A new perspective on the 137Cs retention mechanism in surface soils during the early stage after the Fukushima nuclear accident

The Fukushima Daiichi nuclear power plant accident caused serious radiocesium (137Cs) contamination of the soil in multiple terrestrial ecosystems. Soil is a complex system where minerals, organic matter, and microorganisms interact with each other; therefore, an improved understanding of the interactions of 137Cs with these soil constituents is key to accurately assessing the environmental consequences of the accident. Soil samples were collected from field, orchard, and forest sites in July 2011, separated into three soil fractions with different mineral–organic interaction characteristics using a density fractionation method, and then analyzed for 137Cs content, mineral composition, and organic matter content. The results show that 20–71% of the 137Cs was retained in association with relatively mineral-free, particulate organic matter (POM)-dominant fractions in the orchard and forest surface soil layers. Given the physicochemical and mineralogical properties and the 137Cs extractability of the soils, 137Cs incorporation into the complex structure of POM is likely the main mechanism for 137Cs retention in the surface soil layers. Therefore, our results suggest that a significant fraction of 137Cs is not immediately immobilized by clay minerals and remains potentially mobile and bioavailable in surface layers of organic-rich soils.

The accident at the Fukushima Daiichi nuclear power plant (NPP) in March 2011 released substantial amounts of radionuclides into the atmosphere and consequently caused serious radioactive contamination of terrestrial ecosystems over a wide area of eastern Japan 1 . Of the radionuclides found in these terrestrial ecosystems, 137 Cs, with a physical half-life of 30.1 years, is the primary source of concern because of its potential radiological impact on humans and the ecosystems over the coming decades.
The behavior of 137 Cs in soil after deposition is one of the key factors necessary to evaluate the long-term radiation risks that may be caused by external and internal exposure. It is generally accepted that, once 137 Cs reaches the surface layers of the mineral soil, it is strongly adsorbed (i.e., fixed) by clay minerals 2,3 , rapidly reducing the mobility and bioavailability of 137 Cs in such soils [4][5][6][7][8][9][10] . Konopleva et al. 11 demonstrated that, 19 years after the Chernobyl NPP accident, more than 50% of the Chernobyl-derived 137 Cs remained in the upper 10-cm soil layer in German forest ecosystems. The long-term accumulation of Chernobyl-derived 137 Cs in the surface soil layers has also been observed in grassland ecosystems in European countries 12,13 . In Japan, it has recently been reported that, at a Japanese cedar plantation even 38 years after the fallout from atmospheric nuclear weapons testing, the deposited 137 Cs was still observed, primarily in the uppermost 5 cm of the mineral soil 14 . In addition, there are many observations showing that 137 Cs in mineral soil is more highly concentrated in clay-sized (<2 μm) soil particles than in larger-sized soil particles, reflecting the specific adsorption of 137 Cs onto clay minerals [15][16][17][18][19] . These studies support the overall immobility of 137 Cs in surface soil layers over a long period of time after deposition as a result of the interaction of 137 Cs with clay minerals in the soil layers.
However, there are also studies suggesting that the 137 Cs behavior in the mineral soil cannot simply be evaluated based on the interaction of 137 Cs with clay minerals and there are many factors that can affect the mobility and bioavailability of 137 Cs in soils. These factors, other than the clay mineralogy, include soil physicochemical Cs and adsorbed 137 Cs is easily exchangeable with other cations 20,22 . In addition to direct influences, an indirect influence of soil organic matter on the 137 Cs behavior has also been suggested in that the presence of soil organic matter inhibits the fixation of 137 Cs by clay minerals by limiting the access of 137 Cs to 137 Cs-fixation sites on the clay minerals 22,33 . It has recently been shown that soil aggregation (i.e., the binding of soil particles into aggregates together with soil organic matter) has a significant influence on the mobility and bioavailability of 137 Cs in mineral soils in Japanese forest ecosystems affected by the Fukushima NPP accident 26 . Given that soils (forest soils in particular) in Japan are rich in organic matter 34 and soil organic matter is a dynamic constituent that turns over constantly via litter input and microbial decomposition, an improved understanding of the interactions between 137 Cs, minerals, and organic matter is likely critical to predict the fate of Fukushima-derived 137 Cs in terrestrial ecosystems in Japan.
The aim of the present study is therefore to explore the retention mechanisms of 137 Cs in the surface soil layers of terrestrial ecosystems affected by the Fukushima NPP accident, with a specific focus on the interactions between 137 Cs, soil minerals, and organic matter. We collected surface soil samples from three types of terrestrial ecosystems (a field, CP-2; an orchard, CP-6; and a forest, FR-4) in July 2011 (approximately 3.5 months after the accident) and at the FR-4 site in July 2015 in the city of Fukushima (Fig. 1). The soil samples were separated into three soil fractions with different physical characteristics, a free particulate organic matter (POM)-dominant low-density fraction (fLF); a mineral-associated low-density fraction (mLF); and a mineral-dominant high-density fraction (HF), using a density fractionation method [35][36][37] . The soil fractions were then analyzed for their 137 Cs content, mineral composition, and organic matter (carbon and nitrogen) content.

Results
General properties of the soil samples. Selected physicochemical properties and 137 Cs activity concentrations of the soil samples used in the present study are given in Table 1. The soils at the three sites had similar particle-size distributions. The forest soils (FR-4) were generally higher in organic C content and CEC than the cropland soils. The fractions of 137 Cs extracted with 1 M ammonium acetate (NH 4 Ac, pH 7) from the CP-2 and www.nature.com/scientificreports www.nature.com/scientificreports/ FR-4 soil samples were approximately 10-19% 9 . Radiocesium in this fraction has been considered to be easily exchangeable with other cations such as NH 4 + present in soil solution.
Microscopic observations of the soil physical fractions. The physical fractionation at a density of 1.6 g cm −3 successfully separated the soil samples into three fractions (fLF, mLF, and HF) showing different physical characteristics (Fig. 2). The fLF fraction consisted primarily of plant detritus or POM with recognizable structures. It also appeared that some of the POM in the fLF fraction was coated with fine mineral particles and organic materials. The mLF fraction consisted of finely fragmented organic materials (typically less than 100 μm in size) with less recognizable structures. Some larger-sized plant residues were also observed in the mLF fraction. The HF fraction was dominated by soil mineral particles, some of which appeared to be associated with very fine organic materials.
Mass distributions of the soil physical fractions. In general, the HF fraction accounted for the largest proportion (52-97%) of the soil mass at all sites and depths; however, the mass distribution patterns of the three physical fractions differed between the sites for the different land-use conditions (Fig. 3a). At the field (CP-2) site, the HF fraction made up more than 95% of the soil mass and was independent of the soil depth. At the forest (FR-4) site, the fLF fraction represented a significant proportion (34-38%) of the soil mass in the topmost (0-1 cm) layers and the mass proportion of the fLF fraction decreased with depth to 5-6% in the lowest soil layers. The orchard (CP-6) site also showed a depth-wise trend for the mass distribution similar to that at the FR-4 site; however, the mass proportion of the fLF fraction was smaller than that at the FR-4 site throughout the soil profile. At all sites, the mLF fraction comprised a minor fraction (0.6% up to 7.1%) of the soil mass and showed no consistent decrease or increase in the mass proportion with depth.  (Table 2), the 137 Cs distribution between the physical fractions showed a pattern that was very different from that of the mass distribution between the fractions (Fig. 3b). At the FR-4 site, more than 70% of the 137 Cs was associated with the fLF fraction in the topmost soil layers in both years and the contribution of the fLF to the 137 Cs retention was also remarkable (17-58%) in the underlying soil layers. Similarly, at the CP-6 site, the fLF fraction retained a significant proportion (20-45%) of the 137 Cs in the soil. At the CP-2 site, however, fLF-associated 137 Cs comprised 9% or less of the soil 137 Cs and most of the 137 Cs was contained within the HF fraction. The mLF fraction was a minor (<7%) fraction with respect to the retention of 137 Cs in the soil across all sites.
poM versus mineral particles in the fLF fractions. Via the 137 Cs analysis of the soil physical fractions, it was found that the fLF fractions had the highest 137 Cs activity concentration and consequently retained a significant proportion of the 137 Cs in the soil. In addition, microscopic observations showed that the fLF fractions were composed primarily of POM including plant detritus; however, some of the POM appeared to be coated with fine mineral particles. To explore the mechanisms for the high 137 Cs retention in the fLF fractions, we further separated the fLF fractions into two additional fractions, POM (fLF-POM) and mineral particles (fLF-MP), using a fractionation at a density of 1.9 g cm −3 after ultrasonic washing. www.nature.com/scientificreports www.nature.com/scientificreports/ Even though POM appeared to be the primary component of the fLF fractions in the topmost soil layers ( Table 3, Fig. 4), there was also a non-negligible amount (14-28% on a mass fraction basis) of mineral particles (with a density greater than 1.9 g cm −3 ) in the fractions. At the CP-2 site, the 137 Cs activity concentration was much higher in fLF-MP than in fLF-POM. Conversely, at the FR-4 site, the 137 Cs activity concentration was an order of magnitude greater in fLF-POM than in fLF-MP. At the CP-6 site, the 137 Cs activity concentration was similar in the two fractions. As a result, more than 90% of the fLF-associated 137 Cs was found in fLF-MP for the CP-2 site, whereas for the CP-6 and FR-4 sites, more than 80% of the fLF-associated 137 Cs was found in fLF-POM.    www.nature.com/scientificreports www.nature.com/scientificreports/ Across all sites, the fLF-POM fractions in the topmost soil layers were characterized by a high C content and C/N ratio ( Table 4), indicating that the fractions were rich in relatively fresh, less degraded organic materials 38,39 . Conversely, the fLF-MP fractions generally had low C contents and C/N ratios, indicating that these fractions were dominated by soil mineral materials with a relatively small contribution of more degraded organic materials. These results were consistent with the visual inspection of fLF-POM and fLF-MP (Fig. 4).
The X-ray diffraction (XRD) patterns of the fLF-POM and fLF-MP fractions are given in Fig. 5 and compared to those of the corresponding HF fractions. The major mineral compositions in the soil fractions were quartz, feldspar, and smectite, and these minerals were observed even for the fLF-POM fractions. In general, the mineral compositions were very similar for the three fractions within each site. Minerals specific to radiocesium adsorption (e.g., micaceous clay minerals) were not identified in the fLF-POM fractions at any of the sites.

Discussion
There has long been a general consensus that once 137 Cs is transferred to the mineral soil, it is rapidly and nearly irreversibly fixed by clay minerals 2,40,41 . The fixation of 137 Cs by clay minerals results in limited migration through the soil profile due to chemical and biological processes 10,12-14 and, consequently, a reduced availability for 137 Cs uptake by plants 11,42,43 . According to this consensus, Fukushima-derived 137 Cs in the upper layers of mineral soil is assumed to be largely associated with soil fractions dominated by mineral particles; however, this was not the case in the present study. Most of the 137 Cs was indeed associated with the mineral-dominant HF fraction at the CP-2 site; however, at the CP-6 and FR-4 sites where the fLF fraction had abundantly accumulated, the 137 Cs was largely associated with the fLF fraction and this was also observed more than 4 years after the accident (Fig. 3).
Clearly, the most interesting question here is how such a large amount of 137 Cs can be retained in the fLF fraction in the upper layers of the mineral soil at the fLF-rich cropland (CP-6) and forest (FR-4) sites. Four possible mechanisms merit consideration: (1) nonspecific adsorption by organic materials; (2) incorporation into the POM structure; (3) association with admixed clay minerals; and (4) retention in the soil microbial biomass.
According to mechanism (1), the 137 Cs is largely adsorbed onto nonspecific cation-exchange sites provided by organic materials in the fLF fraction 20,22,44 . The soils at the CP-6 and FR-4 sites are rich in organic materials (organic C) and therefore have a high CEC compared to the CP-2 site (Table 1). However, the fractions of 137 Cs associated with fLF in the soil samples (40-71% of the total 137 Cs for the FR-4 site, Fig. 3b) were much larger than the fractions of 137 Cs extracted with 1 M ammonium acetate (NH 4 Ac, pH 7) from the samples (11-19%, Table 1). This indicates that most of the fLF-associated 137 Cs was present in a form that is not easily exchangeable with other cations, which leads us to reject nonspecific adsorption as the main mechanism for the 137 Cs retention in the fLF fraction.
According to mechanism (2), the 137 Cs is physically trapped in complex structures of degraded POM (Fig. 4) or incorporated into the tissues of fragmented plant debris (dead cedar leaves) in the fLF fraction, which makes the 137 Cs unextractable via the ion-exchange reaction. This mechanism is partly supported by the present result showing that the fLF-associated 137 Cs primarily (81-95%) resides in the fLF-POM fraction that was obtained after ultrasonically washing the fLF fraction to remove the attached mineral particles for the CP-6 and FR-4 sites ( Table 3). This mechanism is also consistent with the finding of Tanaka et al. 45 , who showed that 137 Cs was incorporated into and strongly fixed in the interior tissues of dead cedar leaves that fell on the forest floor prior to the Fukushima NPP accident. Kato et al. 46 showed that, in Japanese cedar forests, the deposition of 137 Cs from the tree canopy to the forest floor occurred primarily via throughfall, not litterfall (i.e., as 137 Cs incorporated into the leaf tissues), during the early stage (<200 days) after the Fukushima NPP accident. Therefore, the incorporation of 137 Cs into the POM structure after deposition to the mineral soil as a dissolved form appears to be a plausible mechanism for 137 Cs retention in the fLF fraction as of July 2011.  Table 3. 137 Cs distribution between the fLF-POM and fLF-MP fractions. a The fraction lost during the separation procedure was not considered when determining the 137 Cs distribution in the fLF fraction. b Errors represent the counting errors in the radiation measurement.  www.nature.com/scientificreports www.nature.com/scientificreports/ According to mechanism (3), the 137 Cs is actually fixed in the clay minerals and the clay minerals are tightly associated with POM 26,47 or rather incorporated into the POM structure. Indeed, the XDR analysis showed that minerals were still present in fLF-POM (Fig. 5). However, the mineral composition of fLF-POM was similar to that of fLF-MP, as well as that of HF, and minerals specific to 137 Cs adsorption (e.g., micaceous minerals) were not found in the fLF-POM fraction. In addition, in the FR-4 soil, the 137 Cs activity concentration of fLF-POM was much higher than those of fLF-MP and HF (Tables 2 and 3). Under these circumstances, this mechanism can only be acceptable if a selective association occurs between POM and 137 Cs-concentrated clay minerals in the soil. However, this is unlikely.
According to mechanism (4), the 137 Cs is retained in the soil microbial biomass in the fLF fraction. Microorganisms in soils have been found to be able to accumulate 137 Cs [48][49][50] . Given that soil microorganisms are generally abundant in organic-rich surface soil layers and that fLF fractions are deficient in clay minerals, which are a strong adsorbent of 137 Cs, it is conceivable that soil microorganisms are responsible for the retention of 137 Cs, particularly in the fLF fraction. Brückmann and Wolters 27 have shown that, on average, 13% (with a large variability between 1% and 56%) of the 137 Cs inventory was immobilized in the microbial biomass in forest-floor organic layers of German forest ecosystems. However, Stemmer et al. 28 reported that the amount of microbial biomass-bound 137 Cs was less than 2% of the 137 Cs inventory in the surface 0-3 cm of mineral soil in an alpine meadow in Austria. Similarly, a minor contribution of the microbial biomass to 137 Cs retention has been reported for subtropical forest soils (0-10-cm depth) in Taiwan 29 . Following the Fukushima NPP accident, it has been shown that the uptake of 137 Cs by soil microorganisms is less important for the retention of 137 Cs in forest surface (0-3 cm) soils in Fukushima compared to ion-exchange adsorption on nonspecific sites provided by abiotic components 8 . In light of these studies, this mechanism may be partly responsible for the 137 Cs retention in the fLF fraction; however, it is clearly not entirely sufficient to explain the observed amount of fLF-associated 137 Cs (20-78% of the total 137 Cs, Fig. 3b) in the soils at the CP-6 and FR-4 sites.
The fLF fraction is relatively less associated with and protected by soil minerals; accordingly, this fraction is considered more easily degradable by soil microorganisms than other fractions [51][52][53] . Therefore, this fraction possibly offers a large but temporary reservoir of 137 Cs in soils. As the microbial decomposition of fLF proceeds, the 137 Cs retained in this fraction can be mobilized, become available to plants, and have the opportunity to be fixed by clay minerals (i.e., to move into the HF fraction) or to migrate downward through the soil profile. In www.nature.com/scientificreports www.nature.com/scientificreports/ the present study, however, the proportion of fLF-associated 137 Cs to the total 137 Cs inventory did not decrease between July 2011 (3.5 months after the accident) and July 2015 (4 years later) at the forest site (Fig. 3b). A similar observation has been reported for surface mineral soils at two forest sites in Fukushima between September 2012 and September 2014 44 . This may be due to the continuous supply of 137 Cs-contaminated POM from the forest-floor organic layer to the mineral soil during the first four years following the accident. Supporting this, studies conducted since the Fukushima NPP accident have shown a rapid (within several years) migration of 137 Cs from organic layers to the mineral soil in Japanese forest ecosystems 10,54-59 . Therefore, it is hypothesized that, at the CP-6 and FR-4 sites, 137 Cs mobilization from the fLF fraction via microbial degradation was compensated for by the 137 Cs supply (as POM) from the organic layers during this period. If this is true, the fLF-associated 137 Cs will decrease with time and with decreasing 137 Cs inventory in the organic layers and the HF-associated 137 Cs will become important in controlling the mobility and bioavailability of 137 Cs in the surface mineral soils.
The mLF fraction had a low 137 Cs activity concentration and a low 137 Cs inventory compared to the other two fractions ( Table 2 and Fig. 3), indicating that finely fragmented organic materials strongly associated with soil minerals (although some might have been occluded in soil aggregates) have a low 137 Cs retention capability. This is in stark contrast with the larger-sized organic materials in the fLF fraction (i.e., fLF-POM, Fig. 4ab), which show a high 137 Cs retention capability, and may partly support the incorporation of 137 Cs into the complex physical structures of larger-sized POM in the fLF fraction (mechanism (2), see above) as the main mechanism for 137 Cs retention in the surface soil layers. Our study demonstrates that, at least until 4 years after the accident, a significant amount of 137 Cs has been retained in larger-sized POM (fLF-POM) in the surface layers of mineral soil in fLF-rich ecosystems. The results of the present study are consistent with the recent finding that a significant proportion (31-55%) of the 137 Cs is present in association with macroaggregates (212-2000 μm) in forest surface soils affected by the Fukushima NPP accident 26 . Given that the 137 Cs currently retained in the fLF fraction may be mobilized and become a source for 137 Cs recycling in such ecosystems in the future, continuous and further investigations of the 137 Cs distribution between the soil physical fractions, especially in fLF-rich ecosystems, are needed to improve our understanding of the 137 Cs retention mechanism in the surface layers of the mineral soil. This is key to accurately assessing the long-term radioecological impacts of the Fukushima NPP accident. physical fractionation of the soil. Dried, sieved (<2 mm), and root-free (visible roots were removed by hand) soil samples were subjected to the following physical fractionation. The fractionation was based on the contrasting densities between soil mineral particles (typically 2.5-3.0 g cm −3 ) and organic materials (<1.4 g cm −3 ) 35,36 . Through fractionation, the soil samples were separated into a relatively mineral-free low-density fraction (fLF), a www.nature.com/scientificreports www.nature.com/scientificreports/ mineral-associated low-density fraction (mLF), and a high-density fraction (HF). This kind of method has been used for studying soil organic matter dynamics in the research fields of soil science and biogeochemistry [35][36][37][51][52][53] .

Methods
The soil samples (approximately 3-5 g) were weighed into 50-ml centrifuge tubes and mixed with 40 ml of a 1.6-g cm −3 sodium polytungstate (SPT) solution. The tubes were gently shaken and centrifuged at 2000 rpm for 30 min, and the floating materials (fLF) were aspirated onto a pre-baked glass microfiber filter (pore size: 0.7 μm). The cycle of gentle shaking, centrifugation, and aspiration was repeated until no fLF remained floating. The fLF fraction was then rinsed with ultrapure water on the filter and dried at 50 °C in an oven. The remaining soil samples were resuspended in 40 ml of 1.6-g cm −3 SPT in the centrifuge tubes, which were then sonicated in an ice bath for 5 min at 70% pulse for a total input of 300 J ml −1 (Sonifier 450, BRANSON, USA); the purpose of this process was to disrupt the aggregates and bonds between the low-density organic materials (mLF) and the high-density mineral particles (HF) in the SPT solution 35,36,53 . The mixture was centrifuged as before. The floating materials (mLF) were very fine and prone to clouding the supernatant, which often caused the filter to clog during aspiration 52,53 . To avoid this difficulty, mLF that was floating on the top of the SPT solution in the tubes was separately collected via pipetting prior to aspiration. The supernatant was then aspirated, and mLF in the supernatant was collected on a pre-baked glass microfiber filter. The cycle of centrifugation, pipetting, and aspiration was repeated until no mLF remained floating. The mLF collected via pipetting (mLF-p) was then rinsed with ultrapure water in a centrifuge tube with repeated centrifugation and dried at 50 °C in an oven. The mLF collected on the filter (mLF-f) was rinsed and dried as before. The mLF-p and mLF-f fractions were mixed to obtain whole mLF samples. After floating off the low-density fractions, the residue (HF) in the centrifuge tubes was rinsed more than five times with ultrapure water via shaking and centrifuging at 2000 rpm for 15 min and then dried at 50 °C in an oven.
The physical fractionation was conducted using three replicates for each of the soil samples, and the physical fractions (fLF, mLF, and HF) obtained from the three replicated samples were mixed together to obtain a sufficient amount of fractionated samples for the 137 Cs analysis. The fractionated samples were photographed using a digital camera (WRYCOM-NF500, WRAYMER, Japan) mounted on a stereomicroscope (SZX7, OLYMPUS, Japan).

Radiocesium analysis.
A radiocesium analysis was conducted following Koarashi et al. 10 . The fractionated soil samples were ground into powder in a mortar, placed into plastic tubes, and analyzed for 137 Cs using a well-type Ge detector (GWL-120, 230, ORTEC, USA). Measurement times were from 2,600 s to 420,000 s, depending on the 137 Cs activity in the samples. The activity concentrations of the 137 Cs were expressed in activity per unit dry weight (Bq kg −1 dw) and corrected for radioactive decay to the sampling date.
The 137 Cs activity in each of the soil physical fractions (fLF, mLF, and HF) was estimated as where A is the 137 Cs activity (Bq) in the fraction, C is the 137 Cs activity concentration (Bq kg −1 ) of the fraction, and M is the mass (kg) of the fraction obtained via physical fractionation. The difference in the soil mass between the unfractionated sample (i.e., prior to fractionation) and the sum of the three physical fractions (i.e., after fractionation) was considered to be the fraction lost during the fractionation procedure and represented <6% of the soil mass. The loss fraction was not analyzed for 137 Cs in the present study, and therefore the amount of 137 Cs in the loss fraction was not considered when evaluating the 137 Cs distribution between the physical fractions of the soil.

Further investigation into the fLF fractions.
To explore the mechanisms for the high 137 Cs retention in the fLF fractions, the fLF fractions were further separated into two fractions (fLF-POM and fLF-MP) via the following procedure. A portion (approximately 0.2 g) of the fLF samples was put into a 50-ml beaker with 20 ml of ultrapure water and was then sonicated for 30 min in a water-bath sonicator (38 kHz, 80 W, tank volume: 1.6 L, US-1, SND Co. Ltd., Japan). The solution was transferred to a 50-ml centrifuge tube and mixed with 20 ml of a 2.8-g cm −3 SPT solution in the tube (i.e., the density of the solution was adjusted to 1.9 g cm −3 ). The tube was shaken well and then centrifuged at 2000 rpm for 30 min. Floating materials in the tube were collected as fLF-POM via pipetting. The remaining supernatant in the tube was discarded and the residue (fLF-MP: mineral particles with a density of >1.9 g cm −3 ) in the tube was rinsed with ultrapure water with repeated shaking and centrifugation. The fLF-POM collected via pipetting was also rinsed with ultrapure water in a tube with repeated centrifugation. Samples of fLF-POM and fLF-MP were dried at 50 °C in an oven, photographed, and analyzed for 137 Cs in the same manner as mentioned above.
The fLF-POM and fLF-MP samples were ground into powder in a mortar, and their mineralogical compositions were measured using an X-ray diffractometer (Ultima IV, Rigaku, Tokyo) with CuKα radiation (40 kV, 20 mA). The XRD pattern was recorded over an angular range of 3-70° (2θ) with a step size of 0.03° and a counting time of 2 s per step. Minerals were identified using the Integrated X-ray powder diffraction software (PDXL Ver.2.4, Rigaku, Tokyo). The corresponding HF samples were also measured for comparison.
Finally, the fLF-POM and fLF-MP samples were analyzed for their total C and N content using an elemental analyzer (vario PYRO cube, Elementar) 39 .