching experiments was 0.21 ± 0.15 ng/L (Table 1), which was lower than that in the effluent from soil cores collected from an e-waste site in Tianjin, despite comparable colloid concentrations in the effluents [38]. This difference is likely due to the much higher BDE-209 content in soil at the site in Tianjin [41].Table 1 BDE-209 concentration in the effluents from soil cores under different hydrodynamic conditions.

Table 1Hydrodynamic condition	Range (ng/L)	Mean ± SD (ng/L)	
Dry–wet cycles	ND−0.48	0.21 ± 0.15	
Freeze–thaw cycles	0.08–0.34	0.19 ± 0.09	
BDE-209, 2,2′,3,3′,4,4′,5,5′,6,6′-decabromodiphenyl ether; ND, below the method detection limit (0.07 ng/L); SD, standard deviation.

3.2 Effects of freeze–thaw cycles on the release of colloids and BDE-209

In the freeze–thaw leaching experiments, water also eluted from only two of the seven ports (Fig. 3). Compared to the dry–wet cycles, the freeze–thaw cycles led to an increase in the released colloidal concentration (Fig. 3 and Table S4). During each freeze–thaw leaching cycle, peak colloidal concentrations (as high as 76 mg/L) were observed in the effluent collected within the first 12 h, which then gradually decreased. This high initial colloid release was primarily attributed to the volumetric expansion of water during the freezing process, exerting mechanical stress on the pore walls, which disintegrates soil aggregates and promotes colloid formation [50,[61], [62], [63], [64]]. The average Dh of colloids in the effluent ranged from 317 ± 25 to 360 ± 16 nm (Fig. 3 and Table S4), lower than that eluted during the dry–wet cycles (Fig. 1 and Table S3), suggesting greater extents of disintegration of soil aggregates during freeze–thaw cycles. The ζ-potential of colloids in the effluent was relatively stable during each freeze–thaw cycle test, but in the first cycle, the ζ-potential was more negative than in other cycles, possibly due to the elution of microbially derived organic substances, as indicated from the excitation/emission peak at 300/375 nm in the EEM spectrum (Fig. S3a) [65,66]. By contrast, in the following four cycles, the effluent only contained low levels of protein-like substances, as indicated by the excitation/emission peak at 220/300 nm in the EEM spectra (Fig. S3b–e) [65,67,68].Fig. 3 Effluent flux, colloid concentration, Dh, ζ-potential, conductivity, and pH at two active ports during freeze–thaw cycle tests. Influent Darcy velocity = 1.5 cm/h, ionic strength = 0.01 mM, pH = 4.4.

Fig. 3

The freeze–thaw cycles appeared to exert negligible influence on the conductivity of the effluent (Fig. 3). During each leaching test, the conductivity declined over time, from 91.6–127.0 μS/cm (on average 110.4 ± 10.4 μS/cm) to 66.9–73.0 μS/cm (on average 70.3 ± 2.2 μS/cm). The relatively low conductivity corresponds to the lower ionic strength of the effluent, which may contribute to the smaller Dh and slightly more negative ζ-potential of the colloids eluted from the freeze–thaw cycle tests. Moreover, the effluent pH remained stable between 7.5 and 7.7 also because of the pH buffering capacity of the soil.

During the freeze–thaw leaching experiments, BDE-209 concentrations in all of the effluent samples were above the MDL of 0.07 ng/L, with values ranging from 0.08 to 0.34 ng/L and, on average, 0.19 ± 0.09 ng/L (Table 1 and Table S4). Notably, compared with the dry–wet cycles, although the freeze–thaw cycles enhanced colloid release, the BDE-209 concentrations in the effluent were not significantly different between the two sets of experiments (Table 1). This result appears to contradict the predictions from the colloid-facilitated release of contaminants. However, further analysis shows that colloid-facilitated release still plays a key role in the release of BDE-209 from the soil under different transient flow conditions, as discussed in the following.

3.3 Mechanisms for differential effects of dry–wet and freeze–thaw cycles on colloid-facilitated release of BDE-209

It has long been recognized that soil colloids play a key role in facilitating the release and subsurface migration of hydrophobic organic pollutants, such as polycyclic aromatic hydrocarbons [[69], [70], [71]], atrazine [72,73], bismerthiazol [74], polychlorinated biphenyls [75], and PBDEs [38,40,76,77]. Consistent with the previous findings, in this study, we observed significant positive correlations (p < 0.01) between the concentrations of colloids and BDE-209 in leachates from soil cores in either the dry–wet or the freeze–thaw cycle tests (Fig. 4), although when data from both sets of cycle tests were pooled together, there was no significant correlation (p = 0.05). Notably, due to the hydrophobicity and extremely low solubility of BDE-209, the slope of the regression curve can, to some extent, indicate the content of BDE-209 bound to colloids in the effluent. The slope of the regression curve for the dry–wet cycle tests was higher than that for the freeze–thaw cycle tests (Fig. 4), suggesting that the colloids eluted during the dry–wet cycle tests had higher BDE-209 content than those eluted during the freeze–thaw cycle tests.Fig. 4 Correlation between colloid and BDE-209 concentrations in the effluent from soil cores, measured in dry–wet and freeze–thaw leaching tests. The dashed lines represent the regression curves.

Fig. 4

The mechanisms for the differential effects of dry–wet and freeze–thaw cycles on the colloid-facilitated release of BDE-209 were examined by characterizing the physicochemical properties of the eluted colloids. Organic carbon can enhance the capability of soil colloids to bind PBDEs [40,41], and different organic contents in the eluted colloids between the dry–wet and freeze–thaw cycles may contribute to the different BDE-209 contents in the colloids. The small amount of colloids in the effluent precludes the quantitative analysis of their organic carbon contents by elemental analysis. However, according to the EEM spectra (Fig. S4), the effluent in the dry–wet cycle tests contained microbially derived and protein-like organic substances at concentrations higher than those in the freeze–thaw cycle tests (except the first cycle). Nevertheless, the effluent sampled from the first cycle of the freeze–thaw cycle tests contained relatively low BDE-209 concentrations (Table S4), despite the abundance of organic matter (Fig. S3). Therefore, other factors likely have led to the differential effects of the dry–wet and freeze–thaw cycles. As mentioned in Section 3.2, the effluent in the freeze–thaw cycle tests contained higher colloid concentrations than that in the dry–wet cycle tests. Moreover, the second to fifth cycles of the freeze–thaw cycle tests yielded higher colloid concentrations than did the first cycle (Fig. 3), despite the much lower organic matter abundance in the effluents (Fig. S3). Therefore, it can be inferred that the freeze–thaw cycles induced the release of more inorganic colloidal components, and it is hypothesized that these inorganic colloids contained low contents of BDE-209, which diluted the overall BDE-209 content in the eluted colloids.

According to XRD characterization results (Fig. 5), the major mineral components of colloids in the effluent from the dry–wet cycle tests were calcite (CaCO3), whereas in the freeze–thaw cycles, the eluted colloids were mainly composed of CaCO3, quartz (SiO2), and albite (NaAlSi3O8). Note that clay minerals are also important components of soil colloids [50], though their characteristic peaks were not observed in the XRD patterns, possibly due to their low crystallinity in the effluents as well as due to the incomplete separation of nanoscale clay particles during the centrifugation (at 1617×g for 25 min). It has been reported that the capillary stress during dry–wet cycles may not mobilize quartz colloids, whereas the mechanical stress during the freeze–thaw processes could break bulk particles of quartz and possibly other primary silicate minerals (e.g., albite) to form the corresponding colloidal particles [50,53,78]. Indeed, TEM analysis (Fig. S5) revealed that particles with smoother surfaces were observed in the effluent from the freeze–thaw leaching experiments, which were likely the colloidal particles formed via the breaking of bulk minerals and exhibiting smooth cleavage planes. These newly formed inorganic colloidal fractions were supposed to contain lower contents of BDE-209, contributing to the different slopes of the regression curves in Fig. 4. Notably, the intercept of the regression curves on the abscissa was higher for the freeze–thaw cycle tests than for the dry–wet cycle tests (Fig. 4), further substantiating the hypothesis that the colloids eluted from the freeze–thaw cycle tests contained higher fractions of particles with extremely low BDE-209 content.Fig. 5 X-ray diffraction patterns of mobilized colloids from dry–wet cycles and freeze–thaw cycles.

Fig. 5

4 Conclusion

Leaching experiments using undisturbed soil cores under dry–wet and freeze–thaw cycle conditions showed significant positive correlations between concentrations of colloids and BDE-209 in the leachate. However, colloids mobilized during the dry–wet cycles contained higher contents of BDE-209 than those in the freeze–thaw cycle tests. These results demonstrate the critical role of soil colloids in facilitating BDE-209 release and downward migration in the subsurface, which varies under different transient flow conditions. The differences can be attributed to varying degrees of alteration in soil structure under these conditions, leading to distinct colloid types. During dry–wet cycles, the drying process lowers soil moisture and induces capillary stress, breaking down soil pore walls and generating colloids with more BDE-209. Conversely, during freeze–thaw cycles, the mechanical stress due to the expansion and contraction of water and ice disrupts soil aggregates more intensely, forming inorganic colloidal particles (mainly primary silicate minerals such as quartz and albite) with low BDE-209 content. These results underscore the significance of transient flow conditions in altering the mobility of PBDEs in the subsurface, emphasizing the necessity of considering these conditions when predicting the fate and risks of PBDEs at contaminated sites. Furthermore, these findings can aid in the design of hydrological intervention measures to curtail the colloid-facilitated spreading of hydrophobic POPs into sensitive subsurface environments, so as to mitigate human exposure to these hazardous substances through groundwater.

CRediT authorship contribution statement

Y.Y.L. and Z.B.H.: data curation, investigation, visualization, writing−original draft. Y.Q.Y.: investigation. L.D.: writing−review and editing. C.J.J.: conceptualization, funding acquisition, investigation, supervision, writing−original draft, writing−review and editing. W.C.: funding acquisition, supervision, writing−review and editing.

Declaration of competing interests

There are no conflicts of interest to declare.

Acknowledgments

This work was supported by the 10.13039/501100012166 National Key Research and Development Program of China (2019YFC1804202 ), the 10.13039/501100001809 National Natural Science Foundation of China (22276101 and 22020102004 ), the 10.13039/501100012226 Fundamental Research Funds for the Central Universities (63233056 ), and the Ministry of Education of China (T2017002 ).

Appendix A Supplementary data

The following is the Supplementary data to this article:Multimedia component 1

Multimedia component 1

Appendix A Supplementary data to this article can be found online at https://doi.org/10.1016/j.eehl.2024.03.002.
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