Dissolved and Particulate Beryllium Isotopes in the Pearl River Estuary: Their Geochemical Behavior in Estuarine Water and Potential Contributions From Anthropogenic Sources

The ratio of cosmogenic ¹⁰Be and its stable isotope ⁹Be has been used as a proxy of long-term continental weathering fluxes and denudation rates, but transport processes of these isotopes from river water to estuarine water and seawater, as well as interference of potential anthropogenic source of ⁹Be on natural ¹⁰Be/⁹Be around populated estuaries are not well constrained. Here, we present results of ¹⁰Be and ⁹Be concentrations of dissolved and reactive particulate phase in the Pearl River Estuary (PRE) and its eight major outlets. The concentrations of Cu, Cd, and Pb are also measured, allowing us to assess their contamination levels and anthropogenic source together with ⁹Be by the geo-accumulation index (I_geo–reac) and enrichment factor (EF). A wide distribution pattern of dissolved ¹⁰Be (137–1,194 at/g_water) and ⁹Be (0.781–8.31 × 10^–12 g/g_water) among these outlets is observed. The distribution coefficients (K_d) of both isotopes between sediment and water are in the order of 10⁵, and on average only 5% of ¹⁰Be exists as dissolved form. Compared with total meteoric ¹⁰Be deposited on the river basin, 23% of the meteoric ¹⁰Be is retained while 38% of ¹⁰Be finally escape the estuary and is transported into coastal seawater. Despite the high contamination levels of Cu and Cd, the lower I_geo–reac and EF values of ⁹Be indicate that ⁹Be is hardly polluted by anthropogenic source. Thus, the ¹⁰Be/⁹Be in the PRE area is mainly associated with natural processes instead of human activities.

Introduction

Beryllium isotopes have played an important role in many geoscience studies. For example, the ¹⁰Be/⁹Be ratio in authigenic phase of marine sediment has been used as a tool to reconstruct the history of Earth’s magnetic fields using the inverse relationship between production rate and geomagnetic dipole values (Frank et al., 1997; Christl et al., 2003; Valet et al., 2014; Simon et al., 2016a, 2018, 2019, 2020). The meteoric ¹⁰Be in soil, river water and suspended sediments also provides a tool to trace erosion rates and denudation rates (von Blanckenburg et al., 2012; Wittmann and von Blanckenburg, 2016).

The cosmogenic nuclide ¹⁰Be (T_1/2 = 1.39 Myr) is produced by spallation of N and O atoms reacted with cosmic rays mainly in the atmosphere, and scavenged to Earth’s surface by wet and dry precipitation (Lal and Peters, 1967). The stable isotope ⁹Be are released from continental silicate weathering, and transported by river water to estuaries and seawater. The riverine input is the main source of ⁹Be to seawater, while the dissolved load only accounts for 13–30% of marine ⁹Be input (Suhrhoff et al., 2019). Thus, the ratio of ¹⁰Be and ⁹Be in seawater is determined by riverine input of ⁹Be and downward flux of meteoric ¹⁰Be (Kusakabe et al., 1991; von Blanckenburg and Bouchez, 2014).

Controlled by mineralogy, particle grain size and the presence of organic and inorganic ligands, the sorption-desorption process of ¹⁰Be and ⁹Be between suspended sediments and water plays an important role in affecting the ¹⁰Be/⁹Be ratios of estuarine water and seawater (Wittmann et al., 2012; Boschi and Willenbring, 2016a, b). The distribution/partition coefficient between particulate and dissolved phases (K_d = S/C, where S and C represent concentrations of particulate Be per mass and in solution at equilibrium, respectively) is used to quantitatively describe distributions under thermodynamic equilibrium conditions (You et al., 1989). Within the applicable range, the K_d of Be increase with pH significantly at pH values ≤6 (You et al., 1989; Brown et al., 1992b), and this increasing trend can extend up to pH∼8 (Suhrhoff et al., 2019). The K_d also decreases with suspended load (Aldahan et al., 1999) as observed with heavy metals (Balls, 1989). Besides silicate weathering, anthropogenic activities, such as mining, metallurgical industry and electronic manufacturing could also release beryllium to river water (Taylor et al., 2003; Shah et al., 2016; Zhuang et al., 2016). Both the natural processes and potential impacts of human activities contribute to the transport of riverine ⁹Be to oceans.

The Pearl River (PR; Zhujiang River) is the third longest river in China, with a total 450,000 km² basin area. The three major tributaries of the PR are the West River (Xijiang River), the North River (Beijiang River), and the East River (Dongjiang River). These three rivers, as well as other small rivers, merge at the Pearl River Estuary (PRE) in southern Guangdong Province. The PRE area, especially the Lingdingyang surrounded by the Guangdong-Hong Kong-Macao Greater Bay Area, is one of the most populated and industrialized area in China. The PRE area receives about 64% of industrial sewage of Guangdong Province (Ma et al., 2005), and the annual sewage discharge into Lingdingyang is much higher than that of Modaomen and Huangmaohai (Dai et al., 2014). Heavy metals of the PRE have been reported to transfer from upstream to delta area (Zhen et al., 2016), and some of them sourced from industrial and domestic sewage (Zhuang et al., 2018).

The early work of beryllium isotope distributions in the PRE waters by Kusakabe et al. (1991) didn’t show any trace of anthropogenic beryllium pollutions. However, this area has experienced tremendous industrial development during the last three decades and there has been no new data in this area. It is unknown if anthropogenic beryllium is released together with other heavy metals which have polluted the PRE (Zhuang et al., 2018). If so, the normalization protocol of ¹⁰Be by ⁹Be in many studies, such as Earth surface erosion processes and the reconstruction of the geomagnetic field parameters, would be challenged. Since beryllium is highly particle-reactive, both dissolved and particulate data of ¹⁰Be and ⁹Be from the PRE are needed in order to gain a full picture of their sources and behavior in transport process.

In this study, we measure distributions of ¹⁰Be and ⁹Be, along with heavy metals (Cu, Cd, and Pb) in dissolved and reactive phases of suspended sediments among eight outlets of the PR and within the Lingdingyang Bay, to assess the partition of beryllium isotopes between particles and water, the export flux of riverine ¹⁰Be into the South China Sea (SCS), and evaluate impacts of potential anthropogenic ⁹Be around industrial cities on ¹⁰Be/⁹Be ratios in the estuarine waters.

Materials and Methods

Study Site

The annual discharge of the major tributaries of the PR, the West River, the North River, and the East River, is 219.7, 42.1, and 23.4 × 10⁹ m³ year^–1, respectively (Dai et al., 2009). The suspended sediment flux of the West River (66.8 × 10⁶ t year^–1) is higher than those of the North River (5.4 × 10⁶ t year^–1) and the East River (2.5 × 10⁶ t year^–1), disproportionately higher than the discharge ratio (Zhang et al., 2011). The West River basin is typically a karst area dominated by carbonate (Cai et al., 2008).

The PR system drains into the northern SCS through eight outlets (river mouths), including Humen (HM), Jiaomen (JM), Hongqimen (HQ), and Hengmen (HE) in the northern part, and, Modaomen (MD), Jitimen (JT), Hutiaomen (HT), and Yamen (YM) in the south (Figure 1). The broadly defined PRE consists of three sub-estuaries, Lingdingyang, Modaomen, and Huangmaohai. Four of the outlets (HM, JM, HQ, and HE) sit along the western shore of the northern part of the Lingdingyang Bay. The Lingdingyang Bay receives river water mainly from the East River, the North River, and small branches of the West River, all of which account for more than half of the PR discharge (Pan et al., 2020). The Huangmaohai Bay sits in the west of the PRE, with two branches of the West River injected via Hutiaomen and Yamen. Between these two bell-shaped sub-estuaries sits the shallow Modaomen Bay, which also receives the water of West River and empties into the northern SCS via Modaomen and Jitimen outlets.

FIGURE 1

Figure 1. Study area and sampling sites of the PRE. The sampling sites at eight outlets and the Lingdingyang Bay are marked with red triangles and blue squares, respectively.

Sample Collection

The river waters and estuarine waters of the PRE were sampled during May 3, 2017 and May 8, 2017 by fish boats. Sampling stations include eight outlets of the PR, and a meridional section of four stations inside the Lingdingyang Bay (Figure 1). At each station, waters were sampled by a 3 L home-made polymethyl methacrylate sampler attached with an iron block at the middle point of river. The sampler was launched repeatedly until the total amount of ∼20 L water for each sample was collected. Both surface (0.5 m depth) and bottom (above 0.5–1.5 m of river bed) waters were sampled and collected in acid-cleaned HDPE bottle. This single-point sampling procedure at the middle point of the vertical profile was reported to cause less than 10% error compared with five-point procedure (Ni et al., 2008). Immediately after sampling, the pH value and salinity were measured with calibrated FiveGo portable pH meter and portable salinity meter, respectively. Each sample was filtered through a 142-mm diameter 0.4 μm nitrocellulose filter, and acidified with 20 mL concentrated HCl to pH < 2 within 2 h of sampling. All used filters were sealed in Ziploc bags and taken to lab, where the suspended sediments were freeze dried for 48–72 h, scraped from the filter and weighed to calculate the suspended load in water. For each filtered sample, an aliquot of 2 L was separated to an acid-cleaned 2 L bottle for ⁹Be and heavy metals measurements. All water samples were not ultra-filtered, so our “dissolved” phase include both “truly dissolved” and colloidal beryllium.

Sample Preparation

Dissolved Beryllium Isotopes

For the determination of the dissolved ¹⁰Be concentrations, the ∼18 L acidified water samples were added with 0.4 mg ⁹Be carrier (AccuStandard^® ICP-MS standard solution) and 120 mg Fe carrier, shook and left for equilibrium for 48 h. Then the pH value was adjusted to 8.5 with ammonium hydroxide, Be(OH)₂ was co-precipitated with Fe(OH)₃ and Mg(OH)₂ after equilibrium and setting for another 48 h. The precipitate was collected in a 50 mL centrifuge tube by decanting and centrifuging, and rinsed by ultrapure water (18.2 MΩ MilliQ^® water) twice. The cleaned precipitate was dissolved in 3 mL 10.2 M HCl, and the chromatography separation procedure of ¹⁰Be was slightly modified from Simon et al. (2016b). Briefly, the Fe was separated by passing samples through a 12 mL column filled with 6 mL Dowex^® 1 × 8 (100–200 mesh) anion exchange resin, the beryllium fraction was collected by 15 mL 10.2 M HCl and evaporated to dryness. After redissolved in 3 mL 10.2 M HCl, the anion resin step was repeated once, evaporated to dryness and redissolved in 1 mL 1 M HCl. Boron and aluminum were separated by loading redissolved sample solution onto a 12 mL column containing 5 mL Dowex^® 50 × 8 (100–200 mesh) cation exchange resin. After eluent of 15 mL 1M HCl, the beryllium fraction was collected in the next 35 mL 1M HCl, evaporated to dryness and redissolved in 10 mL 1M HNO₃ solution. The Be(OH)₂ was precipitated at pH of 8.5, rinsed by pH 8 water for three times, redissolved in 0.1 mL concentrated HNO₃ and converted to BeO at 850°C for 1 h. The BeO was mixed with Nb powder and pressed into cathode-holder for AMS measurements. Blank samples were made by co-precipitate 0.4 mg ⁹Be and 120 mg Fe in 1 L ultrapure water and experienced the same chemical procedures.

The preconcentration and purification method to analyze dissolved ⁹Be concentration was adjusted from Tazoe et al. (2014). Briefly, ∼250 g acidified water sample was neutralized to pH ∼7 and passed through a column contained 1 g Wako-Gel C-300 silica gel, the beryllium (⁹Be) was retained on the column. The matrix residue such as Na, Ca, and Mg were separated by 15 mL ultrapure water and 15 mL 5 mM EDTA solution. The residual EDTA was removed by another 15 mL ultrapure water, and ⁹Be was eluted by 15 mL 0.5 M HNO₃ and collected in a PFA vessel. The eluent was evaporated to dryness and redissolved in 1.2 mL 0.5 M HNO₃. We compared our beryllium recovery with UV-treated samples but didn’t find significant difference, thus the organic matter doesn’t seem to affect the beryllium recovery of our water samples. The recovery rate of preconcentration method for dissolved ⁹Be were tested by adding known amount of ⁹Be to 50 mL pure water, resulting yields of 99.2 ± 2.3%.

Suspended Sediments

The “reactive” authigenic phases [amorphous and crystalline Fe-Mn (hydr)oxides] of sediment are main carriers of ¹⁰Be and non-silicate-bound ⁹Be, and the grain size effects are removed by normalizing ¹⁰Be by ⁹Be (Bourlès et al., 1989; Wittmann et al., 2012). To extract the reactive phase of suspended sediment, the simplified sequential leaching procedure was adopted following that of Wittmann et al. (2012) which was developed from protocols of Tessier et al. (1979) and Wiederhold et al. (2007).

Suspended sediment of 0.1–0.5 g was used for leaching, the volumes of following leaching solutions were based on 0.5 g sediment. The amorphous oxyhydroxides phase was leached by 15 mL of 0.5 M HCl for 24 h at room temperature (water bath of 25°C and keep shaking at 200 rpm). The residue was separated by centrifuging, then rinsed by another 10 mL 0.5 M HCl, vortexed for 30 s and centrifuged. The 15 mL HCl leachate and the following 10 mL rinsing HCl were merged as leachate I, which contains the amorphous oxides fraction and the exchangeable fraction. The corresponding residue I was used for the extraction of crystalline oxide phase. Residue I was further leached by 10 mL 1M HCl in 1M hydroxylamine hydrochloride solution for 4 h at water bath of 80°C (keep shaking at 250 rpm). After centrifuging, the residue was rinsed by 5 mL 0.5 HCl, vortexed for 30 s and centrifuged to separate residue II, and the 5 mL rinsing solution was incorporated with leachate II (the crystalline oxides fraction). Aliquots of 2 mL leachate I and II were separated to analyze ⁹Be and heavy metal concentrations. Since our sequential extraction steps to achieve amorphous and crystalline phases are different from that of Poulton and Canfield (2005), the two phases we extracted are not completely corresponding to theirs.

To determine the reactive ¹⁰Be concentrations, the leachate I and II were spiked with 0.4 mg ⁹Be carrier, added with 5 mL concentrated HNO₃, evaporated to dryness and redissolved in 1.5 mL 10.2 M HCl. The subsequent column chromatography procedures were the same as for water samples.

Measurements

The ¹⁰Be measurements of all samples were performed at the French AMS national facility ASTER (CEREGE). The ¹⁰Be standard used was CEREGE in-house standard (STD-11, ¹⁰Be/⁹Be = 1.191 ± 0.013 × 10^–11) which can be traced back to NIST 4325 Standard Reference Material (Braucher et al., 2015). Blanks of measured atomic ratio were at 10^–15 levels and deducted from measurement results. The ¹⁰Be concentrations were calculated from measured ¹⁰Be/⁹Be ratio and mass of ⁹Be carrier added (Simon et al., 2016b). Uncertainties of measured ¹⁰Be/⁹Be are around 3% and propagated to calculated ¹⁰Be concentrations. The ⁹Be concentrations were measured by a graphite furnace atomic absorption spectrometer (GF-AAS, Thermal Scientific ICE 3500) at Peking University. The detection limit of GF-AAS for ⁹Be measurement is 0.0044 μg/L. The leachates were diluted for ∼50 times before measurements. For both water sample and leachate, standard addition method was used to eliminate the matrix effects. Each measurement was repeated four times and the standard deviation was below 3%. Both ⁹Be and ¹⁰Be concentrations in amorphous and crystalline fractions were measured separately.

The concentration of Cu, Cd, Pb in leachates were also analyzed by GF-AAS, and the concentration of Fe was measured by Flame-AAS. To determine the dissolved Cu, Cd, Pb concentration, ∼10 mL filtered water samples were added with 2 mL concentrated HNO₃, evaporated to dryness and redissolved in 0.5M HNO₃ to fulfill 2–10 times enrichment. The recovery rate of Cu, Cd and Pb is 99 ± 1, 97 ± 2, and 96 ± 4%, respectively. The uncertainties (2σ) of heavy metal measurement are typically below 3%.

Calculation of Geo-Accumulation Index and Enrichment Factor

To evaluate contamination levels of heavy metal, the geo-accumulation index (I_geo) was first introduced by Müller (1969), and defined as the following equation:

$$I_{geo}={\log }_2\frac{{\mathrm{C}}_{\mathrm{n}}}{1.5{\mathrm{B}}_{\mathrm{n}}}(1)$$

where C_n is the measured concentrations of ⁹Be and heavy metals in sediments, B_n is the geochemical background values. The background values of Cu, Cd, and Pb in the study area are selected as 17, 56, and 36 μg/g (MEP (Ministry of Environmental Protection of the People’s Republic of China), 2009), while the background value of ⁹Be is adopted as 1.19 μg/g (CEMS (China Environmental Monitoring Station), 1990). For each metal, the I_geo is assessed for its contamination condition, from uncontaminated (I_geo ≤ 0) to extremely contaminated (I_geo ≥ 5) conditions (Hanif et al., 2016; Zhuang et al., 2018). To calculate the I_geo–reac, the C_n is adopted as the concentrations of metals in the reactive phases of the suspended sediments from the bottom samples. Since both the definition of I_geo and the reported background values are based on the complete digestion of the sediments, the evaluation criterion of calculated I_geo–reac with concentrations in reactive phase should be roughly 1 level higher than that of I_geo. Thus, the uncontaminated criterion is I_geo–reac ≤ −1, and I_geo–reac ≥ 4 would represent extremely contaminated condition.

The enrichment factor (EF) has been used to assess contamination level as well as to discriminate anthropogenic sources of heavy metals (Rahn and McCaffrey, 1979). Because the anthropogenic source of Fe is negligible compared to its high abundance in nature (Zhou et al., 2007), Fe is widely used as one of the reference elements in calculation of EFs, besides Al and Ti (Reimann and Caritat, 2000; Gu et al., 2018). In this study, we choose Fe as the reference element. The concentrations of heavy metals in sediments are normalized to the concentrations of the reference element (Fe), and compared to that of background values as equation (2):

$$EF=\frac{{(C_M/C_{Fe})}_{sample}}{{(C_M/C_{Fe})}_{background}}(2)$$

where Fe is selected as a normalizing element (C_M/Fe)_sample and (C_M/Fe)_background are the concentration ratio of heavy metals and Fe in suspended sediments from bottom water and background value, respectively. The background value of Fe from Guangdong Province soils is 28.8 mg/g (CEMS (China Environmental Monitoring Station), 1990). We assume that the elemental ratio in silicate residue is approximate to that of reac phase to assess the contamination level with our results of reactive phases.

Results

Dissolved ⁹Be and ¹⁰Be Concentrations

The concentration of dissolved beryllium isotopes are listed along with salinity and pH values in Table 1. The units of dissolved concentrations of ⁹Be ([⁹Be]_diss) and ¹⁰Be ([¹⁰Be]_diss) are adopted as g/g_water and at/g_water, respectively. The [⁹Be]_diss range from 0.781 to 8.31 × 10^–12 g/g_water (i.e., 87–921 pmol/kg). They are higher than the average value (0.66 × 10^–12 g/g_water) for the shallow groundwater of the PRE reported by Zhang et al. (2011), but their extremely high concentrations up to >40 × 10^–12 g/g_water reported at several sites were not observed here. The Station HM in the north of the Lingdingyang Bay has the maximum [⁹Be]_diss of both surface and bottom water. The [⁹Be]_diss of the four outlets in the northern part of the Lingdingyang (i.e., HM, JM, HQ and HE) generally decreases from north to south, the same trend as in estuarine water from L1 to L4. The other four western outlets sit in the Modaomen and Huangmaohai Bay (i.e., MD, JT, HT and YM) have lower [⁹Be]_diss. The [⁹Be]_diss in the surface water of Lingdingyang Bay is comparable to the previous work of Kusakabe et al. (1991) in the salinity range of <18.0, and much higher than those in shelf seawaters of the SCS (S > 25). Compared with the riverine [⁹Be]_diss data worldwide, the [⁹Be]_diss of eight outlets are generally higher than those of Ganges-Brahmaputra estuarine water (Brown et al., 1992b), but are one magnitude lower than those in acidic Orinoco and Amazon basins (Measures and Edmond, 1983; Brown et al., 1992a).

TABLE 1

Table 1. Dissolved beryllium isotope concentrations and ¹⁰Be/⁹Be ratios of all stations.

The [¹⁰Be]_diss range between 137 and 1,194 at/g_water, and have a similar distribution pattern as the [⁹Be]_diss with higher concentrations in the bottom waters of Stations HM and L1. The [¹⁰Be]_diss in the Lingdingyang Bay are also comparable to those of Kusakabe et al. (1991), with [¹⁰Be]_diss of 550 at/g_water in the Lingdingyang except for an abnormally high [¹⁰Be]_diss value of 22,060 at/g_water. The bottom enrichment characters of [⁹Be]_diss and [¹⁰Be]_diss are similar to those previously observed by Yang et al. (2003) near the Yangzi River estuary. For all stations except JT in the Modaomen Bay, the [¹⁰Be/⁹Be]_diss vary from 1.08 to 3.11 × 10^–9, while Station JT has the highest [¹⁰Be/⁹Be]_diss of 5.80 × 10^–9 due to its lowest [⁹Be]_diss value. The inconsistent beryllium isotope distributions of eight outlets reflect their distinct weathering environments of source areas, or impact of potential local anthropogenic contribution.

Reactive ⁹Be and ¹⁰Be Concentrations

The concentrations of ⁹Be and ¹⁰Be, as well as ¹⁰Be/⁹Be ratios in the reactive phases (sum of amorphous and crystalline oxides) are listed in Table 2. The subscripts of am-ox, x-ox, and reac refer to the fractions of amorphous oxides, crystalline oxides and reactive phase, respectively. The amount of suspended sediment collected from surface waters of Stations L3, L4, HE, and MD were not sufficient for leaching (<0.1 g), and the am-ox leachate of L2 for ¹⁰Be measurements was unavailable due to an incidental sample loss during the preparation. The ⁹Be concentrations in the reactive phases ([⁹Be]_reac) of all samples range between 1.26 and 2.25 μg/g, higher than the background value of soils in Guangdong Province (CEMS (China Environmental Monitoring Station), 1990). The mean of [⁹Be]_am–ox (1.11 ± 0.14 μg/g) accounts for, on average, 62% of the [⁹Be]_reac (1.78 ± 0.25 μg/g). The ¹⁰Be concentrations in the reactive phases ([¹⁰Be]_reac) range from 1.37 to 2.97 × 10⁸ at/g, with [¹⁰Be]_am–ox taking up 67% of the [¹⁰Be]_reac on average. The [¹⁰Be]_reac is slightly higher than the particulate ¹⁰Be concentration of 0.25–2.48 × 10⁸ at/g reported by Kusakabe et al. (1991) in the PRE (Locations 30 to 34). The (¹⁰Be/⁹Be)_reac range between 1.25 and 2.28 × 10^–9, and the average (¹⁰Be/⁹Be)_am–ox (2.07 × 10^–9) is ∼21% higher than the (¹⁰Be/⁹Be)_x–ox (1.71 × 10^–9). The exceptions are that (¹⁰Be/⁹Be)_x–ox at JM and HQ are observed slightly higher than their (¹⁰Be/⁹Be)_am–ox.

TABLE 2

Table 2. Beryllium isotope concentrations and ¹⁰Be/⁹Be ratios in reactive phases of suspended sediments.

Among the three sub-estuaries of the PR, the Lingdingyang Bay generally has higher [⁹Be]_reac and [¹⁰Be]_reac, but lower (¹⁰Be/⁹Be)_reac ratios. The [⁹Be]_reac and [¹⁰Be]_reac of four outlet stations in the northern part of the Lingdingyang (i.e., HM, JM, HQ, and HE) show marked differences. The average [⁹Be]_reac of Stations HM and JM is ∼18% higher than the average value of all samples, while the corresponding [¹⁰Be]_reac only exceeds ∼4%, giving rise to distinctively low (¹⁰Be/⁹Be)_reac ratios. Station HQ has the highest [¹⁰Be]_reac among all outlets, as well as a (¹⁰Be/⁹Be)_reac ratio much higher than those of HM and JM. As an exception in the Lingdingyang, the [⁹Be]_reac, [¹⁰Be]_reac, and (¹⁰Be/⁹Be)_reac of Station HE are all lower than the average values. From Stations L1–L4 (with increasing salinity), the [⁹Be]_reac and [¹⁰Be]_reac in the bottom suspended sediments decreased by 15 and 10%, respectively.

Comparing to the northern outlets in the Lingdingyang Bay, the southern four outlets in the Modaomen and Huangmaohai Bay have generally lower [⁹Be]_reac and [¹⁰Be]_reac, resulting in higher (¹⁰Be/⁹Be)_reac ratios. There is one exception that Station HT in the Huangmaohai Bay has distinctively lower (¹⁰Be/⁹Be)_reac ratios both in surface and bottom suspended sediments. Just a few kilometers to the west of Station HT, Station YM has an average (¹⁰Be/⁹Be)_reac ratio of 2.23 × 10^–9, ∼32% higher than that of HT. It should be pointed out that YM is also an outlet of the Tanjiang River, rising from west Guangdong Province and also belonging to the PR system.

Concentrations of Cu, Cd, and Pb in Dissolved and Reactive Phases

The concentrations of Cu, Cd, and Pb in both dissolved and reactive phases are summarized in Table 3. The dissolved heavy metal concentrations of all water samples range from 0.379 to 8.75 μg/L for Cu, 0.0165–0.0567 μg/L for Cd, and 0.0353–0.798 μg/L for Pb. Each of the heavy metals show significant variation in spatial distribution. The [Cu]_diss are comparable to the data collected during 2006–2012 by Zhen et al. (2016), but our [Cd]_diss and [Pb]_diss are lower than theirs. The reactive heavy metal concentrations of all suspended sediments are in the range of 46.8–910 μg/g for Cu, 0.0465–2.16 μg/g for Cd, and 49.2–155 μg/g for Pb. All of the reactive heavy metal concentrations are significantly higher than background of soils in Guangdong Province (MEP (Ministry of Environmental Protection of the People’s Republic of China), 2009) and the surface sediments from the shelf areas of the SCS (Zhang and Du, 2005), although these background values are based on total digestions. Compared with previous studies around the PRE (Yang et al., 2012; Ye et al., 2012; Zhuang et al., 2018), our [Cu]_reac, [Cd]_reac and [Pb]_reac at several outlet stations are at extremely high levels, despite that Woods et al. (2012) reported an average Cd concentration up to 1.9 μg/g in surface sediments from the PRE.

TABLE 3

Table 3. Concentrations of Cu, Cd, Pb in dissolved and reactive phases of all samples.

Similar to ⁹Be, the higher heavy metal concentrations both in the dissolved and reactive phases are found at Station HM and JM. The highest [Cu]_diss and [Cd]_diss occur in the bottom water of HM, while the bottom suspended sediment of JM contains the highest [Cd]_reac and [Pb]_reac. However, the extremely high heavy metal concentrations are not detected in the northern Lingdingyang (L1 and L2). Generally, heavy metal concentrations of estuarine waters are lower than those in the outlet stations. The decreasing trend of the dissolved heavy metal concentrations is also observed from Station L1–L4 with increasing salinity.

The highest [Cu]_reac is found in the bottom suspended sediment of Station JT, four times higher than the average value. Moreover, the [Pb]_diss and [Pb]_reac of the bottom JT are also among the highest levels. Station MD, another outlet in the Modaomen Bay, has high [Cd]_diss and [Cd]_reac. The two stations in the Huangmaohai Bay, HT and YM, have high [Cd]_diss but low [Cu]_reac, with Station YM containing the highest [Pb]_diss among all stations.

Apart from the heavy metals, the concentrations of reactive Fe are also analyzed to derive the metal/Fe ratio. Because the anthropogenic source of Fe is negligible compared to its high abundance in nature (Zhou et al., 2007), Fe is widely used as a reference element in calculation of EFs (Reimann and Caritat, 2000). The [Fe]_reac is nearly constant among all samples, with values of most samples around 46 mg/g. Hence, it is appropriate to use Fe as a normalizing element in our study.

Discussion

Partition of Beryllium Isotopes Between Amorphous Oxides and Dissolved Phases

Wittmann et al. (2012) reported that the amorphous oxides of suspended sediments carried 47–74% of ¹⁰Be and 15–27% of ⁹Be relative to the total detrital Amazon River sediments. Compared to the crystalline oxides which record the ¹⁰Be/⁹Be from upper reaches of the river, the amorphous oxides phases have higher specific surface areas and exchange with dissolved phase more actively (Wittmann et al., 2015). Here we calculate distribution coefficient (K_d), as well as partition ratio (k) between amorphous oxides of suspended sediments and dissolved phases, to present proportions of beryllium isotopes carried by suspended sediments. In a certain volume of water, the partition ratio represents the ratio of total amounts of ¹⁰Be (or ⁹Be) carried by suspended sediment and water:

$$k{({}^{10}\text{Be})}_{\mathrm{am}-\mathrm{ox}}=\frac{m{({}^{10}Be)}_{am-ox}}{m{({}^{10}Be)}_{diss}}=K_d{({}^{10}Be)}_{am-ox}=\frac{m_{sediment}}{m_{water}}(3)$$

where m(¹⁰Be)_am–ox and m(¹⁰Be)_diss represent total amounts of ¹⁰Be in amorphous oxides and dissolved phases, C(¹⁰Be)_am–ox and C(¹⁰Be)_diss are their concentrations in corresponding phases, and m_sediment/m_water is the suspended load calculated from mass of dried filter and sampled water. The K_d (¹⁰Be)_am–ox is calculated as a ratio of ¹⁰Be concentrations in sediments (amorphous oxide phase) and water (You et al., 1989; Aldahan et al., 1999). Thus, both the k of ¹⁰Be and ⁹Be are calculated. A higher k(¹⁰Be)_am–ox means that the suspended sediments carried more ¹⁰Be relative to dissolved phase. Since the concentrations of ¹⁰Be and ⁹Be in amorphous oxides have minor variations, the k(¹⁰Be)_am–ox and k(⁹Be)_am–ox mainly depend on their dissolved concentrations and the suspended load.

The distribution coefficients K_d (⁹Be)_am–ox and K_d (¹⁰Be)_am–ox are site dependent, ranging from (1.56–10.8) × 10⁵ and (1.32–9.76) × 10⁵ g_water/g, respectively. Both of them are within the limits of K_d values at neutral and alkaline conditions observed by You et al. (1989) in the laboratory batch experiments. The bottom water of Stations L1, L3, and HM with high suspended load is corresponding to low K_d values (Figure 2). Most of the calculated partition ratios k(¹⁰Be)_am–ox and k(⁹Be)_am–ox are in the range from 3.00–33.5 to 2.98–30.4, with average values of 21.1 and 20.1, respectively. Therefore, only 5% of ¹⁰Be occurs in the water as dissolved form, much less than those in form of amorphous oxide. The extremely high k(¹⁰Be)_am–ox and k(⁹Be)_am–ox, (71.7 and 116) for the bottom water of Station L3 are attributed to the sampling depth that was too close to the river bed and caused a higher suspended load. As shown in Figure 3A, the k(¹⁰Be)_am–ox – k(⁹Be)_am–ox relationship is close to the 1:1 line on log-log coordinates. Excluding the extremely high value of L3, the k(¹⁰Be)_am–ox is, on average, 20% higher than the corresponding k(⁹Be)_am–ox.

FIGURE 2

Figure 2. Partition coefficient (K_d) of ⁹Be and ¹⁰Be versus suspended load. Three bottom waters (HM, L1, and L3) with high suspended load and low K_d values are marked.

FIGURE 3

Figure 3. (A) The partition ratio (k) of ⁹Be and ¹⁰Be between amorphous oxides and dissolved phases. The blue dash line represents 1:1 line. (B) The k-pH plot of both ⁹Be (black squares) and ¹⁰Be (blue dots), the data of HM with pH < 7 are not shown. The extremely high values of L3 were excluded in these plots due to suspended load bias.

Similar to K_d, the k(¹⁰Be)_am–ox and k(⁹Be)_am–ox generally increase with pH at pH > 7 (Figure 3B), with exception of HM. The lowest k(¹⁰Be)_am–ox and k(⁹Be)_am–ox were found at Stations JM and HM, where the river waters are too clear to obtain enough suspended load and precise beryllium concentrations in reactive phase. From Station L1 to Station L4, the k(⁹Be)_am–ox show minor variations and k(¹⁰Be)_am–ox increases ∼50%, despite the greater increase in pH and salinity. Indeed, the K_d of ¹⁰Be and ⁹Be in bottom water have increased by three times from L1 to L4. The reason for decreased amplitude of k value is that the concentration of suspended sediments of L4 is dropped to ∼1/4 compared with L1. Thus, the k instead of K_d could reflect the capacity of suspended sediments in transporting beryllium isotopes more directly.

Since most of the suspended sediments are trapped in the estuary, the k(¹⁰Be)_am–ox from river outlets are higher than the actual percentage of particulate ¹⁰Be that transport from river to seawater. Moreover, the exchange between suspended sediments and dissolved phases cannot reach equilibrium in several weeks (Li et al., 1984), which may introduce large error to the ¹⁰Be budget estimation based on K_d or k.

Export Flux of Riverine ¹⁰Be Into Estuarine Water

At steady state, the ¹⁰Be flux transported by river water into estuary and the SCS should be equal to the deposition flux of meteoric ¹⁰Be that the river basin received. We apply the approach of Wittmann et al. (2015) which was based on framework of von Blanckenburg et al. (2012) to estimate the balance of ¹⁰Be flux between atmospheric deposition and riverine export. Based on the assumption of constant concentrations, discharges and sediment loads, a very rough estimation of ¹⁰Be export flux can be achieved. The equations to calculate total fluxes of atmospheric input ($J_{\text{atm}}^{{}^{10}\text{B}\text{e}}$) and riverine exported (${\text{J}}_{\text{riv}}^{{}^{10}\text{B}\text{e}}$) of ¹⁰Be are listed as follows:

$$J_{atm}^{{}^{10}Be}=A_{riv}\times F_{met}^{{}^{10}Be}(4)$$

$$J_{riv}^{{}^{10}Be}=(J_{riv\mathrm{\_}reac}^{{}^{10}Be}+J_{riv\mathrm{\_}diss}^{{}^{10}Be})\times (1-ex{p}^{(-\mathrm{\lambda }t)})(5)$$

where A_riv is the area of the PR basin, about 4.50 × 10¹¹ m²; $F_{\text{met}}^{{}^{10}\text{B}\text{e}}$ is the meteoric flux of ¹⁰Be, which is derived from ECHAM5-HAM general circulation model as 250 at/m²/s and equals to 7.89 × 10⁹ at/m²/year (Heikkilä et al., 2013); λ is the decay constant of ¹⁰Be, while t is the storage time of sediments. If the storage time is short enough, the right term of equation (5) is close to 1. $J_{\text{riv}\text{\_}\text{r}\text{eac}}^{{}^{10}\text{B}\text{e}}$ and $J_{\text{riv}\text{\_}\text{d}\text{iss}}^{{}^{10}\text{B}\text{e}}$ are fractions of ¹⁰Be carried by reactive and dissolved phases. The $J_{\text{riv}\text{\_}\text{d}\text{iss}}^{{}^{10}\text{B}\text{e}}$ are calculated from the average annual discharges of the PR through each outlets (Q_i) and the corresponding [¹⁰Be]_diss. The $J_{\text{riv}\text{\_}\text{r}\text{eac}}^{{}^{10}\text{B}\text{e}}$ is calculated from [¹⁰Be]_reac and suspended load, therefore the ¹⁰Be fractions in crystalline oxides are also included. Thus, the equation (5) is transformed as sum of eight outlets to equation (6).

$$J_{riv}^{{}^{10}Be}={\mathrm{\Sigma }}_{i=1}^n[({[{}^{10}Be]}_{reac}\times \frac{m_{sediment}}{m_{water}}+{[{}^{10}Be]}_{diss})\times Q_i](6)$$

For each outlet, the [¹⁰Be]_diss and suspended load of surface and bottom water are averaged, while the [¹⁰Be]_reac are weighted average on suspended load of both layers. The [¹⁰Be]_reac of Stations HE and MD are adopted from their bottom values. All of the Q_i values are adopted from Dai et al. (2014). The results of $J_{\mathrm{riv},\mathrm{i}}^{{}^{10}\text{B}\text{e}}$ are listed in Table 4. Hence, the $J_{\text{atm}}^{{}^{10}\text{B}\text{e}}$ and $J_{\text{riv}}^{{}^{10}\text{B}\text{e}}$ calculated using equations (4) and (6) are 3.55 × 10²¹ and 2.74 × 10²¹ at/year, respectively. The exported depositional flux ratio of ¹⁰Be ($J_{\text{riv}}^{{}^{10}\text{B}\text{e}}$/$J_{\text{atm}}^{{}^{10}\text{B}\text{e}}$) is 0.77, that is, 23% of the meteoric ¹⁰Be that is deposited on the PR basin is retained in the basin. Wittmann et al. (2015) pointed out that ¹⁰Be deposited on inactive areas and the decay during sediment storage may also explain the 23% gap between the atmospheric input and export.

TABLE 4

Table 4. The discharge and $J_{\mathrm{riv},\mathrm{i}}^{{}^{10}\text{B}\text{e}}$ from each outlets of the Pearl River and the PRE.

As the major Bay of the PR, the Lingdingyang receives 53% discharge of the PR from HM, JM, HQ, and HE. The riverine ¹⁰Be export from these four outlets is 2.05 × 10²¹ at/year, accounts for 75% of the total riverine export. The data of Station L4 in the southern end of Lingdingyang are adopted to estimate total amounts of ¹⁰Be that escape the Lingdingyang Bay. The discharge of Lingdingyang is the sum from the northern four outlets. Thus, the estimated export ¹⁰Be flux of Lingdingyang is 0.99 × 10²¹ at/year. The ratio of ¹⁰Be flux input into the Lingdingyang Bay and export into the SCS is 49%, i.e., about half of the riverine ¹⁰Be is trapped in the Lingdingyang Bay. Assuming that the Modaomen and Huangmaohai Bay have similar trapped ¹⁰Be proportions as the Lingdingyang, and multiplied by the $J_{\text{riv}}^{{}^{10}\text{B}\text{e}}$/$J_{\text{atm}}^{{}^{10}\text{B}\text{e}}$ ratio of 0.77, the fraction of ¹⁰Be that export into the SCS from the PRE compared to the corresponding deposition of meteoric ¹⁰Be was 38%, although most riverine ¹⁰Be may be further retained in shelf sediments. Therefore, our results are consistent with the current knowledge that marginal seas and continental margins are major sinks of marine ¹⁰Be (Anderson et al., 1990; Yang et al., 2003).

Contamination Assessment of Beryllium and Heavy Metals

The I_geo–reac values are calculated to evaluate the contamination levels of Be (⁹Be), Cu, Cd, Pb in the PRE. Results of I_geo–reac for all metals in suspended sediments are shown in Figure 4. The order of average I_geo–reac of the four metals is Cd (2.45 ± 1.46) ≥ Cu (2.39 ± 1.21) > Pb (0.471 ± 0.469) > Be (−0.048 ± 0.209). Most stations of the PRE are highly polluted by Cu and Cd, while ⁹Be and Pb belong to uncontaminated to moderately contaminated levels.

FIGURE 4

Figure 4. Geo-accumulation indexes (I_geo–reac) of Be (⁹Be), Cu, Cd, Pb in bottom suspended sediments of the PRE.

The highest I_geo–reac of Cu is found at Station JT (5.16, extremely contaminated), while HM, JM, and HE also have I_geo–reac higher than 3 (heavily to extremely contaminated). Heavily contaminated levels (2 < I_geo–reac ≤ 3) are observed at middle part of the Lingdingyang (L2 and L3). The lowest I_geo–reac value of Cu was found at MD (0.876, moderately contaminated), which sits in the same sub-estuary with JT. This distinct contrast of I_geo–reac values from two adjacent outlets indicate that the contamination of Cu is more likely originated from local anthropogenic sources than from upper reaches. The high contamination levels of Cu at HM and JT was also observed by Liu et al. (2014), who attributed these to industrial discharges, copper mining factories and locally high copper consumption.

The highest contamination level for Cd are observed at JM (4.68) and MD (4.21). Indeed, the I_geo–reac of Cd for all eight outlets are above 2, indicating a widespread contamination of Cd among these outlets. The contamination of Cd is mainly sourced from industrial sewage (electroplating and electronic industries) and runoff from agriculture sites in South China (Cheung et al., 2003; Zhuang et al., 2016). The I_geo–reac values of Cd for most estuarine stations in the Lingdingyang are below 1, thus the majority pollution of Cd is not delivered to estuarine waters.

Compared to Cu and Cd, Pb is less polluted in the PRE, with highest I_geo–reac found at JM (1.52) and JT (1.39). The I_geo–reac values of Pb for the rest stations are all below 1, belonging to moderately contaminated conditions. Unlike heavy metals, all I_geo–reac values of ⁹Be are around zero, and no high outliers are found. Thus, the PRE area is generally uncontaminated to moderately contaminated by anthropogenic sources of ⁹Be.

In order to differentiate between natural and potential anthropogenic sources of ⁹Be and heavy metals, the EF values are calculated and shown in Figure 5. We classify the EF values into the following three levels based on criteria adopted by Gu et al. (2012, 2018): mainly natural source (EF < 0.5); natural or anthropogenic sources (0.5 ≤ EF ≤ 1.5); mainly anthropogenic source (EF > 1.5). For the four investigated metals, the ranking of EF values in the PRE is Cd (7.77 ± 6.19) ≥ Cu (7.16 ± 7.57) > Pb (1.34 ± 0.46) > Be (0.896 ± 0.124). Accordingly, Cd and Cu mainly come from anthropogenic pollutions, while Pb and ⁹Be are at the natural or anthropogenic sources level. The spatial distribution of EF of ⁹Be is similar to that of I_geo–reac, and there are no outliers with distinct high EF of ⁹Be among all stations.

FIGURE 5

Figure 5. (A–D) Enrichment factors (EF) of Be (⁹Be), Cu, Cd, Pb in bottom suspended sediments of the PRE.

In summary, the ⁹Be in the PRE are mainly originated from natural sources, none of these investigated stations show evident anthropogenic contaminations. The case of Pb is most similar to ⁹Be, with anthropogenic sources observed at only two stations. Unlike ⁹Be and Pb, the Cd and Cu are widely enriched in suspended sediments, and have extremely high level of contamination at several stations. These results confirm that anthropogenic source of ⁹Be are negligible in the PRE area, hence the beryllium isotopes are eligible in various studies of Earth surface processes.

Conclusion

In this study, distribution of beryllium isotopes in both dissolved and reactive phases of suspended sediments from the PRE are presented. The [⁹Be]_diss and [¹⁰Be]_diss range between 0.78–8.31 × 10^–12 g/g_water and 137–1,194 at/g_water, respectively. We calculated the distribution coefficients (K_d) and partition ratio (k) of ¹⁰Be and ⁹Be between amorphous oxides phase of suspended sediments and the dissolved form. The K_d of both isotopes are at 10⁵ level, and low K_d values are found in three bottom waters with high suspended load. The average k values of ¹⁰Be and ⁹Be are 21.1 and 20.1, respectively. This means that only ∼5% of ¹⁰Be exist in the water as dissolved form, in contrast to the vast inventory of amorphous oxides. In spite of the intense removal of dissolved ¹⁰Be and ⁹Be in the estuary, their proportions in the water and suspended sediments only show minor fluctuations.

The total fluxes of atmospheric meteoric ¹⁰Be deposited on the PR basin, and the ¹⁰Be fluxes export through eight outlets were estimated as 3.55 × 10²¹ and 2.74 × 10²¹ at/year. Thus, 23% of deposited atmospheric ¹⁰Be are retained in the river basin, and only 38% of deposited atmospheric ¹⁰Be could escape the PRE and be transported to the SCS. The calculated I_geo–reac and EF suggest that the PRE area is highly polluted by Cd and Cu, less polluted by Pb, while Be is hardly sourced from human activities. Therefore, our results demonstrate that the ¹⁰Be/⁹Be ratios in the PRE and the northern SCS are mainly controlled by natural processes, and are thus suitable for geoscience studies such as denudation rates by meteoric ¹⁰Be and tracing water masses by (¹⁰Be/⁹Be)_seawater.

Data Availability Statement

The original contributions presented in the study are included in the article/supplementary material, further inquiries can be directed to the corresponding author.

Author Contributions

WK and LZ designed the research and undertaken data analysis and manuscript writing. WK performed the sampling and experiments. GA, DB, and KK responsible for the ¹⁰Be measurements. All authors contributed to the article and approved the submitted version.

Funding

This work was funded by the National Natural Science Foundation of China (No 41776094).

Conflict of Interest

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

Acknowledgments

We thank Xianke Huang for his help during the fieldwork.

References

Aldahan, A., Ye, H. P., and Possnert, G. (1999). Distribution of beryllium between solution and minerals (biotite and albite) under atmospheric conditions and variable pH. Chem. Geol. 156, 209–229. doi: 10.1016/s0009-2541(98)00186-7

CrossRef Full Text | Google Scholar

Anderson, R. F., Lao, Y., Broecker, W. S., Trumbore, S. E., Hofmann, H. J., and Wolfli, W. (1990). Boundary scavenging in the pacific ocean: a comparison of 10Be and 231Pa. Earth Planetary Sci. Lett. 96, 287–304.

Google Scholar

Balls, P. W. (1989). The partition of trace metals between dissolved and particulate phases in european coastal waters: a compilation of field data and comparison with laboratory studies, Netherlands. J. Sea Res. 23, 7–14. doi: 10.1016/0077-7579(89)90037-9

CrossRef Full Text | Google Scholar

Boschi, V., and Willenbring, J. K. (2016a). Beryllium desorption from minerals and organic ligands over time. Chem. Geol. 439, 52–58.

Google Scholar

Boschi, V., and Willenbring, J. K. (2016b). The effect of pH, organic ligand chemistry and mineralogy on the sorption of beryllium over time. Environ. Chem. 13, 711–722. doi: 10.1071/EN15107

CrossRef Full Text | Google Scholar

Bourlès, D., Raisbeck, G. M., and Yiou, F. (1989). 10Be and 9Be in marine-sediments and their potential for dating. Geochim. Et Cosmochim. Acta 53, 443–452. doi: 10.1016/0016-7037(89)90395-5

CrossRef Full Text | Google Scholar

Braucher, R., Guillou, V., Bourlès, D. L., Arnold, M., Aumaître, G., Keddadouche, K., et al. (2015). Preparation of ASTER in-house 10Be/9Be standard solutions. Nucl. Instr. Methods Phys. Res. Section B Beam Interact. With Mater. Atoms 361 (Suppl. C), 335–340. doi: 10.1016/j.nimb.2015.06.012

CrossRef Full Text | Google Scholar

Brown, E. T., Edmond, J. M., Raisbeck, G. M., Bourlès, D. L., Yiou, F., and Measures, C. I. (1992a). Beryllium isotope geochemistry in tropical river basins. Geochim. Et Cosmochim. Acta 56, 1607–1624. doi: 10.1016/0016-7037(92)90228-B

CrossRef Full Text | Google Scholar

Brown, E. T., Measures, C. I., Edmond, J. M., Bourlès, D. L., Raisbeck, G. M., and Yiou, F. (1992b). Continental inputs of beryllium to the oceans. Earth Planetary Sci. Lett. 114, 101–111. doi: 10.1016/0012-821X(92)90154-N

CrossRef Full Text | Google Scholar

Cai, W. J., Guo, X., Chen, C. T. A., Dai, M., and Wang, Y. (2008). A comparative overview of weathering intensity and HCO3-flux in the world’s major rivers with emphasis on the Changjiang, Huanghe, Zhujiang (Pearl) and Mississippi Rivers. Cont. Shelf Res. 28, 1538–1549. doi: 10.1016/j.csr.2007.10.014

CrossRef Full Text | Google Scholar

CEMS (China Environmental Monitoring Station) (1990). Natural Background Values of Soil Elements in China. Beijing: China Environmental Science Press.

Google Scholar

Cheung, K. C., Poon, B. H. T., Lan, C. Y., and Wong, M. H. (2003). Assessment of metal and nutrient concentrations in river water and sediment collected from the cities in the pearl river delta, South China. Chemosphere 52, 1431–1440. doi: 10.1016/S0045-6535(03)00479-X

CrossRef Full Text | Google Scholar

Christl, M., Strobl, C., and Mangini, A. (2003). Beryllium-10 in deep-sea sediments: a tracer for the Earth’s magnetic field intensity during the last 200,000 years. Quater. Sci. Rev. 22, 725–739. doi: 10.1016/s0277-3791(02)00195-6

CrossRef Full Text | Google Scholar

Dai, A., Qian, T., Trenberth, K. E., and Milliman, J. D. (2009). Changes in continental freshwater discharge from 1948 to 2004. J. Clim. 22:2773. doi: 10.1175/2008jcli2592.1

CrossRef Full Text | Google Scholar

Dai, M., Gan, J., Han, A., Kung, H., and Yin, Z. (2014). “Physical dynamics and biogeochemistry of the pearl river plume,” in Biogeochemical Dynamics at Major River-Coastal Interfaces: Linkages With Global Change, eds T. S. Bianchi, M. A. Allison, and W. Jun Cai (Cambridge: Cambridge University Press), 321–352. doi: 10.1017/cbo9781139136853.017

CrossRef Full Text | Google Scholar

Frank, M., Schwarz, B., Baumann, S., Kubik, P. W., Suter, M., and Mangini, A. (1997). A 200 kyr record of cosmogenic radionuclide production rate and geomagnetic field intensity from 10Be in globally stacked deep-sea sediments. Earth Planetary Sci. Lett. 149, 121–129. doi: 10.1016/s0012-821x(97)00070-8

CrossRef Full Text | Google Scholar

Gu, Y.-G., Wang, L.-G., and Gao, Y.-P. (2018). Beryllium in riverine/estuarine sediments from a typical aquaculture wetland, China: Bioavailability and probabilistic ecological risk. Mari. Pollut. Bull. 137, 549–554. doi: 10.1016/j.marpolbul.2018.11.001

PubMed Abstract | CrossRef Full Text | Google Scholar

Gu, Y.-G., Wang, Z.-H., Lu, S.-H., Jiang, S.-J., Mu, D.-H., and Shu, Y.-H. (2012). Multivariate statistical and GIS-based approach to identify source of anthropogenic impacts on metallic elements in sediments from the mid Guangdong coasts China. Environ. Pollut. 163, 248–255. doi: 10.1016/j.envpol.2011.12.041

PubMed Abstract | CrossRef Full Text | Google Scholar

Hanif, N., Eqani, S. A. M. A. S., Ali, S. M., Cincinelli, A., Ali, N. I., Katsoyiannis, A., et al. (2016). Geo-accumulation and enrichment of trace metals in sediments and their associated risks in the Chenab River Pakistan. J. Geochem. Exp. 165, 62–70. doi: 10.1016/j.gexplo.2016.02.006

CrossRef Full Text | Google Scholar

Heikkilä, U., Beer, J., Abreu, J. A., and Steinhilber, F. (2013). On the atmospheric transport and deposition of the cosmogenic radionuclides (10Be). Rev. Space Sci. Rev. 176, 321–332. doi: 10.1007/s11214-011-9838-0

CrossRef Full Text | Google Scholar

Kusakabe, M., Ku, T., Southon, J., Liu, S., Vogel, J., Nelson, D., et al. (1991). Be isotopes in rivers/estuaries and their oceanic budgets. Earth Planetary Sci. Lett. 102, 265–276. doi: 10.1016/0012-821x(91)90022-a

CrossRef Full Text | Google Scholar

Lal, D., and Peters, B. (1967). “Cosmic ray produced radioactivity on the earth,” in Kosmische Strahlung, ed. C. Rays (Berlin: Springer), 551–612. doi: 10.1007/978-3-642-46079-1_7

CrossRef Full Text | Google Scholar

Li, Y.-H., Burkhardt, L., Buchholtz, M., O’Hara, P., and Santschi, P. H. (1984). Partition of radiotracers between suspended particles and seawater. Geochim. Et Cosmochim. Acta 48, 2011–2019. doi: 10.1016/0016-7037(84)90382-x

CrossRef Full Text | Google Scholar

Liu, S. Y., Zhang, Y., Liang, Y. J., Liu, X. Y., Zhao, Y. L., Li, Y., et al. (2014). Spatial-temporal variability and the influencing factors of dissolved copper in riverine runoff at eight outlets in the pearl river estuary Asian. J. Ecotoxicol. 9, 657–662.

Google Scholar

Ma, Q., Hu, M., Zhu, T., Liu, L., and Dai, M. (2005). Seawater, atmospheric dimethylsulfide and aerosol ions in the pearl river estuary and the adjacent northern South China Sea. J. Sea Res. 53, 131–145. doi: 10.1016/j.seares.2004.06.002

CrossRef Full Text | Google Scholar

Measures, C. I., and Edmond, J. M. (1983). The geochemical cycle of 9Be: a reconnaissance. Earth Planetary Sci. Lett. 66, 101–110. doi: 10.1016/0012-821x(83)90129-2

CrossRef Full Text | Google Scholar

MEP (Ministry of Environmental Protection of the People’s Republic of China) (2009). Background Value of Soil Environment in China. Beijing: China Environmental Monitoring Press.

Google Scholar

Müller, G. (1969). Index of geoaccumulation in sediments of the Rhine River. Geojournal 2, 108–118.

Google Scholar

Ni, H.-G., Lu, F.-H., Luo, X.-L., Tian, H.-Y., Wang, J.-Z., Guan, Y.-F., et al. (2008). Assessment of sampling designs to measure riverine fluxes from the pearl river delta, China to the South China Sea. Environ. Monit. Assess. 143, 291–301. doi: 10.1007/s10661-007-9982-x

PubMed Abstract | CrossRef Full Text | Google Scholar

Pan, J., Lai, W., and Devlin, A. T. (2020). “Circulations in the pearl river estuary: observation and modeling,” in Estuaries and Coastal Zones-Dynamics and Response to Environmental Changes, eds J. Pan, W. Lai, and A. T. Devlin (IntechOpen).

Google Scholar

Poulton, S. W., and Canfield, D. E. (2005). Development of a sequential extraction procedure for iron: implications for iron partitioning in continentally derived particulates. Chem. Geol. 214, 209–221. doi: 10.1016/j.chemgeo.2004.09.003

CrossRef Full Text | Google Scholar

Rahn, K. A., and McCaffrey, R. J. (1979). Compositional differences between arctic aerosol and snow. Nature 280, 479–480. doi: 10.1038/280479a0

CrossRef Full Text | Google Scholar

Reimann, C., and Caritat, P. D. (2000). Intrinsic flaws of element enrichment factors (EFs) in environmental geochemistry. Environ. Sci. Technol. 34, 5084–5091. doi: 10.1021/es001339o

CrossRef Full Text | Google Scholar

Shah, A. N., Tanveer, M., Hussain, S., and Yang, G. (2016). Beryllium in the environment: whether fatal for plant growth? Rev. Environ. Sci. Bio. Technol. 15, 549–561. doi: 10.1007/s11157-016-9412-z

CrossRef Full Text | Google Scholar

Simon, Q., Bourlès, D. L., Thouveny, N., Horng, C.-S., Valet, J.-P., Bassinot, F., et al. (2018). Cosmogenic signature of geomagnetic reversals and excursions from the réunion event to the matuyama–brunhes transition (0.7–2.14 Ma interval). Earth Planetary Sci. Lett. 482, 510–524. doi: 10.1016/j.epsl.2017.11.021

CrossRef Full Text | Google Scholar

Simon, Q., Suganuma, Y., Okada, M., and Haneda, Y. (2019). High-resolution 10Be and paleomagnetic recording of the last polarity reversal in the Chiba composite section: age and dynamics of the matuyama–brunhes transition. Earth Planetary Sci. Lett. 519, 92–100. doi: 10.1016/j.epsl.2019.05.004

CrossRef Full Text | Google Scholar

Simon, Q., Thouveny, N., Bourlès, D. L., Nuttin, L., Hillaire-Marcel, C., and St-Onge, G. (2016a). Authigenic 10Be/9Be ratios and 10Be-fluxes (230Thxs-normalized) in central baffin bay sediments during the last glacial cycle: paleoenvironmental implications. Quater. Sci. Rev. 140, 142–162. doi: 10.1016/j.quascirev.2016.03.027

CrossRef Full Text | Google Scholar

Simon, Q., Thouveny, N., Bourlès, D. L., Valet, J. P., and Bassinot, F. (2020). Cosmogenic 10Be production records reveal dynamics of geomagnetic dipole moment (GDM) over the Laschamp excursion (20–60ka). Earth Planetary Ence. Lett. 550:116547. doi: 10.1016/j.epsl.2020.116547

CrossRef Full Text | Google Scholar

Simon, Q., Thouveny, N., Bourlès, D. L., Valet, J.-P., Bassinot, F., Ménabréaz, L., et al. (2016b). Authigenic 10Be/9Be ratio signatures of the cosmogenic nuclide production linked to geomagnetic dipole moment variation since the brunhes/matuyama boundary. J. Geophy. Res. Solid Earth 121, 7716–7741. doi: 10.1002/2016JB013335

PubMed Abstract | CrossRef Full Text | Google Scholar

Suhrhoff, T. J., Rickli, J., Crocket, K., Bura-Nakic, E., and Vance, D. (2019). Behavior of beryllium in the weathering environment and its delivery to the ocean. Geochim. Et Cosmochim. Acta 265, 48–68. doi: 10.1016/j.gca.2019.08.017

CrossRef Full Text | Google Scholar

Taylor, T. P., Ding, M., Ehler, D. S., Foreman, T. M., Kaszuba, J. P., and Sauer, N. N. (2003). Beryllium in the environment: a review. J. Environ. Sci. Health Part A 38, 439–469.

Google Scholar

Tazoe, H., Yamagata, T., Obata, H., and Nagai, H. (2014). Determination of picomolar beryllium levels in seawater with inductively coupled plasma mass spectrometry following silica-gel preconcentration. Analy. Chim. Acta 852, 74–81. doi: 10.1016/j.aca.2014.09.015

PubMed Abstract | CrossRef Full Text | Google Scholar

Tessier, A., Campbell, P. G., and Bisson, M. (1979). Sequential extraction procedure for the speciation of particulate trace metals. Analy. Chem. 51, 844–851. doi: 10.1021/ac50043a017

CrossRef Full Text | Google Scholar

Valet, J. P., Bassinot, F., Bouilloux, A., Bourles, D., Nomade, S., Guillou, V., et al. (2014). Geomagnetic, cosmogenic and climatic changes across the last geomagnetic reversal from Equatorial Indian Ocean sediments. Earth Planetary Sci. Lett. 397, 67–79. doi: 10.1016/j.epsl.2014.03.053

CrossRef Full Text | Google Scholar

von Blanckenburg, F., and Bouchez, J. (2014). River fluxes to the sea from the ocean’s 10Be/9Be ratio. Earth Planetary Sci. Lett. 387, 34–43. doi: 10.1016/j.epsl.2013.11.004

CrossRef Full Text | Google Scholar

von Blanckenburg, F., Bouchez, J., and Wittmann, H. (2012). Earth surface erosion and weathering from the 10Be (meteoric)/9Be ratio. Earth Planetary Sci. Lett. 351-352, 295–305. doi: 10.1016/j.epsl.2012.07.022

CrossRef Full Text | Google Scholar

Wiederhold, J. G., Teutsch, N., Kraemer, S. M., Halliday, A. N., and Kretzschmar, R. (2007). Iron isotope fractionation in oxic soils by mineral weathering and podzolization. Geochim. Et Cosmochim. Acta 71, 5821–5833. doi: 10.1016/j.gca.2007.07.023

CrossRef Full Text | Google Scholar

Wittmann, H., and von Blanckenburg, F. (2016). The geological significance of cosmogenic nuclides in large lowland river basins. Earth Sci. Rev. 159, 118–141. doi: 10.1016/j.earscirev.2016.06.001

CrossRef Full Text | Google Scholar

Wittmann, H., von Blanckenburg, F., Bouchez, J., Dannhaus, N., Naumann, R., Christl, M., et al. (2012). The dependence of meteoric 10Be concentrations on particle size in Amazon River bed sediment and the extraction of reactive 10Be/9Be ratios. Chem. Geol. 318, 126–138. doi: 10.1016/j.chemgeo.2012.04.031

CrossRef Full Text | Google Scholar

Wittmann, H., von Blanckenburg, F., Dannhaus, N., Bouchez, J., Gaillardet, J., Guyot, J. L., et al. (2015). A test of the cosmogenic 10Be(meteoric)/9Be proxy for simultaneously determining basin-wide erosion rates, denudation rates, and the degree of weathering in the Amazon basin. J. Geophys. Re. Earth Surf. 120, 2498–2528. doi: 10.1002/2015JF003581

CrossRef Full Text | Google Scholar

Woods, A. M., Lloyd, J. M., Zong, Y., and Brodie, C. R. (2012). Spatial mapping of pearl river estuary surface sediment geochemistry: influence of data analysis on environmental interpretation. Estuar. Coastal Shelf Sci. 115, 218–233. doi: 10.1016/j.ecss.2012.09.005

CrossRef Full Text | Google Scholar

Yang, Y. L., Kusakabe, M., and Southon, J. R. (2003). 10Be profiles in the East China sea and the okinawa trough. Deep Sea Res. Part II Top. Stud. Oceanogr. 50, 339–351. doi: 10.1016/s0967-0645(02)00458-7

CrossRef Full Text | Google Scholar

Yang, Y., Chen, F., Zhang, L., Liu, J., Wu, S., and Kang, M. (2012). Comprehensive assessment of heavy metal contamination in sediment of the pearl river estuary and adjacent shelf. Mari. Pollut. Bull. 64, 1947–1955. doi: 10.1016/j.marpolbul.2012.04.024

PubMed Abstract | CrossRef Full Text | Google Scholar

Ye, F., Huang, X., Zhang, D., Tian, L., and Zeng, Y. (2012). Distribution of heavy metals in sediments of the pearl river Estuary, Southern China: implications for sources and historical changes. J. Environ. Sci. 24, 579–588. doi: 10.1016/S1001-0742(11)60783-3

CrossRef Full Text | Google Scholar

You, C. F., Lee, T., and Li, Y.-H. (1989). The partition of Be between soil and water. Chem. Geol. 77, 105–118. doi: 10.1016/0009-2541(89)90136-8

CrossRef Full Text | Google Scholar

Zhang, W., Shousheng, M., Zhang, Y., and Kaimin, C. (2011). Temporal variation of suspended sediment load in the Pearl River due to human activities. Int. J. Sediment Res. 26, 487–497. doi: 10.1016/s1001-6279(12)60007-9

CrossRef Full Text | Google Scholar

Zhang, Y., and Du, J. (2005). Background values of pollutants in sediments of the South China Sea. Acta Oceanol. Sin. 27, 161–166.

Google Scholar

Zhen, G., Li, Y., Tong, Y., Yang, L., Zhu, Y., and Zhang, W. (2016). Temporal variation and regional transfer of heavy metals in the Pearl (Zhujiang) River, China. Environ. Sci. Pollut. Res. 23, 8410–8420. doi: 10.1007/s11356-016-6077-7

PubMed Abstract | CrossRef Full Text | Google Scholar

Zhou, F., Guo, H., and Hao, Z. (2007). Spatial distribution of heavy metals in Hong Kong’s marine sediments and their human impacts: a GIS-based chemometric approach. Mari. Pollut. Bull. 54, 1372–1384. doi: 10.1016/j.marpolbul.2007.05.017

PubMed Abstract | CrossRef Full Text | Google Scholar

Zhuang, Q., Li, G., and Liu, Z. (2018). Distribution, source and pollution level of heavy metals in river sediments from South China. Catena 170, 386–396. doi: 10.1016/j.catena.2018.06.037

CrossRef Full Text | Google Scholar

Zhuang, W., Chen, Q., Gao, X., Zhou, F., Wang, M., and Liu, Y. (2016). Characterization of surface sediments from the beijing-hangzhou grand canal (Zaozhuang section). China: assessment of beryllium enrichment, biological effect, and mobility. Environ. Sci. Pollut. Res. 23, 13560–13568. doi: 10.1007/s11356-016-6837-4

PubMed Abstract | CrossRef Full Text | Google Scholar