Philippa Kaur

Drainage area and branching ratio: a matter of scaleGeomorphological and ecological viewpoints on river networks generally differ owing to discordant definitions of the fundamental unit (the node) used to analyze them. From a geomorphological perspective, the determination of a river network entails the definition of an observational scale. Real river networks can be extracted from digital elevation models (DEMs) via algorithms for flow direction determination such as D8 (i.e., each pixel drains towards the lowest of its 8 nearest neighbors53). After the outlet location has been specified (and hence the upstream area A spanned by the river network), the first observational scale required is thus the pixel length l of the DEM, which defines the extent of a network node. A second scale is then needed to distinguish the portion of the drainage network effectively belonging to the channel network. The simplest but still widely used method53 defines channels as those pixels whose drainage area exceeds a threshold value AT. Hydrologically based criteria to determine the appropriate value for AT exist54; however, for the sake of simplicity, we here consider AT as a free parameter.BBTs and RBNs are random constructs, and as such they do not satisfy the optimality criterion of minimizing total energy expenditure, which is the fundamental physical process shaping fluvial landscapes. Furthermore, neither of these networks is a spanning tree, which is a key attribute of real fluvial landforms10: in fact, in both BBTs and RBNs, the extent of the drained domain is not defined. As a result, the drainage area at an arbitrary network node cannot in principle be attributed, unless by using the number of upstream nodes as a proxy. This has practical implications from an ecological viewpoint because drainage area is the master variable controlling several attributes of a river, such as width, depth, discharge, or slope3,55, which in turn impact habitat characteristics and the ecology of organisms therein56.In BBTs and RBNs, branching probability p has been defined35,38,45,46,47 as the probability that a network node is branching, i.e. connected to two upstream nodes. As such, the branching probability of a realized river network (be it a real river or a synthetic construct) could be evaluated as the ratio between the number of links NL constituting a network and the total number of network nodes N; if a unit distance between two adjacent nodes is assumed, the denominator equals the total network length. We note that the former definition of branching probability only holds in the context of the generation of a synthetic random network; it is in fact improper to refer to a “probability” when analyzing the properties of a realized river network. We clarify this aspect by introducing the concept of branching ratio pr for the latter definition (pr = NL/N). Moreover, in the case of BBTs, p and pr do not coincide (see Methods). Importantly, p and pr have no parallel in the literature on fluvial forms, nor do they refer to any of the well-studied measures of rivers’ fractal character.The choice of different observational scales for the same drainage network results in different values of NL and N, and hence of pr. Remarkably, the very same drainage network can result in river networks that virtually assume any value of pr (ranging from 0 to 1) and N (up to the upper bound A) depending on the choice of AT and A (the latter corresponding to a given l value when measured in the number of pixels; Fig. 1d–i); networks with low AT/A ratios result in high N (Fig. 2a), while networks with low AT result in high pr (Fig. 2b). Furthermore, pr does not identify the inherent (i.e., scale-independent) branching character of a given river network in relation to other river networks. In fact, by extracting different river networks at various scales (i.e., various AT values) and assessing the rivers’ rank in terms of pr, one observes that rivers that look more “branching” (i.e., have higher pr) than others for a given AT value can become less “branching” for a different AT value (Fig. 3). We therefore conclude that branching probability is a non-descriptive property of a river network, which by no means describes its inherent branching character, and depends on the observational scale.Fig. 2: Variation of N and pr as a function of observational scales for OCNs and real river networks.a Expected value of number of network nodes N as a function of threshold area AT and total drained area A (from Eq. (1)); the white dots indicate the values of AT and A used to generate the OCNs used in this analysis. b Expected value of branching ratio pr as a function of AT and A (from Eq. (1)); symbols as in a.Full size imageFig. 3: Values of branching ratio as a function of AT for the 50 real river networks analyzed in this study.a Natural values of pr in logarithmic scale. b z-normalized branching ratios (i.e., for each AT value, values of pr are normalized so that they have null mean and unit standard deviation), which better shows how rivers rank differently in terms of pr for different observation scales (i.e., AT). Lines connect dots relative to the same river. For visual purposes, rivers that rank first, second, second-to-last or last in at least one of the AT groups are displayed in colors; the other rivers are displayed in grey.Full size imageScaling is also crucial when looking at river networks from an ecological perspective. In this case, the relevant scale determining the dimension l of a node is the extent of habitat within which individuals (due to e.g. physical constrains) can be assigned to a single population57,58; the riverine connectivity and ensuing dispersal among these populations give rise to a metapopulation at the river network level. The specific spatial scale largely depends on the targeted species (e.g. being larger for fish than for aquatic insects), and it is conceivably much larger than (or, at least, it has no reason to be equal to) the pixel size of the DEM on which the river network is extracted. Since the evaluation of pr depends on the number of nodes N, which, in turn, is defined based on the scale length l, the resulting pr of a river network under this perspective would depend on the characteristics of the target taxa, which is inconsistent with the alleged role of pr as a scale-invariant property of river networks.Note also that using the ecological definition of l (i.e., spatial range of a local population) to discretize a real river network into N nodes, and from there calculate the branching ratio pr = NL/N, is problematic. Indeed, this would imply an elongation of all links shorter than l (which constitute a non-negligible fraction of the total links, under the assumption of exponential distribution of link lengths51), hence preventing a correct estimation of the connectivity patterns (i.e., distances between nodes) and the resulting ecological metrics of the river network (see section Ecological implications).From an ecological perspective, it could be reasonable to consider AT as a parameter expressing how a particular taxon perceives the suitable landscape, rather than a value to be determined from geomorphological arguments: for instance, large fishes inhabit wide and deep river reaches, and do not access small headwaters56. In this case, imposing a large AT would result in a coarser, less branching network constituted by few main channels (Fig. 1f, i), which could mimic the potentially available habitat for such species. Conversely, aquatic insects inhabit also small headwaters17,59, therefore their perceived landscape would resemble the finely resolved networks of Fig. 1d, g, characterized by low AT and higher (apparent) pr.Topology and scaling of river networks and random analoguesTo verify the topological (i.e., Horton’s laws on bifurcation and length ratios) and scaling (i.e., probability distribution of drainage areas) relationships of the different network types, we extracted from DEMs 50 real river networks encompassing a wide range of drainage areas (Fig. 4), and we generated 50 OCNs, 50 RBNs and 50 BBTs of comparable size (see Methods).Fig. 4: Location of real river basins used in the analysis.River basins are shown in dark grey; countries in light grey. Rivers’ numbering is sorted in ascending order according to drainage area values.Full size imageTypical values3,7,60 for the bifurcation ratio RB lie between 3 and 5, while length ratios (RL) range between 1.5 and 3.5. As expected, we observed that the real rivers and OCNs used in our analysis have RB and RL values within the aforementioned ranges (Fig. 5a, b). The same is true for RBNs, while the RB and RL values found for BBTs are lower than the typical ranges. This finding holds regardless of the scale (subsumed by AT) at which real river networks and OCNs are extracted (Supplementary Figs. 1 and 2). Remarkably, BBTs fail to satisfy Horton’s laws despite the statistical inevitability of such laws for any network argued by ref. 61. To this regard, we note that the networks analyzed by ref. 61 did not include constructs where all paths from the source nodes to the outlet have the same length, which is the defining feature of BBTs (Fig. 1a).Fig. 5: Comparison of topological and scaling properties of the different networks.a Scaling of number of network links Nω as a function of stream order ω for the various network types (rivers and OCNs obtained with AT = 20 pixels; RBNs and BBTs derived accordingly – see Methods). b Mean link length Lω (in units of l) as a function of ω. Networks used are as in panel a. c Scaling of drainage areas: probability P[A ≥ a] to randomly sample a node with drainage area A ≥ a as a function of a. The displayed trend lines are fitted on the ensemble values for the 50 network replicates, by excluding nodes with drainage area larger than 2000 pixels (cutoff value marked with a black solid line). The scaling coefficients β reported correspond to the slopes of the fitted trend lines. Extended details on all panels are provided in the Supplementary Methods.Full size imageWhile the power-law scaling of areas in OCNs (Fig. 5c) has an exponent β ≈ 0.45 that closely resembles the one found for the real rivers (β ≈ 0.46) and within the typically observed range8,10 β = 0.43 ± 0.02, drainage areas of RBNs scale as a power law with an exponent β ≈ 0.51, which departs from the observed range. Conversely, BBTs do not show any power-law scaling of areas. Scaling exponents of drainage areas fitted separately for each real river network yielded values in the range 0.36÷0.57 (Supplementary Table 1). In particular, we observed that these values tend to the expected range β = 0.43 ± 0.02 for increasing values of A, expressed in number of pixels (Supplementary Fig. 3), hence implying that highly resolved catchments are required in order to properly estimate β. Interestingly, the observed values of Horton ratios and scaling exponent β for RBNs are compatible with the values RB = 4, RL = 2, β = 0.5 predicted for Shreve’s random topology model3,60,62, which is actually equivalent to a RBN with infinite links.Ecological implicationsWe compared the different network types via two metrics that express the ecological value of a landscape for a metapopulation: the coefficient of variation of a metapopulation CVM and the metapopulation capacity λM. The coefficient of variation of a metapopulation63 is a measure of metapopulation stability (a metapopulation being more stable the lower CVM is), while the metapopulation capacity42,64 expresses the potential for a metapopulation to persist in the long run (persistence being more likely the higher λM is). Both measures are among the most universal metrics describing dynamics of spatially fragmented populations24,40. In order to assess the impact of the two landscape features mostly affecting metapopulation dynamics, i.e. spatial connectivity and spatial distribution of habitat patches, we calculated these metrics for the four network types under two different scenarios: uniform (CVM,U, λM,U) and non-uniform (CVM,H, λM,H) spatial distribution of habitat patch sizes. In the first scenario, CVM,U and λM,U assess stability and persistence (respectively) of a metapopulation solely based on pairwise distances between network nodes; in the second scenario, CVM,H and λM,H depend on the interplay between pairwise distances and spatially heterogeneous habitat availability (namely, downstream nodes being larger than upstream ones).We found that the values of CVM (be it derived with uniform (CVM,U) or nonuniform (CVM,H) distributions of patch sizes) obtained for OCNs match strikingly well those of real rivers (Fig. 6). These CVM values are consistently lower than those found for RBNs, while values of CVM for BBTs are even higher. Notably, this result holds for different values of AT (and hence different pr values) at which real rivers and OCNs are extracted (Fig. 6a–c; g–i), and for values of mean dispersal distance α (see Methods) spanning multiple orders of magnitude (Supplementary Figs. 4–7).Fig. 6: Comparison of values of metapopulation metrics across river network types and observational scales (AT).a–c CVM,U. d–f λM,U. g–i CVM,H. j–l λM,H. Boxplot elements are as follows: center line, median; notches, (pm 1.58cdot {{{{{{{rm{IQR}}}}}}}}/sqrt{50}), where IQR is the interquartile range; box limits, upper and lower quartiles; whiskers, extending up to the most extreme data points that are within ±1.5 ⋅ IQR; circles, outliers. Metapopulation metric values were obtained by setting α = 100 l. Note that in Eq. (1), given A = 40, 000, AT = 20 results in E[N] ≈ 4574, E[pr] ≈ 0.228; AT = 100 yields E[N] ≈ 2231, E[pr] ≈ 0.098; AT = 500 results in E[N] ≈ 1088, E[pr] ≈ 0.042.Full size imageFor a constant α value, the CVM of real rivers, OCNs and RBNs decreases as the resolution at which the network is extracted increases (i.e., AT decreases; see Fig. 6 and Supplementary Figs. 4–7). This is expected63, since a decrease in AT corresponds to an increase in N (Fig. 2a), leading to a decrease in CVM. Indeed, a larger ecosystem, constituted of more patches, has the potential to include a larger (and more diverse) number of subpopulations, which increases stability at a metapopulation level through statistical averaging–a phenomenon widely known as the portfolio effect65. We also found that BBT networks do not generally follow the above-described pattern of decreasing CVM with increasing N; rather, the CVM of BBTs increases with N when the mean dispersal distance α is set to intermediate to high values (Fig. 6 and Supplementary Figs. 5–7), and only when α is very low (e.g. α = 10 l as in Supplementary Fig. 4) and a uniform patch-size distribution is assumed does CVM,U follow the expected decreasing trend with increasing N.However, we need to warn against the conclusion that river networks with higher values of pr (and hence lower AT, see Fig. 2b) are inherently associated with higher metapopulation stability. Indeed, our result was obtained by changing the scale at which we observed the same river networks, and not by increasing the river networks’ size. If the number of network nodes (and, consequently, the branching ratio pr) is determined by the scale at which the landscape is observed, one cannot directly assume that any of such nodes is a node (or patch) in the ecological sense, i.e. the geographical span of a local population: the extent of such patches should be determined based on the mobility characteristics of the focus species, and should be independent of the scale at which the river network is observed. In contrast, we note that, if different river networks spanning different catchment areas (say, in km2) are compared, all of them extracted from the same DEM (same l and same AT in km2), then the larger river network will appear more branching (i.e., have larger pr). Indeed, by selecting catchments with larger A (in km2) for fixed l and AT (in km2), one moves towards the top-left corner of Fig. 2a, b (i.e., perpendicular to the level curves AT/A). The apparent higher “branchiness” of the river network with larger A will result in lower values of CVM; however, the higher metapopulation stability of the larger network will not be due to its (alleged) inherent more branching character, but only dictated by its larger habitat availability.We observed that metapopulation capacity λM values of OCNs (be it evaluated under uniform (λM,U) or non-uniform (λM,H) patch-size distribution assumption) are the closest to those of real rivers, while RBNs (and even more so BBTs) generally overestimate λM with respect to real rivers and OCNs (Fig. 6d–f; j–l). This result holds irrespective of the choice of AT and for intermediate to high values of α (Supplementary Figs. 5–7). When the mean dispersal distance is instead set to very low values (α = 10 l – Supplementary Fig. 4) and the river network is extracted at a high resolution (i.e., low AT), the metapopulation capacity of OCNs under assumption of uniform patch-size distribution (λM,U) is underestimated with respect to that of real rivers. A likely explanation for this apparent mismatch is that, for low values of AT, the number of nodes N tends to be somewhat higher for the extracted river networks used in this analysis than for OCNs (Supplementary Fig. 8), and the effect of the different dimensionality of real rivers and OCNs in the metapopulation capacity estimation tends to be more evident as the mean dispersal distance decreases. Interestingly, such mismatch is absent when a non-uniform patch size distribution is assumed, as λM,H values for OCNs match those for real rivers regardless of the mean dispersal distance value and the river network resolution (Fig. 6; Supplementary Figs. 4–7).The OCN construct encapsulates both random and deterministic processes, the former related to the stochastic nature of the OCN generation algorithm, and the latter pertaining to the minimization of total energy expenditure that characterizes OCN configurations. As such, OCNs reproduce the aggregation patterns of real river networks. From an ecological viewpoint, this implies that both pairwise distances between nodes and the distribution of patch sizes (expressed as a function of drainage areas, or of a proxy thereof such as the number of nodes upstream) are much closer to those of real networks than is the case for fully random synthetic networks as BBTs and RBNs. In particular, BBTs and (to a lesser extent) RBNs tend to underestimate pairwise distances with respect to real rivers and OCNs, as documented by a comparison of mean pairwise distances across network types (Supplementary Fig. 9a–c). Our analysis shows that the connectivity structure of these random networks (subsumed by the matrix of pairwise distances) is too compact with respect to that of real rivers, which leads to an overestimation of the role of dispersal in increasing the ability of a metapopulation to persist in the long run, but also an increased likelihood of synchrony among the different local populations, which results in higher instability.Comparison of patch size distributions among the network types expressed in terms of CVM,0 (i.e., the portion of CVM,H that uniquely depends on the distribution of patch sizes and not on pairwise distances) shows that, while for coarsely resolved networks (AT = 500) no clear differences in CVM,0 emerged, for highly resolved networks (AT = 20) BBTs heavily underestimate the CVM,0 of real rivers and OCNs, while RBNs slightly overestimate it (Supplementary Fig. 9d–f). As a result of the interplay of differences in distance matrices and patch size distributions, BBTs and (to a lesser extent) RBNs generally tend to overestimate the coefficient of variation of a metapopulation and the metapopulation capacity of real rivers and OCNs in both scenarios of uniform and non-uniform patch size distribution. The only exception to this trend occurs for the metapopulation capacity λM,H of very large BBTs (corresponding to AT = 20) in the case of very high dispersal distances (α = 1000 l – Supplementary Fig. 7): here, the patch-size effect (i.e., underestimation of CVM,0) predominates over the distance effect (i.e., overestimation of mean dij), resulting in an underestimation of λM,H with respect to real rivers and OCNs.Our results were derived under a number of simplifying assumptions. In particular, we acknowledge that, while the distance matrix of a landscape and the distribution of patch sizes have in general important implications for metapopulation dynamics, other factors not considered here, such as Euclidean between-patch distance48, fat-tailed dispersal kernel66 and density-dependent dispersal67 could also play a relevant role in this respect. However, it needs to be noted that, especially with regards to the assessment of the Moran effect in metapopulation synchrony (i.e., increased synchrony in local fluvial populations that are geographically close but not flow-connected48), the use of OCNs allows integration of Euclidean distances in a metapopulation model, while this is not possible for RBNs and BBTs, where Euclidean distances are not defined. Moreover, if a larger degree of realism is required for a specific ecological modelling study, such as heterogeneity in abiotic factors (e.g. water temperature or flow rates), the use of OCNs as model landscapes allows a direct integration of these variables, as they can conveniently be expressed as functions of drainage area3,55. In contrast, this is not possible for RBNs or BBTs, because only OCNs verify the scaling of areas (Fig. 5c), while RBNs and BBTs lack a proper definition of drainage areas.Our comparison of synthetic and real river networks showed that riverine metapopulations are more stable and less invasible than what would be predicted by random network analogues. Conversely, the use of OCNs as model landscapes allows capturing not only the scaling features of real rivers, but also drawing ecological conclusions that are in line with those that could be observed in real river networks. We thus support the use of OCNs as analogues of real river networks in theoretical and applied ecological modelling studies. While we found that BBTs are highly inaccurate in reproducing ecological metrics of real river networks and should be therefore discarded altogether in future modelling applications, RBNs show a certain degree of similarity with OCNs and real river networks in this respect; moreover, RBNs (as is the case for any random tree61) satisfy Horton’s laws on bifurcation and length ratios. A relevant advantage of RBNs over OCNs is that their generation algorithm is at least one order of magnitude faster49. Therefore, we acknowledge that RBNs could be considered as a suitable surrogate for real river networks as null models in cases where a large number of network replicates is required. To this end, we encourage researchers exploiting synthetic river networks (whether they be OCNs or RBNs) to always clarify the observational scales (that is, total area drained, size of a node, area drained by a headwater) subsumed by the synthetic network and which give rise to a certain complexity measure (i.e., branching ratio). Only in such a way could the predictions from these studies be compared with real river networks.In conclusion, our results advocate a tighter integration between physical (geomorphology, hydrology) and biological (ecology) disciplines in the study of freshwater ecosystems, and particularly in the perspective of a mechanistic understanding of drivers of persistence and loss of biodiversity. More

Data UsedWe used cadastral maps from 1858 to reconstruct the LULC structure of Aksu and Kestel for the mid-nineteenth century. General Staff of the Ottoman Army produced these maps in 1:10,000 scale. These maps were one of the earliest attempts of creating cadastral maps in the Ottoman Empire. The images of historical maps scanned at 1270 dpi resolutions are provided by the Turkish Presidency State Archives of the Republic of Turkey – Department of Ottoman Archives (Map collection, HRT.h, 561–567). Individual tiles of cadastral maps are of a 1:2,000 scale. However, these maps are kept separated from their accompanying cadastral registers or documentation regarding their production process in the archives. There are no studies on the production of these cadastral maps, but few studies used them until now35,36.The LULC structures of Aksu and Kestel for the mid-twentieth century were generated using aerial photographs from June 23, 1955, with a scale of 1:30,000. These aerial photographs were captured by the US Navy Photographic Squadron VJ-62 (established on April 10, 1952, re-designated to VAP-62 on July 1956, and disestablished on October 15, 1969). The squadron conducted mapping and special photographic projects worldwide37. Lastly, the VHR satellite images of WorldView-3 (WV-3) with 0.3 m of spatial resolution, collected on September 6, 2020, were used to show the up-to-date LULC patterns of Aksu and Kestel.MethodologyFigure 2 shows the flowchart of steps followed in this study to detect the LULC changes. The workflow includes three phases: preprocessing, LULC mapping, and statistical analysis of LULC changes.Figure 2Flowchart of the processing steps for the LULC change analysis for Kestel.Full size imageData preprocessingOrthorectification is the first important step in ensuring accurate spatial positioning among the multi-temporal and multi-source images, eliminating geometric distortions, and defining all data sets on a common projection system. To align the multi-modal geospatial datasets, we first performed the orthorectification of the satellite images and then we used the orthorectified satellite images as reference for the georeferencing of the cadastral maps and aerial photographs.Satellite imagery orthorectificationWe first pan-sharpened the WV-3 images by applying the PANSHARP2 algorithm38 to fuse the panchromatic (PAN) image of 0.3 m spatial resolution with four multispectral bands (R, G, B, and near-infrared (NIR)) of 1.2 m. We then geometrically corrected the pan-sharpened (PSP) WV-3 imageries using an ALOS Global Digital Surface Model with a horizontal resolution of approximately 30 m (ALOS World 3D – 30 m), rational polynomial coefficients (RPC) file, and additional five ground control points (GCPs) for the refinement. As a geometric model, we used the RPC model with zero-order polynomial adjustment39, and orthorectified images were registered to the Universal Traverse Mercator (UTM) Zone 35 N as the reference coordinate system.Georeferencing of scanned cadastral maps and aerial photographsWe used orthorectified WV-3 imageries as a reference for the geometric correction of the historical cadastral maps and the aerial photographs. We selected the spline (triangulation) transformation, a rubber sheeting method, useful for local accuracy and requiring a minimum of 10 control points, as the transformation method to determine the correct map coordinate location for each cell in the historical maps and aerial photographs. The spline transformation provides superior accuracies for the geometric correction of the historical geospatial data40,41.The lack of topographic properties of planimetric features in the historical cadastral maps and the inherent distortions of the aerial photographs due to terrain and camera tilts causes difficulties in precise georeferencing of these data sets. It increases the number of required ground control points (GCPs) for optimal image rectification. Adequate and homogenously distributed GCPs, identified from cadastral maps and aerial photographs, can ensure precise spatial alignment among different geospatial data. The best locations for GCPs were border intersections of agricultural fields and roads. As for artificial objects, places of worship and schools, which are highly probable that have remained unchanged, are other optimal locations for GCPs to perform the accurate geometric correction. Figure 3 displays samples of GCPs selected from cadastral maps and aerial photographs. We obtained 2.11 m or better overall RMSE (Root Mean Square Error) values for the geometric correction of the historical maps and aerial photographs.Figure 3Examples of GCPs selection (red crosses in blue circles) on (a), (c) Cadastral maps and their counterparts on (b), (d) Aerial photographs.Full size imageLULC classification schemeWe defined our classification scheme by analyzing the LULC classes distinguished in all three datasets (i.e., cadastral maps, aerial photographs, and WV-3 imageries). We used the classification scheme shown in Table 1. We also present codes and names of the land cover (LC) classes derived from Corine LC nomenclature42.Table 1 Classification scheme of the study.Full size tableThe legends provided on the historical cadastral maps of Aksu and Kestel delineate 15 LULC categories, which are: (1) buildings, (2) home gardens, (3) roads, (4) arable land, (5) gardens, (6) mulberry groves, (7) chestnut groves, (8) olive groves, (9) vegetable gardens, (10) forest, (11) courtyards, (12) vineyards, (13) arable fields, (14) cemeteries, (15) watercourses. Categorizing the land cover types of cadastral maps is limited with the classes available in the map legend. The legend of cadastral maps categorizes the forested area in one class named “forest”. Therefore, it was not possible to use third-level LC sub-categories in our classification schema for forest area which is separating forested areas into three subclasses (3.1.1, 3.1.2, and 3.1.3) according to the type of tree cover. Although some of the third-level LC sub-categories could be extracted from the cadastral map legend, it was not possible to extract all third level agricultural classes from single-band monochromatic aerial photographs. Although the spatial extent of fruit trees as a permanent crop could be determined from aerial photographs, it was not possible to classify these trees into different fruit types (e.g. 2.2.1 Vineyards, 2.2.2 Fruit trees and berry plantations, 2.2.3 Olive groves). Limitation on the number of forest classes is due to the historical cadastral map content; whereas limitation on the number of agricultural classes is mainly offset by the aerial photographs which have only one spectral band and we did not have any field survey or ancillary geographical data that could help the specific identification of fruit trees.Our primary focus is to find out agricultural land abandonment, deforestation/afforestation, urbanization, and rural depopulation within the historical periods. Therefore, most of the second level LULC classes are sufficient for our purpose. LULC changes within the third class level such as the conversion of third level agriculture classes among each other were not aimed to analyze in this research. Our datasets allow us to use Level 3 CORINE classes for the artificial surfaces. These classes are useful to determine residential area implications of rural depopulation or urbanization, one of the focused transformation types for our analysis.We re-organized and categorized the LULC types used in the cadastral maps, with minimum possible manipulation (only for 2.4 and 3.2 LC classes) according to the classification scheme, as shown in Table 2.Table 2 Correspondence between Corine Land Cover and historical cadastral maps nomenclature.Full size tableLULC mappingAfter aligning all geospatial data, we used the georeferenced cadastral maps, aerial photographs, and satellite images for the LULC mapping. We set the spatial extent of the selected regions based on boundaries digitized from the cadastral maps of 1858. Then we detected historical LULC changes within these extents for all geospatial datasets covering 1900 ha and 830 ha of the Aksu and Kestel regions, respectively. Figures 4 and 5 show the selected extents from the historical maps, aerial photographs, and satellite images of the Kestel and Aksu sites, respectively.Figure 4Geospatial dataset for the Kestel study region. (a) 1858 Cadastral map, (b) 1955 aerial photo, and (c) 2020 WV-3 satellite image (finer details shown in the inset images highlighted by Blue boxes).Full size imageFigure 5Geospatial dataset for the Aksu study region. (a) 1858 Cadastral map, (b) 1955 aerial photo, and (c) 2020 WV-3 satellite image (finer details shown in the inset images highlighted by red boxes).Full size imageDigitization of cadastral maps-1858 LULC mapsWe visually interpreted and manually digitized the geographic features on the historical maps and created vector data for each class. The road features in cadastral maps lack topological properties. They also include spatial errors possibly generated in the process of surveying and map production. Therefore, we cross-checked digitized road segments by visual inspection of the road data of the aerial photographs from 1955. We then further modified road polygons to represent accurate road widths. Afterward, we categorized vectorized features of the cadastral maps into the LULC classes defined in Table 1. Finally, we created the vectorized 1858 LULC map. Figure 6 presents the vectorized 1858 cadastral maps of Aksu and Kestel.Figure 6Vectorized cadastral maps of (a) Kestel and (b) Aksu with Red and green lines showing the vector boundaries.Full size imageObject-based image analysis of aerial photographs-1955 LULC mapsAt the second stage of LULC mapping, we performed the segmentation and classification of the aerial photographs using an object-based approach for generating the 1955 LULC map. The object-based image analysis (OBIA) approach in LULC mapping provides advantages over the traditional per-pixel techniques such as higher classification accuracy, depicting more accurate LULC change, and differentiating extra LULC classes33,43,44. We used the eCognition® software (Trimble Germany GmbH, Munich) to implement an object-based image analysis (OBIA). The OBIA approach contains two phases including the segmentation and classification phases that are performed to locate meaningful objects in an image and categorize the created objects, respectively.Multiple ancillary datasets have been used to support different phases of OBIA. The Open Street Map (OSM) vector data, an open-source geospatial dataset (http://www.openstreetmap.org/), has been utilized as ancillary vector data in OBIA to improve the classification of the remotely sensed images. Sertel et al. (2018) used OSM as a thematic layer for road extraction7. Since there are several limitations in extracting the roads from aerial imagery, the OSM road network data could be useful. A majority of unpaved roads in single-band aerial photographs can easily be misclassified as homogeneous areas of arable lands. Precise detection of the roads from monoband aerial photographs without multi-spectral information is difficult. Therefore, we overlaid the OSM road network data with the aerial photographs to extract the revised aerial road vectors through visual interpretation and manual digitization.We segmented the 1955 aerial photographs with the integration of 1858 LULC map produced from cadastral maps. We implemented the multi-resolution segmentation algorithm. In this segmentation method, a parameter called scale determines the size of resulting objects, and the shape and compactness parameters determine the boundaries of objects. The segmentation process of the aerial photographs was performed at multiple stages with various scale, shape, and compactness parameter values. At the initial stage, we segmented the regions according to the 1858 LULC map and we utilized large-scale parameters. The scale parameter was set to 100 and the shape parameter and the compactness were set as 0.7 and 0.3, respectively. At this stage, we focused on interpreting the objects that have not changed between 1858 and 1955. We classified the segments using the thematic layer attribute (LULC classes defined by the cadastral maps) with the highest coverage. Image objects in which the land surface has changed during 1858–1955 period were detected by visual interpretation and unclassified for further segmentation. We followed this approach to reduce the manual effort. We defined unchanged objects between 1858 and 1955 and assigned the same classes of 1858 LULC map to the objects in 1955 aerial photographs. We then segmented the remaining segments, the last time into smaller objects with the scale parameter set as 25, the shape parameter set as 0.2, and the compactness set as 0.8.We classified the remaining unclassified objects through the development of rulesets. An object can be described by several possible features as explanatory variables which are provided by eCognition. In the classification ruleset, different features and parameters can be defined to describe and extract object classes of interest and thresholds for each feature can be defined by the trial-and-error method. We tested sets of variables for the classification of the monoband aerial photographs. Object features such as the mean value of the monoband, texture after Haralick, distance to neighbor objects, shape features (e.g., rectangular fit and asymmetry), and extent features (e.g., area and length/width) were the most useful alternatives. The classification process of the parcels of the aerial photographs with LULC change started with the classification of roads constructed between 1858 and 1955 by utilizing the aerial road map. The watercourse class was the most difficult to classify since shrubs or trees mostly covered the watercourses. These areas were misclassified as forest or agricultural land. Therefore, experts in historical map reading with local geographical information performed the detection and classification of the water course class and interpreted by the cadastral map (1858) and the google map (2020). After roads and watercourses, we classified forest and agricultural lands using the optimal thresholds for the brightness feature. We calculated the thresholds using the single band of the aerial photograph combined with the area and rectangular fit features. The heterogeneous agricultural areas class principally occupied by agriculture with significant areas of natural grass and trees within the same object are separated from the arable lands using the standard deviation of the digital number (DN) values of the aerial photographs. The texture feature helped classify the permanent crops. The brightness, shape, asymmetry, and distance to road class features were the best-performing ones for classifying the remaining artificial surfaces. The manual interpretation was performed for the classification of sub-classes of artificial surface class, including the continuous/discontinuous urban fabric, industrial, commercial, and transport units, mine, dump and construction sites, and artificial, non-agricultural vegetated areas. Since these land use classes contain one or more land cover and land use categories (e.g., artificial non-agriculture land or industrial or commercial units), finding the optimal threshold and exact feature for distinguishing the subclasses of artificial surfaces is difficult. Especially in the case of using the single-band aerial photographs, manual interpretation was required.Object-based image analysis of satellite images-2020 LULC mapsWe segmented WV-3 satellite images using multi-resolution segmentation algorithm and ancillary geographic data. Similar to the aerial road map, the road network of the study region in 2020, named, WV-3 road map, was extracted by overlaying the OSM road data with the WV-3 satellite image. In the segmentation process of the WV-3 image, we used the vector boundaries of the classified aerial photograph (the 1955 LULC map) and the WV-3 road map as ancillary thematic layers. We opted for the same segmentation and classification approach used for the aerial photographs for the WV-3 image.Firstly, we segmented the satellite image into spectrally homogeneous objects using vector data of the 1955 LULC map by applying large-scale parameters. We implemented scale parameter values of 300, 200, 100, and 50 to find the optimal scale to classify objects that have not changed between 1955 and 2020. The best multi-resolution segmentation configuration was the scale of 100 and the shape and compactness parameters of 0.3 and 0.7, respectively. We classified the segments using the thematic layer attribute (LULC classes defined by the aerial maps) with the highest coverage. Segments with LULC change, e.g. the image objects in which the land surface has changed during 1955–2020 period were detected by visual interpretation and unclassified for further segmentation. As a result, we excluded the objects which were remained unchanged during 1955–2020 by assigning the prepared labels which were allocated in the previous step during the classification of 1955 aerial photographs. We then segmented the remaining objects into smaller objects to identify the changed areas in detail. At this step, the scale, shape, and compactness parameters were set as 25, 0.2, and 0.8, respectively.Except for the additional sets of variables utilized to classify the WV-3 images, we applied the rule-set developed for the classification of the aerial photograph for the classification of the remaining objects of 2020 satellite images. The additional sets of variables include the mean of G, B, R, and NIR and two spectral indices, the Normalized Difference Water Index (NDWI), and the Normalized Difference Vegetation Index (NDVI). NDVI was calculated as the normalized difference of reflectance values in the red and NIR bands; whereas , NDWI was determined as the normalized difference of reflectance values of the green and NIR bands. Through the logical conditions, objects having specified values of NDVI and NDWI can be assigned to vegetation and water classes, respectively. The use of NDVI facilitated the delineation of terrains covered by vegetation and the NDWI improved the extraction of water bodies due to its ability to separate water and non-water objects. We separated different sub-classes of agricultural areas and forests by using optimal thresholds for NDVI which were defined by a trial and error method. Also we utilized assigning the optimal threshold to NDWI to separate water bodies from other land covers. In addition, the mean blue band layer was useful in classifying the artificial surfaces. We assessed the accuracy of each classification using error matrices (overall, user’s and producer’s accuracies, and Kappa statistics)45,46.Estimating LULC changes and LULC conversionsAfter the production of LULC maps of Aksu and Kestel for 1858, 1955, and 2020, the vector data of the LULC maps were used to quantify the LULC conversions for two different periods which are 1858–1955 and 1955–2020. To compare the LULC maps of study areas between two different dates of each study period, we provided detailed “from-to” LULC change information by calculating the LULC change transition matrix computed using overlay functions in ArcGIS.We overlaid LULC maps of 1858 and 1955 and intersected the vector boundaries of the 1858 and 1955 LULC maps to determine the conversion types of LULC classes (from which class to which class). Similarly, to quantify the LULC changes between 1955 and 2020, we overlaid the 1955 and 2020 LULC maps. Then we created transition matrices and performed statistical analysis utilizing the matrices. Finally, we discussed the main LULC change types and the driving factors of the changes in the selected study areas. More

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Animal husbandryCarukia barnesi polyps were available in culture from the James Cook University Aquarium, spawned from medusa originally collected near Double Island, North Queensland, Australia (16° 43.5′ S, 145° 41.0′ E) in 2014 and 20158. Populations exponentially increase through asexual reproduction8. Detached buds and swimming polyps were collected from the main culture, and transferred into 6-well tissue culture plates in natural filtered seawater. Plates were maintained in darkness to inhibit algae growth at 27 °C in a constant temperature cabinet. Buds and swimming polyps were left to develop and attach to well bottoms, at which point they were then fed freshly hatched Artemia nauplii and water changed 2–3 times per week. Lids remained attached to tissue culture plates to negate water evaporation and maintain a stable salinity. Polyps were maintained in this way for a minimum of 4 months before experiments began, with all individuals matured with the ability to asexually reproduce further buds. To preserve water quality15 polyps were starved for two days prior to experiment start and were not fed for the duration of the trials. One day prior to the experiment start, all immature buds and polyps were removed from wells, leaving approximately 5–10 mature polyps attached to the substrate for analysis.Preparation of reagentsReagentsSix chemicals were trialed in the current study to induce metamorphosis in C. barnesi polyps. Four indole containing compounds were chosen that have previously been trialed with other cubozoan species: 5-methoxy-2-methyl-3-indoleacetic acid, 5-methoxyindole-2-carboxylic acid, 2-methylindole16 and 5-Methoxy-2-methylindole15,16. Along with the retinoic X receptor 9-cis-retinoic acid and lugols solution.Indole compound treatmentsChemical concentrations of indoles documented in the literature were used to conduct preliminary concentration tests. Fifty mM stock solutions were prepared with 100% ethanol, which was diluted with filtered seawater to the desired experimental concentrations: 50 μM16, 20 μM and 5 μM15. Due to high fatality rates at all of these concentrations when used in this study on C. barnesi, all concentrations were diluted. Fifty mM stock solutions of 5-methoxy-2-methyl-3-indoleacetic acid, 5-methoxyindole-2-carboxylic acid, 2-methylindole and 5-Methoxy-2-methylindole were prepared with 50% ethanol (50% Milli-Q® water) and stored at − 20 °C. Fifty mM stock solutions were diluted with filtered seawater to the experimental concentrations of 5 μM, 1 μM, 0.5 μM, 0.1 μM and 0.05 μM. The carrier solution of 50% ethanol (50% Milli-Q® water) was diluted to the equivalent of the experimental concentrations listed above for use as a control, and incorporated into data as concentration 0. Seventeen ml of solution was added to polyps to fill each well of a 6-well plate.Iodine treatment (lugols solution)Aqueous iodine in the form of Lugols solution (0.37% iodine and 0.74% potassium iodide (sigma product information)) was prepared with equivalent concentrations of moles iodine/iodide: 1.5 μM, 3 μM, 6 μM, 12 μM and 24 μM. Filtered seawater only was used a control for this treatment and incorporated into data as concentration 0. 17 ml of solution was added to polyps to fill each well of a 6-well plate.Retinoid treatmentTo reduce ethanol associated fatality of polyps 0.015% ethanol in Milli-Q® water was used to prepare a 1 mM stock solution of 9-cis-Retinoic acid. The 1 mM stock solution was diluted with filtered seawater to the experimental concentrations of 5 μM, 1 μM, 0.5 μM, 0.1 μM and 0.05 μM. The carrier solution of 0.015% ethanol (Milli-Q® water) was diluted to the equivalent of the experimental concentrations listed above for use as a control, and incorporated into data as concentration 0. 17 ml of solution was added to polyps to fill each well of a 6-well plate.Metamorphosis trialsPrimary trialsExperimental concentrations of reagents were added to C. barnesi polyps growing in the wells of sterile 6-well tissue culture plates. One plate was used per chemical, per concentration, in which five wells functioned as replicates containing the chemical being trialed, whilst the sixth well contained only the control medium. Five concentrations were run for each of six chemicals; 30 plates in total.The filtered seawater the polyps were growing in was exchanged for the experimental chemical on day 0, and was not changed for the duration of the trial. Lids remained attached to tissue culture plates to negate water evaporation and hence salinity changes.Polyps in each well were photographed each day through a dissection microscope over a period of 34 days. Results were then recorded from the photographs, categorised (Fig. 6) as the number of polyps which displayed:Tentacle migration: one of the key signs of metamorphosis in this species, polyp tentacles merge, migrating to form four distinct corners in a square shape8.Detached medusa: a medusa formed and detached from the polyp, recorded regardless of health.Mobile detached medusa: a healthy medusa formed and detached from the polyp, with the ability to swim.Polyp survival: this was then used to calculate the number of polyps which survived the treatment which did not metamorphose.Optimisation trialThe optimal chemical and concentration was then deduced by choosing the combination that produced the largest percentage of healthy detached medusa, in this case 5-methoxy-2-methylindole at 1 μM. A final trial was then run with this to determine if length of chemical exposure could optimize healthy medusa yield. Three replicates of a minimum of five polyps were used per treatment, in which in 1 μM of 5-methoxy-2-methylindole (in seawater) was added to polyps for 24, 48, 72, 96 and 120 h, before the solution was changed to fresh seawater. A sea water only control was also run. The total number of healthy detached medusa were recorded each day.Data analysisAll statistical analysis was conducted in IBM SPSS Statistics Ver28. Graphs were produced in Microsoft Excel 2016 and OriginPro Graphing and Analysis 2021.Primary trialsThe effect of chemical, concentration and time was analysed using a repeated measures three-way ANOVA for four sets of data gathered during the metamorphosis process: percentage of polyps to display tentacle migration, percentage of polyps to have medusa detach, percentage of polyps to have healthy swimming medusa detach, percentage survival of polyps that did not metamorphose. Percentage data was arcsine square root transformed prior to analysis. Mauchly’s Test of Sphericity indicated that the assumption of sphericity had been violated on all four sets of data and therefore, a Greenhouse–Geisser correction was used.Optimisation trialDifferences in the mean percentage of healthy medusa produced at different exposure times was analysed using ANOVA. Differences between means were elucidated using a Post hoc Tukey pairwise comparison test (Tukey HSD alpha 0.05). More

Our study reveals zinc isotopes to be a promising trophic indicator in sharks and other fishes in general, similar to previous studies featuring both terrestrial and marine mammals9,10,11,12,13. We analysed the Zn isotope values for extant sharks spanning captive/aquarium and wild individuals from various localities and found close correspondence with their respective trophic level. Further, Zn concentration and isotope composition suggest preservation of this biological signal in fossil specimens with little diagenetic alteration. A survey of fossil shark teeth spanning the Miocene–Pliocene reveal similar δ66Zn values and variation as found in related (e.g., congeneric) extant sharks with similar dentition and ecology. Our δ66Zn results indicate high trophic levels for Otodus and perhaps a trophic change in C. carcharias, the great white shark.Zinc isotopes in extant sharks and teleostsAs with mammals9,10,11,12,13, bioapatite δ66Zn values in wild extant elasmobranchs and teleosts show overall lower values with increasing trophic level (Fig. 1; Supplementary Fig. 1). Both δ66Zn and δ15Ncoll correlate with FishBase17 trophic levels, despite differences in species’ geographic origin and tissue types sampled (Spearman’s correlation r = −0.87, p = 5.65E–16, n = 48 and r = +0.42, p = 1.47E–5, n = 40, respectively). There is no statistically significant relationship between bioapatite δ66Zn values and δ13Ccoll, but there is one between wild fish δ66Zn and δ15Ncoll values from the same tooth or individual (R2 = 0.28, p = 6.89E–4, n = 38; Supplementary Fig. 2). Both proxies thus generally reflect trophic levels. Large apex predatory sharks (e.g., Carcharodon carcharias, Isurus oxyrinchus, and Lamna ditropis) have significantly more negative δ66Zn values than lower trophic level teleosts and the plankton-feeding basking shark (Cetorhinus maximus). In particular, the shortfin mako shark (I. oxyrinchus) and great white shark (Carcharodon carcharias), both apex predators18,19, have much lower δ66Zn values than in any previously recorded extant vertebrate species (enameloid up to –0.71 and –0.63‰, respectively). These low δ66Zn values are likely due to the larger number of trophic levels in the marine ecosystem than in terrestrial food webs (terrestrial mammal enamel lies typically between 0 and +1.6‰9,11) and perhaps differences between marine and terrestrial Zn isotope baselines.Fig. 1: Zinc isotope (δ66Zn) composition of extant elasmobranch and teleost fish teeth and gill raker.Specimens come from off the coast of KwaZulu-Natal (KZN) South Africa, New Jersey (NJ), California (CA), North Carolina (NC), Iceland (IS), Norway (NO), Florida (FL), Cyprus (CY), Massachusetts (MA), Alaska and Israel. Aquarium sharks are from the New York (NY) and Tokyo (TYO) Aquariums and the Eilat (Israel) Underwater Observatory Park. Pisciculture S. aurata individuals are numbered and plotted individually to visualise the homogeneity among control-fed individuals compared to wild elasmobranchs and teleosts. Silhouettes are not to scale. Measurement uncertainty is indicated at the 2 SD level. Samples are colour-coded following their genus, regardless of locality. Source data are provided as a Source Data file.Full size imageAbsolute enameloid δ66Zn values vary by up to 1.26‰ among the extant species analysed from various oceanographic areas (Fig. 1). Our results also demonstrate large variability in enameloid δ66Zn values among extant sharks within the same region; for example, there is a 0.88‰ difference between mean values of Carcharodon carcharias and the bull shark (Carcharhinus leucas; Fig. 1) from KwaZulu-Natal (South Africa, KZN). In contrast, enameloid δ66Zn from the same species (e.g., Carcharodon carcharias, Carcharhinus obscurus) demonstrate a low isotopic variability, independent of geographic location (Fig. 1).We observe uniformity in the enameloid δ66Zn values of five gilt-head bream (Sparus aurata) individuals fed on a controlled fish pellet diet in pisciculture cages located offshore of Central Israel, with values within the measurement uncertainty of each other (–0.01 ± 0.01‰). As with δ66Zn, the δ15Ncoll and δ13Ccoll values are distinct from those of wild teleost individuals caught nearby in Haifa Bay, reflecting the artificial pelleted diet of the pisciculture individuals (Fig. 1, Supplementary Figs. 3, 4). Strongly contrasting the homogenous control fed S. aurata δ66Zn values, we observe a higher δ66Zn variability among (and even within) wild and aquaria elasmobranch individuals (fed with wild-caught fish and cephalopods). For instance, two teeth of a single tiger shark (Galeocerdo cuvier) individual (–0.52 and –0.27‰, Fig. 1) have a variability higher than the total variability among the three KZN G. cuvier individuals. Galeocerdo cuvier is well known for its highly opportunistic prey selection20. Therefore, the δ66Zn value of bioapatite is likely highly responsive to an individual’s diet at the time of tissue formation, and as shark teeth form and replace continuously, enameloid δ66Zn values can vary among teeth of a single individual. Thus, although fish can absorb Zn via their gills, waterborne Zn absorption appears to have a negligible effect on elasmobranch tooth δ66Zn values, in line with Zn incorporated into soft and skeletal tissues in natural environments being predominantly derived from dietary gastrointestinal uptake7,8.Carcharhinus enameloid δ66Zn values are high relative to sharks with similar bulk δ15Ncoll values, which contrary to the here analysed Carcharhinus species more regularly consume pelagic prey offshore, oceanic and on the continental shelf (e.g., Galeocerdo cuvier)21. This discrepancy may relate to Carcharhinus species inhabiting neritic waters where they feed primarily on demersal/benthic, freshwater-brackish-coastal prey22,23,24,25. While the diet of KZN G. cuvier and Carcharodon carcharias can also include reef-associated or demersal prey, pelagic organisms are typically more important by mass, especially in adult individuals20,26. Zinc isotope variability among marine organisms and their tissues is largely unknown, currently limiting our ability to identify specific food items based on shark enameloid δ66Zn values beyond generally observed trophic level effects. Whether higher Carcharhinus enameloid δ66Zn values relate to specific prey species (and trophic level) or general differences in basal organic matter source between a primarily neritic food web compared to a more open marine pelagic food web remains unclear (Supplementary Discussion 1). However, we observe no difference in δ13Ccoll values that imply a more terrestrial carbon signal in the KZN Carcharhinus species relative to sympatric species, arguing against differences in the basal organic matter source amongst the KZN shark species (Supplementary Fig. 3).A previous study on Arctic marine mammal bones suggested a higher geographic independence of δ66Zn values from baseline variability compared to δ13Ccoll and δ15Ncoll values13. Likewise, fish taxa with similar diet composition, habitat use and/or trophic level, have a similar range of bioapatite δ66Zn values regardless of their geographic locality (Fig. 1), indicating that δ66Zn may allow worldwide dietary and trophic level comparability with limited marine baseline variation. Further studies will need to expand our knowledge on δ66Zn variability in extant marine vertebrates as well as the effects of baseline on marine vertebrate enameloid δ66Zn values, especially compared to dentine δ13Ccoll and δ15Ncoll values. Nevertheless, the high taxa-specific and perhaps baseline-independent δ66Zn values suggest δ66Zn is an independent indicator of trophic level and an asset for present and past food web reconstructions in the marine realm.Deep-time zinc isotope preservation in fossil enameloidFossil enameloid has δ66Zn values and Zn concentrations ([Zn]) in the range of extant elasmobranch species, arguing against significant diagenetic modification (Figs. 2, 3, Supplementary Fig. 5). Fossil shark teeth examined herein are from Germany, Malta, Japan, North Carolina (USA) and Florida (USA) covering the Early Miocene (Burdigalian, 20.4–16.0 Ma), Miocene-Pliocene transition (Messinian-Zanclean boundary, ca. 5.3 Ma), and the Early Pliocene (Zanclean, 5.3–3.6 Ma; Fig. 2; Supplementary Note 2). Importantly, extant and fossil elasmobranch enameloid δ66Zn values (–0.71 to +0.28‰ and –0.83 to +0.27‰, respectively) differ from: (1) previously reported values of terrestrial mammal enamel (0 to +1.6‰);9,11 and (2) sedimentary carbonate δ66Zn values of the fossil sites (+0.34 to +0.49‰, Supplementary Fig. 6, Supplementary Table 1). These differences support a preserved biological signal in fossil enameloid.Fig. 2: Zinc isotope (δ66Zn) composition of fossil elasmobranch enameloid.Teeth are from the Early Pliocene of Japan, North Carolina (NC) and Florida (FL), Pliocene to Miocene transition of Florida, the Early Miocene of North Carolina, Germany and Malta. For more details on the sample background, see Supplementary Data 1, Supplementary Note 2. Silhouettes are not to scale. Measurement uncertainty is indicated at the 2 SD level. Samples are colour-coded following their genus, regardless of locality. Source data are provided as a Source Data file.Full size imageFig. 3: Zinc isotope (δ66Zn) composition of fossil and extant elasmobranch enameloid of selected taxa combined from Figs. 1 and 2.Fossil teeth are from multiple locations and ages. Grey silhouettes indicate extant teeth. The boxes for n  > 5 represent the 25th–75th percentiles (with the median as a horizontal line) and the whiskers show the 10th–90th percentiles. Box plots (and n) do not include aquarium teeth (open squares). Otodus spp. includes all O. chubutensis (dark blue) and O. megalodon teeth (light blue) analysed and samples are otherwise colour-coded following their genus. Silhouettes are not to scale. For more details on the samples, see Figs. 1 and 2 and Supplementary Data 1. Source data are provided as a Source Data file. Measurement uncertainty is indicated at the 2 SD level.Full size imageWe observe the same within tooth Zn spatial concentration pattern in extant and Miocene tiger shark (Galeocerdo spp.) teeth, with Zn being more enriched in the outer enameloid than close to the enameloid-dentine junction. If significant diagenetic Zn exchange had occurred throughout the enameloid, this original Zn concentration pattern would not be preserved in the fossil tooth (Supplementary Fig. 7). Additionally, both extant and fossil shark enameloid show the same variation in [Zn] according to their taxonomy, with carcharhiniforms generally having higher [Zn] than lamniforms (Supplementary Fig. 5), again arguing against significant diagenetic enameloid Zn exchange. For the European Miocene sites, δ18OP analyses were also conducted on a subset of teeth, where their enameloid appears to be generally well-preserved as suggested by δ18OP values demonstrating species-specific relative in-vivo temperature ranges as expected compared to the habitat use of equivalent modern species27 (Supplementary Fig. 8).To discern the effects of diagenetic Zn alteration, we compare visually pristine appearing enameloid with areas sampled along fractures and dentine of the same tooth (Supplementary Figs. 6 and 9, Supplementary Table 2). Our results demonstrate that the diagenetic Zn exchange in fractured enameloid leads to higher δ66Zn values than in the pristine enameloid of the same tooth, whereas we observe no differences in enameloid δ66Zn profiles of modern teeth (Supplementary Figs. 9 and 10, Supplementary Table 3). Likewise, the diagenetically more susceptible fossil dentine shows higher δ66Zn values as reflected by a significantly higher and more variable dentine-enamel δ66Zn offset (+0.78 ± 0.33‰, n = 13) than observed for extant teeth (+0.22 ± 0.1‰, n = 23; Supplementary Discussion 2, Supplementary Figs. 6 and 11). For the fossil enameloid shown here, in-vivo δ66Zn values must be at least as low as their current values, indicating limited to no alteration. Consequently, δ66Zn analysis of fossil enameloid can enable deep-time dietary reconstructions.The homogeneity in δ66Zn values for the same species or genera independent of locality and geological age is a remarkable observation (Figs. 2, 3), not only limiting the likelihood of Zn diagenetic alteration but also arguing for minimal variability in habitat-specific food web baselines or a strong homogenisation of δ66Zn values at low trophic levels. There are still some limitations, such as the absence of reported δ66Zn values of marine non-mammalian vertebrates for comparison outside this study, the limited sample size for some species, and uncertainties regarding Zn isotope baseline variability. However, our extensive δ66Zn dataset includes not only multiple species from different localities and periods with distinct differences in dietary Zn uptake among extinct elasmobranch species, but also direct overlap in extant and fossil δ66Zn values of the same genus and/or lifestyle. This spatial and temporal coherence suggests that it may be possible to use the same interpretative framework on extant and fossil elasmobranch assemblages globally (Figs. 1, 2, 3), and our remaining discussion is based on this assumption.Zinc isotopes and ecology of Miocene-Pliocene sharksAbsolute and relative δ66Zn values among some taxonomic groups show no statistical variation with geologic age and locality (e.g., Carcharias spp., Galeocerdo spp.), indicating relatively stable trophic levels and ecological niches throughout time and space. For example, most extinct elasmobranchs with a slender tearing, grasping tooth morphology (e.g., Carcharias) have δ66Zn values that can be directly compared to modern equivalents (e.g., Carcharias taurus, Isurus oxyrinchus, Lamna ditropis). This type of dentition and corresponding tooth morphology are adapted to restrain small, active prey—like fish and cephalopods28,29,30. However, there are differences among the δ66Zn values for these types of elasmobranchs within the Early Miocene of Germany, with Mitsukurina lineata and Pseudocarcharias rigida having higher mean δ66Zn values compared to Araloselachus cuspidatus (Fig. 2). Indeed, post hoc Tukey pairwise comparisons draw out A. cuspidatus as distinct from most species for the Germany (Early Miocene) assemblage, including those with a similar grasping tooth morphology (Supplementary Table 4). Our δ66Zn values indicate that A. cuspidatus was likely a higher trophic level piscivore than M. lineata and P. rigida, supported by the larger tooth size of A. cuspidatus.Zinc isotope values within the Galeocerdo lineage show no statistical variability with age nor locality, suggesting tiger sharks occupied a similar trophic level and ecological role in the marine ecosystem since at least the Early Miocene (Fig. 3). Notably, our results imply that the increase in body size from G. aduncus to the modern G. cuvier did not change its overall trophic level, which is in line with the highly similar tooth morphology between the two species31.For Carcharhinus, the Early Miocene Malta assemblage is drawn out as statistically different from extant wild Carcharhinus spp. (Supplementary Table 5). Still, Carcharhinus spp. in both extant and fossil assemblages always have higher mean δ66Zn values than other sympatric predatory sharks and are drawn out as statistically different from sympatric shark species in each fossil assemblage (Figs. 2, 3, Supplementary Tables 6–8). Based on similarities in tooth morphology and δ66Zn values among extant and extinct Carcharhinus spp., we suggest that extinct taxa also primarily occupied a neritic-coastal habitat feeding upon demersal-benthic prey22,23,24,25. For the Carcharhinus teeth from Malta, this interpretation is supported by lower δ18OP values than sympatric species, indicating a higher water temperature or lower salinity: i.e., a shallow and/or brackish water habitat (Supplementary Fig. 8). Consequently, the uniformly higher δ66Zn values of extant and fossil Carcharhinus spp. indicate the consumption of food items distinct from other measured sympatric species already during the Early Miocene and Early Pliocene.Absolute δ66Zn values for Otodus spp., along with values relative to sympatric species, indicate megatooth sharks were apex predators feeding at a very high trophic level (Figs. 2, 3). In all Early Miocene assemblages, mean O. chubutensis δ66Zn values are among the lowest compared to sympatric species, including the lowest bioapatite δ66Zn value measured to date (–0.83‰). Mean O. chubutensis δ66Zn values are as low as extant Carcharodon carcharias (respectively, –0.57 ± 0.18‰, n = 19 and –0.57 ± 0.05‰, n = 4). Noteworthy, Games-Howell pairwise comparisons indicate the lower extant C. carcharias δ66Zn values as distinct from most fossil Carcharodon populations, possibly indicating a dietary shift in the Carcharodon lineage (Supplementary Table 9). Early Pliocene values from O. megalodon from Japan also demonstrate very low mean δ66Zn values (–0.62 ± 0.11‰, n = 5) that are statistically different from the Atlantic O. megalodon populations sampled from Florida and North Carolina, which have higher mean δ66Zn values (respectively, –0.34 ± 0.11‰, n = 11; –0.38 ± 0.11‰, n = 7; Fig. 2, Supplementary Table 10).Possible explanations for the observed spatial and temporal variability in Otodus and Carcharodon δ66Zn values in our study are differences in prey consumption (and trophic level) or baseline variation. Additionally, we cannot rule out other factors such as interpretive limitations due to sample sizes. For example, extant C. carcharias can exhibit some degree of dietary individuality32, yet we only have δ66Zn data from two individuals (4 teeth) from two localities. Still, the low δ66Zn values in both extant C. carcharias compared to other extant sharks is in line with the generally high trophic level estimates of this species18. Particularly for O. megalodon from Japan where we have only one species analysed, we cannot exclude the possibility of either differences in δ66Zn baseline or regionally different prey species. However, the absence of significant δ66Zn differences within many taxa amongst locations and geological ages implies negligible differences in δ66Zn food web baselines (Figs. 2, 3). Therefore, the observed spatial and temporal variability in δ66Zn values likely demonstrates true dietary differences amongst Otodus and Carcharodon populations both geographically and temporally, with important implications for each species’ feeding ecology and evolution both on a local and global scale.Otodus and Carcharodon in the Early Miocene are represented by O. chubutensis and C. hastalis, respectively, with statistically significant higher mean δ66Zn values from the latter (Figs. 3, 4, Supplementary Tables 11, 12). The mean δ66Zn value for all O. chubutensis is the lowest of all mean values recorded in our fossil shark dataset (Fig. 3), suggesting that O. chubutensis could occupy a higher trophic position than C. hastalis. Importantly, differences between δ66Zn values of O. chubutensis and C. hastalis do not appear related to a different ratio of juveniles to adults in either species, as our results do not record an ontogenetic diet shift (Supplementary Fig. 12). We observe no correlation between the total body length of Otodus spp., Carcharodon spp. and their respective δ66Zn values (Supplementary Fig. 12), likely, because each examined specimen had already surpassed the body size for which ontogenetic dietary shifts, if any, occur.Fig. 4: Results from post-hoc Games-Howell pairwise comparisons of δ66Zn values of enameloid from fossil and extant Otodus spp. and Carcharodon spp.All assemblages and ages are combined for a given species, except for extant C. carcharias. Extant C. carcharias teeth are indicated with grey silhouettes. a Includes all O. megalodon populations, whereas (b) excludes the Japanese (Pacific) population, focusing on Atlantic and Tethys/Paratethys populations only. The boxes for n  > 5 represent the 25th–75th percentiles (with the median as a horizontal line), and the whiskers show the 10th–90th percentiles. Significance level is indicated by “*” (p value < 0.05), “**” (p value < 0.005), “***” (p value < 0.0005) and “****” (p value < 0.00005). Measurement uncertainty is indicated at the 2 SD level. See also Supplementary Tables 11, 12. Otodus chubutensis (dark blue) and O. megalodon teeth (light blue) are coloured separately. All other samples are colour-coded following their genus. Source data are provided as a Source Data file. Silhouettes are not to scale.Full size imageWhen including only Otodus spp. from the Atlantic and Paratethys/Tethys regions, we observe a statistically significant difference between O. chubutensis and O. megalodon (Supplementary Table 12, Fig. 4b). During the Early Pliocene, the Otodus lineage represented by O. megalodon shows a considerable increase in the mean δ66Zn value for the Atlantic populations, hinting at a reduced trophic position for the megatooth shark lineage in the Atlantic. At the same time, the Early Pliocene C. carcharias remains at the same trophic level as C. hastalis (Figs. 2–4, Supplementary Tables 11, 12). Although the extant sample size is limited, our results are intriguing because the mean δ66Zn value for extant C. carcharias places it at a trophic level that would be higher than the Atlantic Early Pliocene O. megalodon (Figs. 3, 4, Supplementary Table 11, 12).Extant Carcharodon carcharias is a predatory shark whereby larger individuals regularly feed on high trophic level marine mammals33. Although Neogene Carcharodon and Otodus were likely opportunistic in their prey selection similar to many extant apex predatory sharks33, fossil evidence of bite marks suggests that both taxa fed largely on marine mammals such as cetaceans (mysticetes and odontocetes) and pinnipeds1,2,4,34,35,36,37,38. However, in the majority of cases, it remains unclear if these feeding events on mammals document active hunting or scavenging and how important each prey taxa were to their overall diet. Early Pliocene C. carcharias and O. megalodon δ66Zn data suggest that lower trophic level mammal prey such as mysticetes (and perhaps herbivorous sirenians) may have been an important food item for both species. Mysticetes are filter-feeders and likely to have higher tissue δ66Zn values than piscivorous odontocetes or pinnipeds, similar to the higher δ66Zn values in the plankton-feeding extant Cetorhinus maximus compared to piscivorous sharks (Fig. 1). Bite marks on Late Miocene–Early Pliocene mysticetes bones from both Carcharodon carcharias and O. megalodon1,4,34,38 corroborate at least occasional feeding events.Now extinct small- and medium-sized mysticetes (e.g., Cetotheriidae and various small-sized Balaenidae and Balaenopteridae) were abundant during the Early Pliocene39,40 and were thus available as prey for large sharks, i.e., Otodus megalodon4 and Carcharodon carcharias1. In contrast, Early Miocene cetacean fossils are dominated by toothed cetaceans, where the Early Miocene European and North American sites sampled in this study lack any mysticete remains41,42,43,44. The Early Miocene Otodus (and modern C. carcharias) lower δ66Zn values (higher trophic level) may partly be related to the lack of lower trophic level mammals (e.g., mysticetes) available as prey. Mysticetes became more abundant following a diversity plateau during the mid-Miocene39,45. Subsequently mysticetes remains become more prominent in the Late Miocene to Early Pliocene fossil assemblages from North Carolina and Florida studied herein43,44, where mysticetes were, together with other mammals (e.g., odontocetes), possibly preyed upon by O. megalodon and C. carcharias.For the Early Pliocene of North Carolina, where we have δ66Zn values for both Otodus megalodon and Carcharodon carcharias, our results suggest largely overlapping trophic levels for both species. Feeding at the same trophic level does not necessarily imply direct dietary competition, as both species could have specialised on different prey with similar trophic levels. However, at least some overlap in food items between both species is likely, as also indicated by fossil bite marks1,4,34,38. Extant predatory sharks typically feed on a wide range of food items33, and there is evidence for generalist feeding, as well as, in some cases, specialisation at lower trophic levels for extant C. carcharias32. Higher dietary individuality and the opportunistic nature of apex predators are possible explanations for the range of δ66Zn values observed in both species (–0.61 to –0.04‰ in Pliocene North Carolina).The extinction of Otodus megalodon could have been caused by multiple, compounding environmental and ecological factors46,47, including climate change and thermal limitations48, the collapse of prey populations4 and resource competition with Carcharodon carcharias15 and possibly other taxa not examined here (e.g., carnivorous odontocetes). The δ66Zn results presented here indicate the potential of trophic change, where we find evidence for a decrease in the mean trophic position from O. chubutensis to O. megalodon in the Atlantic and an increase in trophic position for C. carcharias from the Early Pliocene to its extant form. If these trophic dynamics are accurate, then there is a possibility for the competition of dietary resources between these two shark lineages15. Our results also support the hypothesis of Otodus size-driven co-evolution and co-extinction with mysticetes4, indicated, at least for the Atlantic assemblage, by a shift towards lower trophic level prey from the Early Miocene to the Early Pliocene within the Otodus lineage. In general, our study demonstrates δ66Zn to be a powerful, promising tool to investigate the trophic ecology, diet, evolution, and extinction of fossil marine vertebrates. More

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Choice of biowaste treatment plants and sample identifiersCompost samples were collected from four central municipal biowaste treatment plants (denominated as #1 to #4) in Baden-Wurttemberg, Germany (Table 1). All plants used a state-of-the-art two-stage biowaste treatment process comprising of (a) anaerobic digestion/biogas production and (b) subsequent composting of the solid digestate to produce a high-quality mature compost sold for direct use as fertilizer in agriculture. The composts were regularly analyzed by an independent laboratory for quality and residual contamination and consistently fulfilled the quality requirements of the label RAL-GZ 251 Gütezeichen Kompost of the German Bundesgütegemeinschaft Kompost e.V. (www.gz-kompost.de). Plants #1 and #3 produce in addition a liquid fertilizer, which is separated from the solid digestate at the end of stage a) by press filtration and which is also intended for direct use on agricultural soil (replacement of liquid manure). In case of plants #1, #3, and #4 up to 25 wt% of shrub/tree cuttings were added to the solid digestate for composting. All plants used sieving (typically with a 12 or a 20 mm mesh) at the end of the process to assure the necessary purity of their finished composts. Whenever technically possible, we as well took samples of the pre-compost immediately before this final sieving step to evaluate its contribution to the removal of residual BPD fragments. For analysis, composts were passed consecutively through two sieves with mesh sizes of 5 mm and 1 mm, yielding two fragment preparations for IR-analysis namely a > 5 mm fraction corresponding to the contamination by residual “macroplastic” (5 mm is a commonly used upper size limit for “microplastic”, anything larger is macroplastic) and a 1–5 mm fraction corresponding to the regulatory relevant residual contamination by microplastic. The lower limit of 1 mm rather than 2 mm was chosen in anticipation of the expected changes in regulation, where the replacement of the 2 mm limit by a 1 mm limit is imminent.Table 1 Technical data of the investigated plants and incidence of BDP fragments in the sampled composts.Full size tableOccurrence of plastic fragments  > 1 mm in the sampled compostsComposting times of 5–9 weeks were used in the investigated plants (Table 1), which is shorter than the 12 weeks indicated in EN 13432 for the 90% disintegration of a compostable plastic material, but a realistic time span for state-of-the-art technical waste treatment. Since we were not in a position to estimate the quantity of BDP entering the plants, since for technical reasons we were unable to obtain a representative sample, we cannot say, whether any residual BDP detected by us in the finished composts was due to a yet incomplete disintegration process or whether it corresponds to the 10% material still permissible by EN 13432 even after the full composting step. However, in 7 out of the 12 sampled composts and pre-composts fragments with chemical signatures corresponding to the BDPs poly (lactic acid) (PLA) and poly (butylene-adipate-co-terephthalate) (PBAT) were identified in the > 5 mm and/or the 1–5 mm sieving fractions using FTIR analysis3 (Fig. 1; Table 1). All recovered fragments appeared to stem from foils, bags or packaging, since they were thin compared to their length and width (see Suppl Figure S1 for typical examples). Fragments with overlapping signatures, most likely PBAT/PLA mixtures or blends, were also found (see Suppl Figure S2 for the interpretation of the spectra). In addition, the recorded BDP fragment spectra (Fig. 1A) showed high similarity to the FTIR spectra of commercial compostable bags sold in the vicinity of the biowaste treatment plants (Fig. 1B), which together with the geometry of the recovered fragments led us to assuming that the majority of the BDP entered the biowaste in the form of such bags.Figure 1FTIR spectra of BDP fragments from composts and commercial bags. (A) BDP fragments recovered from the composts and (B) the commercial compostable bags. Fragments were coded as follows: p or f for pre-compost or finished compost, followed by the plant number (#1 to #4), an indication of the size fraction ( > 5 mm or 1–5 mm) in which the fragment was found, and finally, the fragment number. Fragment F#1_5mm_4 therefore represents the 4th fragment collected in the  > 5 mm size fraction from the finished compost of plant number 1. Bags were arbitrarily numbered 1–10, see Suppl Table S1 for supplier information. The spectra (in grey) of the reference materials for PLA and PBAT are given as basis for the interpretation. Spectra in red refer to test samples consisting only of PBAT, while those in blue indicate samples composed of PBAT/PLA mixtures.Full size imageThe BDP fragments were found alongside fragments of commodity plastics (mostly PE) in all cases. Finished composts tended to contain fewer and smaller fragments than the corresponding pre-composts. The final sieving of the pre-composts to prepare the finished composts hence appears to be quite effective in removing such fragments, in particular those from the > 5 mm size fraction (Table 1) and for that reason has become state-of-the-art in preparing quality composts (contamination by plastic fragments > 2 mm of less than 0.1 wt%). Given that the size of the fragments is a crucial factor regarding ecological risk, we analyzed the sizes (length Î width) of the BDP fragments in comparison to that of the plastic fragments with signatures of commodity plastics such as PE (Fig. 2). BDP fragments found in a given compost sample tended to be smaller than the fragments stemming from non-BDP materials, which may indicate that BDPs degrade faster or tend to disintegrate into tinier particles than commodity plastics. This may also explain why in the compost from plant #2, no BDP fragments were found in the particle fraction retained by the 5 mm sieve ( > 5 mm fraction), while 19 such particles were found in the fraction then retained by the 1 mm sieve (1–5 mm fraction). Interestingly, plant #2 is the only one included in our study that uses no mechanical breakdown of the incoming biowaste. This reduces the mechanical stress on the incoming material. Mechanical stress can alter the properties of plastic foils such as the crystallinity whereby crystallinity has been shown to influence the biological degradation of BDP such as PLA7.Figure 2Size distribution of plastic fragments  > 1 mm. (A) Fragments found in the finished compost from plant #1, (B) in the finished compost from plant #2, and (C) in the pre-compost from plant #3. For reasons of statistical relevance, only samples containing more than 20 BDP fragments per kg of compost were included in the analysis.Full size imageMaterial characteristics of BDP fragments in comparison to those of commercial biodegradable bagsIn order to verify whether the BDP fragments recovered from the composts differed from the compostable bags in any parameter with possible relevance for biodegradation and environmental impact16, the physico-chemical properties of bags and fragments were studied in detail. Since we wanted to have a maximum of information of the BDP fragments, size/weight was a limiting factor in selecting fragments for analysis. Fragments of at least 1 mg were required for the FT-IR analysis. 5 mg-fragments could be analyzed in addition by 1H-NMR, while the full set of analytics (FT-IR, 1H-NMR, and DSC) required at least 10 mg of sample.For insight into the chemical composition, 1H-NMR spectra of the commercial bags and all suitable BDP fragments were compared (Fig. 3). In case of material mixtures and blends, the 1H-NMR analysis allows quantification of the PBAT/PLA weight ratio in the materials and also of the ratio of the butylene terephthalate (BT) and butylene adipate (BA) units in the involved PBAT polyesters.Figure 31H NMR spectra of BDP fragments from composts and commercial bags. (A) BDP fragments recovered from the composts and (B) the commercial compostable bags. Fragments were coded as follows: p or f for pre-compost or finished compost, followed by the plant number (#1 to #4), an indication of the size fraction ( > 5 mm or 1–5 mm) in which the fragment was found, and finally, the fragment number. Bags were arbitrarily numbered 1–10, see Suppl Table S1 for supplier information. The spectra (in grey) of the reference materials for PLA and PBAT are given as basis for the interpretation. Spectra in red refer to test samples consisting only of PBAT, while those in blue indicate samples composed of PBAT/PLA mixtures. (C) Chemical structures of PLA and PBAT, chemical shifts of the protons are assigned as indicated in the reference spectra in (B).Full size imageThe 1H-NMR spectra corroborate the FTIR measurements in that all investigated commercial bags were made from PBAT/PLA mixtures of varied composition (Table 2). By comparison, some of the fragments, for instance, f#1_5mm_4, appeared to consist of only PBAT. Other fragments, e.g., f#1_1mm_9, were mixtures of PLA and PBAT (Table 2). However, even in the case of PBAT/PLA mixtures, the average PBAT content tended to be higher in the fragments than in the bags, while the BT/BA monomer ratio in the respective PBATs, was also significantly higher in the fragments than in the bags. If we assume the fragments to stem from similar compostable bags as the ones included in our comparison, this would mean that during composting of such a bag, the PLA degrades more quickly than the PBAT, whereas within a given PBAT polyester, the BA unit is more easily degraded than the BT unit. Evidence can indeed be found in the pertinent literature that PLA has faster biodegradation kinetics than PBAT, while BT is more resistant to mineralization than BA17,18.Table 2 Composition of commercial compostable bags and BDP fragments recovered from the composts as analyzed by 1H-NMR.Full size tableNext, differential scanning calorimetry (DSC) was used to analyze fragments compared to commercial bags in regard to the presence of amorphous vs. crystalline domains, a parameter expected to affect biodegradation kinetics and therefore the putative environmental impact of the produced microplastic16 upon release into the environment with the composts. Whereas amorphous domains show glass transition, crystalline domains show melting, both of which can be discerned by the respective phase transition enthalpy in the DSC curves (Fig. 4).Figure 4DSC curves of BDP fragments and compostable bags #1 and #7. Curves for the reference materials (in grey) for PLA and PBAT are given for comparison. Curves were recorded during the first heating run (temperature range: − 50 °C to 200 °C, heating rate: 10 °C min−1). (A) and (B) curves in red refer to test samples consisting only of PBAT, while those in blue indicate samples composed of PBAT/PLA mixtures. Fragments were coded as follows: p or f for pre-compost or finished compost, followed by the plant number (#1 to #4), an indication of the size fraction ( > 5 mm or 1–5 mm) in which the fragment was found, and finally, the fragment number.Full size imageThe curve for the reference PBAT shows a glass transition temperature (Tg) of − 29 °C and a broad melting range between 100 and 140 °C for the crystalline domains, while that of the PLA reference shows a glass transition temperature of 58 °C and a narrower melting peak between 144 °C and 162 °C. The curve for commercial bag #1, which had a comparatively high PLA content, shows a pronounced melting peak in the expected range; the same is the case for fragment p#3_5mm_1 and to a lesser extent for fragment p#3_5mm_9, two fragments, which also have high PLA contents. The DSC curves of the other fragments and bag #1 are undefined in comparison, which is due to their high PBAT content. According to the DSC curves, most of the investigated materials are semicrystalline, i.e., contain both amorphous (glass transition) and crystalline (melting) domains. However, the DCS data alone allow only a qualitative discussion of the differences between fragments and bags.To obtain quantitative data on the crystallinity differences, wide angle X-ray scattering (WAXS) spectra were recorded. WAXS requires fragments at least 3 cm long, which restricted the number of fragment samples to three, all of which were found in pre-compost samples. The corresponding curves are shown in Fig. 5A–C. The spectra of the commercial biodegradable bags are shown in Suppl Figure S3. Foils were in addition prepared by heat pressing from the reference materials for PLA and PBAT in order to include them into the WAXS measurements (Fig. 5D). While the foils produced from the PBAT reference material produced crystallinity peaks at 16.2°, 17.3°, 20.4°, 23.2°, and 24.8°, the foil prepared from the PLA reference material showed only an amorphous halo at 15.5° and 31.5°, which is in accordance with values published in the literature19. A more pronounced crystallinity peak was obtained in the case of an additionally annealed PLA foil.Figure 5WAXS curves with Lorenz fitting for (A) fragment p#3_5mm_1, (B) fragment p#3_5mm_9, and (C) fragment p#4_5mm_2. (D) WAXS curves for foils produced from the PBAT and PLA reference materials; the percent values indicate the crystallinity. The dash lines are the fitting peak curves for the XRD spectrum. Crystallinity can be obtained by dividing the integration area of the fitted peaks by the integration area of the entire spectrum. Fragments were coded as follows: p or f for pre-compost or finished compost, followed by the plant number (#1 to #4), an indication of the size fraction ( > 5 mm or 1–5 mm) in which the fragment was found, and finally, the fragment number.Full size imageIn case of the fragments and bags, the peaks of PLA and PBAT overlapped to some extent in the WAXS spectra, but by conducting Lorenz fitting using Origin software, the overall crystallinity could be calculated as follows:$$chi = { 1}00% , *{text{ Aa}}/left( {{text{Aa }} + {text{ Ac}}} right)$$where χ is the crystallinity and Aa and Ac represent the areas of the amorphous and crystalline peaks.Using this equation, crystallinities of 55% (fragments p#3_5mm_1), 34% (p#3_5mm_9), and 34% (p#4_5mm_2) were calculated for the fragments. The foils prepared in house for the reference materials had similar crystallinities (43% in case of the annealed PLA foil and 26% of the PBAT foil), while the simple PLA foil was amorphous. By comparison, for eight of the commercial bags, crystallinities in the range from 1% to 7% were calculated, whereas these values were 14% and 15% for the remaining two bag types (Suppl Figure S3).The high crystallinity of the larger fragments recovered from the pre-compost samples suggests that crystalline domains of BDP materials may indeed disintegrate more slowly than the amorphous ones, as prior studies on microbial biodegradation have suggested7,8. Admittedly, such large fragments per se would not enter the environment, since the final sieving step used to prepare the finished composts is quite efficient at removing them. However, it is tempting to extrapolate that residual BDP in general are remnants of the more crystal domains of the original material, even though experimental proof of this assumption is at present not possible. 10 wt% of a BDP bag is allowed to remain after standard composting. It is usually assumed that any such residues continue to degrade with comparable speed. However, should these residues correspond to the more crystalline domains, rather than degrading with similar speed as the bulk material, the more crystalline fragments can be expected to persist for a much longer and at present unpredictable length of time in the environment, e.g. when applied to the soil with the composts; in particular, when they are also enriched in PBAT and BT units as suggested by our analysis of the chemical composition. Data from the use of biodegradable foils in agriculture show that the degradation in the environment may take years20. Altogether this may have unforeseen economic and environmental consequences, especially when considering the high fraction of BDP fragments < 5 mm. Putative consequences include changes in soil properties, the soil microbiome and therefore in plant performance21, a factor indispensable for worldwide nutrition.Residues of BDP fragments  1 mm were found in the collected LF samples. This is hardly surprising, given that the LF is produced by press filtration of the digestate after the anaerobic stage. Such a filtration step can be expected to retain fragments > 1 mm in the produced filter cake, which goes into the composting step, leaving the filtrate, i.e. the LF, essentially free of such particles. Anaerobic digestion is currently not assumed to contribute significantly to the degradation of BDP17,22, but the process conditions (mixing, pumping) may promote breakdown of larger fragments, particularly when additives such as plasticizers23 leach out of the material.Since the residual solids content of the LF is low (plant #1: 8.6 wt%, plant #3: 5.8 wt%), a combination of enzymatic-oxidative treatment and µFTIR imaging originally developed for environmental samples from aqueous systems24,25 could be adapted for the analysis (size and chemical signature) of particles in the LF down to a size of 10 µm. The corresponding data are compiled in Table 3. In all cases, residual fragments from PBAT-based polymers represented the dominant plastic fraction in the investigated samples; i.e. approximately 53% of all plastic particles in the LF from plant #1 (11,520 BDP particles per liter) and 65% in the case of plant #3 (12,480 BDP particles per liter). Liquid manure is applied several times a year to fields at a concentration of 2–3 L m−2. According to our analysis > 20,000 BDP microparticles of a size ranging from 10 µm to 500 µm enter each m2 of agricultural soil whenever LF is applied on agricultural surfaces.Table 3 Microplastic fragments (BDP/all) found per liter of liquid fertilizer.Full size tableDue to the complexity of the matrix, a similar analysis of individual plastic fragments  1 mm. Six compost samples representing the more contaminated ones based on the content of fragments > 1 mm, namely, f#1, f#2, p#3, f#3, p#4 and f#4 (nomenclature: f or p for finished or pre-compost, followed by plant number), were extracted with a 90/10 vol% chloroform/methanol mixture. The amounts of PBAT and PLA in the obtained extracts were then quantified via 1H-NMR (Table 4). Briefly, the intensity of characteristic signals in the extract spectra of the compost samples (see Suppl Figure S4) were compared to peak intensities produced by calibration standards of the pure polymer dissolved at a known concentration in the chloroform/methanol. All samples and standards were normalized using the 1,2-dichloroethan signal at 3.73 ppm as internal standard. See also Suppl Figure S5 for an exemplification of the quantification of the PBAT/PLA ratios. Based on the amounts of PBAT and PLA extracted from a known amount of compost, the total mass concentration (wt% dry weight) of these polymers in the composts was calculated.Table 4 Evidence of PBAT and PLA residues caused by fragments  2 mm. Moreover, residues of PBAT and PLA were found in all investigated compost samples, including the finished compost from plant #4, which had shown no contamination by larger BPD fragments (Table 1). The pre-compost from that plant had shown a few contaminating BDP fragments in the > 5 mm fraction. However, in regard to the fragments More