Although eco-acoustic monitoring has the potential to deliver biodiversity insight on vast scales, existing analytical approaches behave unpredictably across studies. We collated 8,023 audio recordings with paired manual avifaunal point counts to investigate whether soundscapes could be used to monitor biodiversity across diverse ecosystems. We found that neither univariate indices nor machine learning models were predictive of species richness across datasets but soundscape change was consistently indicative of community change. Our findings indicate that there are no common features of biodiverse soundscapes and that soundscape monitoring should be used cautiously and in conjunction with more reliable in-person ecological surveys.
Anthropogenic pressures are impacting biodiversity globally 1 . Declines in species richness catch headlines 2 but changes in community composition can have just as devastating ecological effects 3 . To design evidence-based conservation measures, monitoring biodiversity is essential and listening to the sounds produced by an ecosystem (acoustic monitoring) holds promise as a scalable and inexpensive way to achieving this 4 .
Many species contribute to an ecosystem’s soundscape, whether through producing vocalizations (for example, birdsong), stridulations (for example, cricket chirps) or as they interact with the environment (for example, the buzz of a bee). Streaming 5 or recording 6 soundscapes across huge scales is now common 7,8,9 , yet interpreting the audio to derive biodiversity insight remains a challenge. Automating the identification of stereotyped sounds in audio recordings 10 can provide species occurrence data on large scales, building a bottom-up picture of biodiversity. However, vocalization detection algorithms rely on vast amounts of training data 11 , meaning even state-of-the-art models are only able to reliably detect vocalizations from the most commonly found species 12 . An alternative top-down approach is to use the features of an entire soundscape to infer biodiversity 13 . Entropy-based acoustic indices 13 or embeddings from machine learning models 14 have both been used to predict community richness or ecosystem intactness with some success when calibrated using trusted independent sources of ground-truth biodiversity data (hereafter, ground-truth data) 15,16 . Nonetheless, soundscape features which correlate positively with biodiversity at one site can have an inverse relationship in another 17,18 and no reliable single metric has been found 19 . The inability of both vocalization detection and soundscape approaches to generalize has meant that acoustic monitoring has only provided ecological insight in already well-studied regions and the technology’s transformative potential has remained largely unfulfilled.
We collated 8,023 short (1–20 min) soundscape recordings collected concurrently with manually recorded avifaunal community data from four diverse datasets: a temperate forest in Ithaca, USA (n = 6,734, one site), a varied tropical rainforest landscape in Sabah, Malaysia (n = 977, 14 sites), an agricultural tea landscape in Chiayi, Taiwan (n = 165, 16 sites) and a varied montane tropical forest and grassland landscape in the Western Ghats, India (n = 147, 91 sites). In India, Malaysia and Taiwan, avian community data were collected by in-person point counts performed by experts, whereas citizen science checklists from eBird were used for the USA 20 (Methods). For each recording, we calculated two common types of acoustic features: (1) a 128-dimensional convolutional neural network (CNN) embedding (learned features, LFs) 14,21 and (2) 60 analytically derived soundscape indices (SSIs) 22 .
By training a machine learning model on the full-dimensional learned feature vectors, we were able to predict species richness within datasets with relative success (coefficient of determination R 2 = 0.32, 0.50, 0.22 and 0.14 for India, Malaysia, Taiwan and the USA, respectively; Fig. 1d). However, except for in one case (trained on Taiwan, tested on India, R 2 = 0.13), the models were unable to predict richness when evaluated on datasets that they were not trained upon (R 2 < 0, Fig. 1d). There was no significant correlation between dataset sample sizes and within-dataset or mean cross-dataset R 2 values (Pearson’s correlation P > 0.05). Similar results were found when using only the 60 SSIs or all 188 features (128 LFs + 60 SSIs) together to predict richness (Extended Data Fig. 1). Our results indicate that even within a single dataset—but especially when looking across datasets—soundscapes with similar levels of avian diversity do not share similar acoustic features.
Rather than attempting to predict species richness directly, we investigated the relationship between pairwise Euclidean distances between the mean acoustic features of each audio recording (soundscape change) and pairwise Jaccard distances between their associated avian species communities (community change). We found strong significant correlations between soundscape change and community change within each dataset (Spearman correlation Mantel test, P ≤ 0.001 for all; Fig. 1e). There were more examples where communities changed but soundscapes did not than the converse, indicated by clustering of points in the upper left of Fig. 1e. No significant correlations were found between soundscape change and change in species richness in any of the four datasets (Extended Data Fig. 2). For each of the 14 sites in Malaysia—the only dataset with sufficient replicates at individual sites to test for this relationship—there was a significant correlation between soundscape change and community change (P ≤ 0.03). The decision to use LF (Fig. 1e) or SSIs (Extended Data Fig. 3) to track soundscape change did not alter our findings.
In all four datasets, we found that certain acoustic features correlated with avian richness. However, features did not behave consistently across datasets and most were only correlated with richness in one or two datasets (explaining inconsistencies seen in the literature 19 ; Fig. 1a–c). Without access to independent sources of trusted ground-truth biodiversity data, ascertaining which features were most suitable for each dataset would be impossible, limiting the transferability of this approach. We then found that a machine learning model trained on compound indices was able to produce within-dataset predictions of species richness with some success. However, again, generalizability was not achieved as models did not produce informative estimates when applied to datasets they were not trained upon (Fig. 1d). The diversity of flora, fauna, survey designs and recording equipment across the datasets might, perhaps, make these results unsurprising. Indeed, further studies may show that limited generalizability can be achieved when study biomes and survey methodologies are closely matched. Nevertheless, given the current state-of-the-art, our results stress that however well an acoustic feature or machine learning model may perform in one scenario, without access to high-quality ground-truth data from the exact location being studied, soundscape methods should not be used to generate predictions of species richness.
In contrast to the unpredictable behaviour of the models producing estimates of avian richness, we found that soundscape change correlated with community change in all datasets (Fig. 1e). However, this approach was also limited, as there were many examples where similar soundscapes were associated with very different avian communities. One reason could be that non-biotic sounds (for example, motors) or non-avian biotic sounds (for example, insects and amphibians) were larger contributors to soundscapes than were birds 15,23 . Our ability to predict avian richness was therefore more likely to have been based on latent variables measuring habitat suitability (for example, vocalizing prey or nearby water sources) rather than being driven by the vocal contributions of birds directly.
Despite having access to vast amounts of manually collected avifaunal point count data, even within datasets, soundscape predictions of avian richness and community change struggled with accuracy. This result suggests that even if we can collate standardized global soundscape datasets and train models using ground-truth data from every biome on earth, we may still be left wanting. Rather than relying on automated predictions in isolation, a more reliable approach might be to use soundscapes to collect coarse but scalable data, which can be used to direct more detailed ecological studies. For example, data from large-scale acoustic monitoring networks could be used to direct in-person surveys towards only the sites with unexpected and prolonged soundscape changes or large shifts in predicted biodiversity. Such an approach could result in more efficient use of limited expert resources to ensure that conservation interventions are deployed in a timely and focussed manner. Indeed, due to issues surrounding interpretability and accountability, we will probably always require expert verification of autonomous monitoring outputs before policy and management practices are modified 24 . Proceeding with more realistic expectations around how soundscapes can best contribute large-scale biodiversity monitoring efforts will be essential to maximizing the transformative potential of the technology.
Data were collected from 91 sites in the Western Ghats between March 2020 and May 2021. Sites were a mixture of montane wet evergreen and semi-evergreen forests (29), montane grasslands (8), moist deciduous forests (7), timber plantations (17), tea plantations (20), agricultural land (8) and settlements (2). Sites had a mean separation distance of 18.9 km, with the closest two being 824 m apart.
In total, 147 15-minute point counts were conducted (mean 1.6 per site). All point counts were conducted between 06:00 and 10:00 as these were the hours with highest expected avian activity. A variable-distance point count approach was followed and all bird species heard, seen and those that flew over (primarily raptor species) were noted. A total of 119 avifaunal species were recorded across all point counts.
Single channel audio was recorded during each point count using an Audiomoth device raised 1–2 m from the ground 6 . Recordings were saved in WAV format at a sampling rate of 48 kHz.
Data were collected from a varied tropical landscape at the Stability of Altered Forest Ecosystems (SAFE) project 25 in Sabah between March 2018 and February 2020. The 14 sites spanned a degradation gradient: two in protected old growth forest, two in a protected riparian reserve, six in selectively logged forest (logging events in 1970s and early 2000s), two in salvage logged forest (last logged in early 2010s) and two in oil palm plantations. Sites had a mean separation distance of 7.6 km, with the closest two being 583 m apart.
In total, 977 20-minute avifaunal point counts were performed across 24 hours of the day (59–80 per site). During point counts, all visual or aural encounters of avifaunal species within a 10 m radius of the sampling site were recorded. Species identifications and names were validated using the Global Biodiversity Information Facility 26 . A total of 216 avifaunal species were recorded across all point counts.
Single channel audio was recorded during each point count with a Tascam DR-05 recorder mounted to a tripod raised 1–2 m from the ground (integrated omnidirectional microphone, nominal input level −20 dBV, range 20 Hz–22 kHz). Recordings were saved in WAV format at a sampling rate of 44.1 kHz.
Data were collected from 16 tea plantations in the Alishan tea district, located in Chiayi County of Taiwan, between January and November 2022. The tea plantations spanned an elevation gradient from 816 to 1,464 m and were surrounded by secondary broadleaf forests or coniferous plantations.
In total, 176 10-minute avifaunal point counts were conducted, on average once per site per month. Every survey was conducted within 3 hours after sunrise. During point counts, the species of every bird individual visually or aurally detected was recorded and its horizontal distance from the observer was estimated as 0–25, 25–50, 50–100, >100 m or flying over. Our process for matching audio recordings to point counts is provided later in the Methods. A total of 81 avifaunal species were recorded across all point counts used in this study.
Stereo audio was recorded at a sampling rate of 44.1 kHz for one of every 15 minutes at each site throughout the sampling period. Wildlife Acoustics Song Meter 4 devices were used (mounted to a tree trunk or tripod raised 1–2 m from the ground) and data were saved in WAV format.
Data were collected from a single site in a temperate forest at Sapsucker Woods, Ithaca, USA, from January 2016 to December 2021.
In the absence of standardized point counts across such a long duration, we used eBird checklists to determine avian communities (only possible since Sapsucker Woods is a major hotspot for eBird data). Data were filtered to only keep checklists which were complete, of the ‘travelling’ or ‘stationary’ types, located within 200 m of the recording location and lasted less than 30 minutes (to minimize bias from varying sampling intensities). We combined occurrence data from all eBird checklists that started during each audio recording to create one pseudo point count per audio file. Since we only use occurrence data, double counting of the same species was not an issue. This resulted in a total of 6,734 point counts and 231 species.
Audio was recorded continuously throughout the sampling period using an omnidirectional Gras-41AC precision microphone raised approximately 1.5 m from the ground and digitized with a Barix Instreamer ADC at a sampling rate of 48 kHz. Files were saved in 15-minute chunks and in WAV format.
For each point count, we used the field observations to derive community occurrence data (that is, presence or absence) since not all datasets had reliable abundance information. To calculate avian richness, we calculated the unique number of species encountered within each point count (α-diversity).
We used two approaches to calculate acoustic features for each recording. The first was using a suite of 60 SSIs, using the scikit-maad library (v.1.3) 22 . Specifically, we appended together the feature vectors from the functions maad.features.all_temporal_alpha_indices and maad.features.all_spectral_alpha_indices. In all cases, we used default arguments for generating features, since tuning parameters to each of the datasets would have been a cumbersome and inherently qualitative process susceptible to introducing bias. A full list of features is given in Supplementary Table 1.
The second approach was to use a learned feature embedding LF from the VGGish CNN model 21 . VGGish is a pretrained, general-purpose sound classification model that was trained on the AudioSet database 27 . First, to transform raw audio into the input format expected by the pretrained VGGish model 21 , audio was downsampled to 16 kHz, then converted into a log-scaled mel-frequency spectrogram (window size 25 ms, window hop 10 ms, periodic Hann window). Frequencies were mapped to 64 mel-frequency bins between 125 and 7,500 Hz and magnitudes were offset before taking the logarithm. Data were inputted to the CNN in spectrograms of dimension 96 × 64, producing one 128-dimensional acoustic feature vector for each 0.96 s of audio.
For all acoustic features (LFs and SSIs), mean feature vectors were calculated for each point count. For Malaysia, India and the USA, recordings had a 1:1 mapping with point counts, so features were averaged across each audio recording. For Taiwan, where audio data were sparser, features were averaged across all audio recordings that begun within a 1-hour window centred around the start time of the point count, resulting in an average of four 1-minute audio recordings per point count. Using window sizes of 40 or 20 minutes around the point count start time in Taiwan did not change our results (Fig. 1d reproduced with different window sizes and Extended Data Fig. 4).
Univariate Pearson correlations were calculated between each of the 60 SSIs and 128 LFs with avian richness across all point counts within datasets. Significance thresholds were determined by calculating 100 null correlation coefficients between shuffled features and avian richness. Null correlation coefficients were aggregated for each dataset across all 60 SSIs and 128 LFs (total null values 18,800). Accounting for a Bonferroni multiple hypothesis correction, the threshold for significance at P = 0.05 was taken as value 18,784 in an array of sorted absolute null correlation coefficients in ascending order. For each dataset, all real correlations between acoustic features and avian richness with an absolute correlation coefficient greater than this threshold were determined to be significant. Lines of best fit in Fig. 1c were determined by fitting a first-order polynomial to the data on avian richness (x axis) and the single acoustic feature being considered (y axis).
We used a random forest regression model to make predictions of avian richness using the full-dimensional acoustic feature vectors. In Extended Data Fig. 1b, the mean of all 60 SSIs and 128 LFs were combined to create one 188-dimensional feature vector per audio recording. The scikit-learn implementation RandomForestRegressor was used with a fixed random seed and other default parameters unchanged. We randomly selected 70% of point counts in each dataset for training and held back 30% for testing. Splitting train and test data randomly represents a ‘best-case’ scenario since, within datasets, training data distributions will closely match test data distributions. We measured how well predicted richness matched true richness using the coefficient of determination (R 2 ). Both within and across datasets (that is, for all results in Fig. 1d and Extended Data Fig. 1), models were always trained on the training data and scores were evaluated on the test data.
We used Pearson’s correlation coefficient to test whether there was a link between dataset sample sizes and classifier performance. For within-dataset performance, we used the diagonal values of Fig. 1d. For cross-dataset performance, we took the mean of each row in Fig. 1d, excluding the diagonal values.
To measure community change between two point counts we used the Jaccard distance metric, where 0 indicates no change in community and 1 indicates no shared species between communities. To measure soundscape change we calculated the Euclidean distance between the mean acoustic feature vectors for each point count. Distances were computed in the full 128-dimensional space for the LFs (Fig. 1e) and 60-dimensional space for the SSIs (Extended Data Fig. 3). Performing these two operations between all point counts within each dataset, we derived two matrices representing pairwise community change and pairwise soundscape change. Since community change was bounded between 0 and 1, a Pearson correlation was not appropriate. Therefore, the Spearman correlation between these two matrices was measured using a one-tailed Mantel test, with P values derived empirically (using the skbio implementation skbio.stats.distance.mantel).
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